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ok:, phcnorh • i",ine, thi.!:>er.d.roIc) and fOUl fo< goalS (m"",",d . f"nl>
11.1..1 SWmF. Mosl of lhe swine ope.-.'ioo, ;n North Ame";c. are cJoo;e c<>n· finenlen! <>p
maximum ,"eiglll &:nn in ",'cry ,hon 'ime. 1hc pig>'r< rou· tinely ""wurmcd aod 'reated for p .... 'il;o ;nfo>,a,ion •. 1lJe prodIlClS in u>e i",,'ode vano.., pen""'Mn. coum.p/Ios. 00rameetin. amitraz. and tetnochlorvinph
owncr<. lky ore now routinely used f()( i,!IOCI,rod nematode con'rul. in d<.>g" 8.1.6 CAT S
11>< exjX>SUJe of pesticide. in cal, .'n""'l'"'""l1o" l!>al of dogs, ... nlajot" of in,ecticid<.-;
cl."
'0
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,o.
,o. """d.
,i"".. , W~re analy,,,d hy chmrna'ogra ph;'; ,«bniques, I\:"idde, cau",d n% of ,be poi"",ingc"-"",. ",be'euall ",be,,",;.; ,ubsl:.neCS c.used 22%. The .0i"",1< affected were mainly c,". dogs, shtcp. bin!s, . nd bet:; (Alllonioo " ai" 1997), De'pile lhe lack 0( mono .pecifoc SI.,i"":,, il secm. crnain I",,' poiS
'0
I.,g.,
f"",.,..
.rr"""
pcstici
'0
PJ"''I"'"
OIl'
8.2,1 1.\1\'1I.01'EII nOS E A, abou, LlO .m. a ,,''''''er sprayed 54 milk cows "'ilh dichlor"OS aI a """age 45 litllCS g,-...., th,n inlOnded. The prope' OO"gc, which had been u"d fo< some 'ime, had pm"on en_ lire ly ,.fe alld tIad gi"en 8000 fly cont""'. leading ' 0 impl"O'-'«I mil~ prod"",ioo. n.. VI"""Y e.ec,,;'.. dose prot!",od ,",'ore poi"",itij! wilhin • fe'" min",",,- I'ortunatoly • • 1l ,he CO"" eo,.."".] in • fow hour> . • lthoogh some cow. ""re in ",,-ere danKer fmm coo,'ulsi"", fo, some ,ime, It wos fouod ,hat the hod ,"Lon dkhl"""',,, from Ih. ",'ren.@:c",,,.ine,,whieh ..... lhe ...". ,i>:. and eo!or .. lhe one he nad "sed bofore. T he early morning lighl ""d been dim .nd lhe "",Lo, may ha,·c been """py ,,·hen he rrera,ed tbe .pray (KmPI' anJ Onden. 1964). An on,""",u,«I .moonl of H% ",alal.ion a
,e_
"",.le,
,,,,.,«1
8.U TRf.An :D S IlliD USED AS FEED
8.2.2 IMP ROPER C O-,"IPOUNDS OR FORMULATI0 1
PyrelhroKl • • re widely used as in«:oli,.,,! moinly by ~Ju · C",<midali"" in the li,..,r. The
I'"
.....,k.
""",,0<1,
""Hlo;j, ,ft..-
,.-.em.,.....
'0'
.non
A large oo lbreak of poisoning """urrcd a"",ng ani "",I, fed t""led >«rl vain on 1"'0 oollo
""''''"I}"
8.2.4 ACClVEI>TAI. A1'>"D MA U C IOUS PO lSO:>l lNG
Paraqu.u i, a re'lrided use herbi,-ide tha! i. "-"t",,,,,,ly ""i< to companion "nim.l •• nd I;"~ .. ock "'I,,,n illj!<sied. Etc,'on heif.... wrn: mode"''''l)' poi1Or>ed by p. raqu.t'p"'yod un grass along a dilch be.ide which !he heif.... ",".!Ud on Ul
.x!"""'"
~
(04 _
ri<wttJI died aft«.ruing an .,... !ru.ed
_
fairway Oft .... day of OflPIi
pi,.,""
B.l-S IMI'II.OI'O( Sl·OKAGF. TItree bon.eo In a t..lI loo Lnodcd .... lid 011 • m""bi,.., used 10 dl",it... .. poiWJONl Brt,,',",n thttn. 'hojI «lOt. ....-.cd O.M of
cm•.
t,
,..1<_
.'P<"<''''
1.1.'
1 ~IP ROrER
DlSi'OSAL OF pt:Sfl CIO[
CO~TA I :': ERS
A/'iO I'I'.STIC IOt-: I'OISOSEI) o\N I ~ I ALS
oni ...... lite .. ~ oJurin& jot" (I01e .plOd<. Suott • Nnwioot """lied in tIoc poi"""", of ........ and by bepla
(Oo
"">1<, ............
_I,
U.8 LAYI'EII.SONS WITlI .... O TRAI NING USING rESTIClDFS Sl>unly .tier the introd"",;,,,, .nd •• 'U";.... "ploy"",nl of lo..ic in"",,;ride. in agric ultural pn><:.k • • il btt ..... apparrnt """ """idcr.........""'..."» ~ for ~ .... of lbest. poIenl chemiatl •. Train;n, p
"".mpI>o
al:tlefi>l;.
A pa<1ially.mpricd bot; of pIIom< d~ ia • Iiold .....11td in f.,... hrifm dyinc It>d (I01e "'-'i"l typical dooIi ....uc:noe. inltibitioot P"i_inc'fuor 211 heir.,. co
.""ir
.,od
1.2.7 [.AR(;[..S(;A I.r. PESTICIDE USE
'i'"
The ..- """"""" follacy i . .........ion I"'" "'"'-. spay for""""" )""'"- i. con COlI' ...... 10 be "'"" in ....
"lI> .....
'""'" ... ay _ """""". Akboltlh W1e1 . ... '" Sjl«ik for· mul.,,,,,,, and u""l1y ..,.ify d",. i!:tu..,.. of ~ ....., .. in ohemical ingr
rnmpan'"
=
">&i"
dJ..,,,
,mal.
The con.oIidarioo of ......ltwaI pncIi<eo """ tI1IM prod",". until "'" hWt ....,.-,aIilY """ ""'"' ....."'.... IIId ...,(oocublc rq;,iooo of food ....
,,,,»od Iee"n,
""ini.,
n,ba_.
U Sp:cioll
1!tI< :
..,d wid......-
16'
"""y.
U ' CAIlELt:ssNES.... TIle ~ majority ofpe>ticido ",obkl1l> in """"'...., anima!o reiul, from itnorance or miirnamg""",n,-i n shott. <....1......" "" ,lie u",r', p.lt. For c,"mple. pbbelod tonCaintn ...... bttD' f"",",III IOU"" of I/'OUbIe. Spr.iyin, of ... inLIb ..·illl. ~nd :m.. ",..., 10 be lh< ""'"red 0Dr io lh< ,-h of nnolessfteu. This si .... tion "" il$ ~Id in otoildmo', hair " 'idI ponlllion. cases. the use of"," """'" compound ' - bO
h.-'.
.two"""""
'0
•• 8.J SPECIAl. PRO BLEMS ENCOUNTE RED W IT H P[!o,"TICIDES
,~ ..... icoIoCiocaJ ...........,h hh!C<)' of • • """"'" c~rtial ""'" pooIrn<W1<m lesions. and """,I" of tllelOp1"'" hiJhly ",nifie..,.. ,he defini'i.e prOof of eau", Is of,en the idon'i~· c.. ion and qttan'ificaliun 01. """"ifie I"'>licido '" to' it:oIocic.1 Conc.n'B';"'" in .... appropriat" I)ioJollitll or <"'ironmcnlal lamplcs(OduI>C.I999).
8.J.l S \'l;TEMIC
PESf I CIDE.~
The inrroducIion of inj
ic.1 I".",,,,,, •. such os mi ..... ;ni litis« 01 tbo cat,le Jruh HJP
"""'lid$.
"""""""1ori.....
U.I A(.vn : TOXICITIES
In the .... majooity of poi>Ollinas. the ~i"'"m
"""""itr,.
(o.",.ilo,.,
,i.,.
fi,·.
pal"""
..".,....
."""'at
.«ecI•.
..". SYS'.mic orll'""""",ph." e<>mpoU nd. p""."' ",me .p"d.1 problem. if '''''y lire DOt applied .. _wrop
0'0'.-.
""i""".
Sp«1/iully ....,.;11'·. 10 Ihei:r pn:rpenieo Co...~i"".I 01.. 1935~ Systemic orpnophoophal,,1II\I$I be appIioodduriD~ tbc fall_· ......... ...", inInDIII ........"" arc ill their mow micraoory pIIlMI>. If ac>PIied lOO cmy. "'" . . ,.. an: .... if applied ID<> late. ,be Io<:arion oflho parWt< ... ben a"... ked by ,~ in;ccticid< mal' prod .... xli.,.,\ and of,.n f. ..1 """'plica. 'ion\ doe '" ,he rele..., of ,,,,inl from dud I. " ·.... Time uf .p-
,,,IrIe_
,«",,;,..,
,r.e...
rlie. ,iOtl obvioo
An i"..."woo ODd poo
fWI'I)W
270
CHAP1l'R S
"'''icK\< U.. in V
H.3.3 I)[LAY[O NEUROTOXICITY
M,., ","'h"in~ ,he 3Cul< "sns of org.nophosp!lal. po;'on;n!. • fu"he, di>lillCl manif.""ion of <x[>O
.11"
'0
ha.
8.3.4 BJO LOGICAL ~10RAG[, EXCIU:1'10." . ANI) RESIlIUI'.s The "lIim,,,, challenge in u"ng P"'llicides in do."..,ic ani· m.l, i. to "void !he occurrellCe of ct..mic.l ""idu., in .nimal Jlloo",," intended for human con,umption. To !hi, end. .00 ncrrlion are studied in dome"ic ani"",l. early in pesliddc
"1<",,,"
"al.. 191~), The",, ""food I>as ~el •• Mud.,. en,ur< thot pesticide. are used in agricultural practice in OOCOId with I.bel J«:<mI_ mend:rtio", ond ,,:moo chc:mic.1 "Wlic.lioo . The occa.ior,.1 mi,u\e of , uoh peSlicid<, is u",ally quickly
.,,,,y,
,iorr,'
8.4 UNIQUE1'\ESS OF PESTICIDE USE IN DOMESTIC ANIMALS 8.4. 1 SI'[C IK~ DfFFER~:NCIi.S !'eSl",id< use in demestic .nim.Io i",,,he> lar&< numberS (O( diffe",nt
h.,,,
,ha,
'1Ia'
. *u o.nmo. E>ornpI< '" """'" u;ft"..... .. ~._y 11'1<1 "").'00I0cY
or.,..,.. A""""".) .. '''"'' «V-> rm.c"",
'" _ " " ' " .. ...-.of """'"
",id _
ol "" di~", 0'' '''
' - " ol dir~ ~< "'"
_ . . - boo", ... "")"." "v"""' "'" Ii<"""""""..,. ,.,..iN< (.. ,"-"'1 i.
......... of " ' . . -
~
U_ "....". .m pIl
ho"..' ..- f" .. M,
c-"""ol o... """"'" ","'.•y
w _, ....,. ""'''' C ;rpvnghted material
·xpre,sed dinically •• v.riation, in ",n
"''11<,.., llIOIIOI!a,!ric anim.ls
(hors.e. ,wine. dog. ,ot) have •1"""",h, ph)'Siologkally ond bioc"'mically ,imil.. '0 of human •• cOHlo. , .... p. ond goo" h.,·~. uniqu. pllJ1 of the d i· g..ti"" t""" It .... rum"n) tb.t serve< • • a f""""nlaJion v.! for
tb.,
c"",."ning ""lIul ..... fong< int" proIoin pItt""""" n.e <",,"er· • ion i, oarried out by TU""'" mic""",,!!""i""" which utili,o!be cellul""".nd foragc nu,ricn" to build microbial amioo ""ids 'hat .....ub"'luondy digo"cd in the ruminant"< mull inle"in< (Oehm<: ond B.1ITctt. 19~6), n.e rumen h., a reducing cnviron · ment witb • pH varying from 3_3 (0 7.3, dependi,,! on the diO! c"",,,med . Ruminant< on a h i~_ h carOO/t)'rlMe '.Iioo, . , in fe.:o!tydrne b.... kdown, Gratinll.,i"",I. h,,'e." almost neu!ral ",men pH bccau"" tbe rum
f"" ...,
8.4..1
~1r.rABO I . I Sl\l
BiOl,,,,,,form.ti,,n or m
",lfaJe ,onjugale>. where .. pig,.re c""", i.lIy "",ive in glu. c"ronid. "" ""tivity .0<1 the abil ity to produce 'OfIj "~a ... of glucuronic ",,;,J. It.,..... "nd dol' tend (0 h,sc compa .. ble m · id.ti,'o """,lI.:mi.m•. loading to good ,ulfale ond glucuronide format;"'n. and..--.: c'p"blc of biotran,forminll pe&licides 1>01m,lIy
0<"""
IU.4 A .... IMALS AS SENTINELS Qf' PI(STfCIDE USE
The "'ide'p,ead u'" of pe,lici ... , in don .. ,!ic .ni"",l .. JXIr1i""Iarly li ... 10
272
CIlAP'J'ER S P,,,,i< i"" U", in Vo«rinary MOOicin<
conclude
""are
'""cc
0""'"
II.4.S AGG IU:SSIO N ANI)
ANTJCII 0 l.ISI<:ST[ RA S I<:S Sign. a,«)Ciated ""ith IOC"'" lo,icil~ from "'iaroop/>.>
,.1
,.,ne
.".uit.
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coo,"'"
8.S TRKATl'tIENT OF PESTICIDE TOXICITlES IN DOMES TIC ANIMAI--S The I",alm!:nl and m.nagemen! of pe
"""O<"'f')'
1\1""''
In>C'.
REFERENCES
r1""""
AJo ' =~ A. A, M",<W)' " .... "'" .. the - ,I<, ..-I """"'" "" .. , 0/ ... "" 101_", C~_ ~ ;;.,,,..,,.,.. ,K;..! J •.1a-
t'''"' ' [,,,,,,", 1'"'_"""
A""', <J. A.. aod S... ~. I. A. I, "ew=< Vrt· ......,. Theo..".. R
. . . . r_. .
c....J "-.Mi...
'P"'"'" ............ ....".
w.."" """'- "
~ ...... _
..... )"'1 .. """"'''"''. ... 11_ T"""".
, , - , K.. TWbo, S.. Go"" J. P" ..-I....,.. . .... R. O W1~ OoJ-'bJo"'" ",!
"*'
c__ rox"".
R.,
..,. 'J ' -lll . Ko.;. T. lt (,997) . ...,.,. ~,...,.,...... I, "C...-..t n.,_ .. F........ M~"'- (N. I!. " -. «I." ".". " pp. ___ 11 ) . " _. "",,,,,,,,.
....
K""!'I'- f . 1'1.. _
"'*loo. A. p, ,,96-1). A«>I
_ _ _ ;' <. _ _
K_. Il. A .. "*, H""",-, l . (' m) _ ""_~ .. ...,."""" ... ""'....... ,..:..,.q ( 977). A.. J, \w, N... 0.1. 1 ~1l _ 1 ~" ...... ~ ...... N.... H. ('970) . H@poioOoIio& ... """oIH.,m
M. il086). I, -..-.... .. the Th.""""" uf _ ," {L, Ft;\>ttJ,<J. F. ~ ..... C. ~_ .""-~ 1>04 .... ,\w. "- ... J_ .... l . F.h.,;".. 10._, 8 _ , K.. 1kV_,o..,,,,,. B. <J.,." N.. !;!w"",,,- w, C" S<>i<. >;. I .......
M
17. ll-17. _.D.,~m.J..
1""""""'" t)y . _
__ ... _
,~,{ I 'I%'
'!<no..,.
",,0;.,
.
1',"'*'..-. n .'15-l36
, _ _ C. O'l'i8). ""'"'''''''''''''''' ""''''''"'' io
H_ T""""- ,,, 'l'_'''-
"""',. ;W.
ot.t.t.....
""""'n. P. I>. G. 097tli. 0._...,.... ,. ,k" di>pOOOl 01' ... _ ~". ..,. m . Bud. W B. I''nO). l
. ...
150, 1_1472
....r,.
_..m.." .. ......... '.r<""" O/f... koo , ..k , T..-197""'Ni<. U ..."- , ... , "'-,,, l'oo>, .... 0" .... c-......, T, L.1 19O!.\). f_ "'"""" ;,. .. Ik . - ;, - . ""'''. ;;., 11. ,.. 11oricoI. n.2I1'_1&2
,m"
C""""",,, R. w .. M""""". M. ' .. ""'"- A. 10. ___ s~, E.l..1 A """'" 0/ """"",,, ........"""', " ............... _ici
",." 'd. 37, )7t-)'"
l... T-... " .. D ..
10._
R. J.. .... 9 _ S. J. 11_)
~.......,.., .. _ r i I y .. 'be '"""" "'''un>po
'Ok
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0;.,0_ J .. Pm. R. L .. [I.df,-. K. J .. _ J .. "" "
V. lt 0'Il10).
_*",V.N·_,-I·..· _ mu"!) "." ,,,,., ..
, ... ca.. J. A.. ... UN, A",.,. '''"' 100 l W, """""">,, Ill. E. , '''''l. ~ ""'_.. .. . Do',. I,
m-.n
"-="- M. N .. " ' - "
.... ",0, ""' ... A_
A .. M~ . .... .... E'",,- A. ' ""Ol.
"""".con-
_ .. "" ......... of"""""" <""'-'pi ..... 0"'- 1lWc... , . NO-,.. "",,".10. .. ..-le...... R. (1" " T". ,,~.......t ..,. .........., ....... "'" .,,,.....,..., A "'...... roo.L Md;t, C_ , '- 11O-17!, _ "-, '915), _ . , . 0/ ....... , ....... p..>Jo<'" «l\<, ... of """,. k " . - "11'''' ,. ~ "'" _ _ ,,;,Hili t7. 2Z5-111 ....... M. t 1¥1' ). _ " """""l' ,...,.,...,. ""' .. _0..-... .-...i_ _ ("lot ;..,.,.,. 0;..--). , • ..,-"""Oy o(lI<>vy MeuI ... ""' E... "",,· ""","IF. W. Odvu<, "", ~ Port " .... 16'_JOl o.u..-. _ YoB. , la",. K. M .. T-. ~ . K , c..-, K. P.. _ C. R.. MtOomoo. D M" "'" R""""","- R. C. ('"' ' ~ .....,. 01 <..... _ ,_ """;" ., ~ ...,. •. " , "'" .,.,""', .,. of ~ , " dKhOoo '''_"Y'''''''' >cid _ idn. J. ",od. C""'" I.". 1.1, 1Z2!>-I11' . - . I ( J. O%'J). Men...,- ~ .. ,,,",,,,, 'W .... ~""" f .... . ""'" ..,.
SIo>""""" po;.ao;o. '" """"'-l ,,"'"icOO, .. . "" o;ro - . ",,,,, ~;o., .. """"
A,,.,.. 158."" 1. _ ' "
v...ru...,- A"",",,,,, o...r... ....." _., '" "Z-
01 A, ...
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_ . L ... 0"""1. h""""
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.... .. D. <J .. ..", " ' - .... 11. ( 197];
~
.......... ", ..."'J.A ... y,,, ..... """'.
ClOIdo ... "- B" ..... " - . D. H· 1' m). c _ _ po=otioo 0 1 _ " fuo,J ... r>o;" fWJ S-4. 1,~ . Cloo-t. D. E., S......,., H. I!., .... fin'. f , M, 0 ""). ChIo
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",,...., v..
I y~
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"'<,G. W...... Ooooush. H. w.. ab. o m, "1'",,,, ~;"u..< AN_
....,."A"""""" Pr<>!.!.mloo
FoooJ "," .. ~ "''''''''- U . I .. K"" ....
"""""""-
~ , . .,....,.. R. I .. M"",,", " . R......
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""""'''r ",",,,......
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"'-'.... <_"""'.....
.... E ........... ,.- {M. V.• """'. ....). pp. 07--63. TloJ""
A"'mal ""_
"d_,,<, F W. I ,,,"'c). R=no od'.......... "'~ 0 / _ I. ~ C1ioKaI 'Th.-.g." \\01. , ID
-
pp.
0.""'" f. ....
od. C""""",,, Vd. ""i.. U,,","", 01 ,,.-,........
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CHAPTER
9 Pesticide Use Practices in Integrated Pest Management Frank G. Zalom University of California, Davis
Integrated pest management (IPM) has become broadly accepted as an approach to effectively manage insect, disease, nematode, weed, and vertebrate pests in many parts of the world. However, the complexity of IPM is not always appreciated, and its definition can be distorted to reflect the agendas of organizations and individuals that have come to embrace it. IPM is the management of pests in a systems framework, rather than a tactic or group of tactics for a specific pest or pest group. Many IPM tactics, although they may reduce chemical use, remain chemically intensive. IPM has been described as a continuum, with IPM systems ranging from those that are chemically intensive to those that embrace measures that prevent or avoid pest problems and utilize more biologically based tactics. A minimum level of IPM requires the use of scouting and decisions based upon established action thresholds. Moving IPM along the continuum toward the use of more biologically based methods of managing pests remains a challenge that requires extensive interaction between agricultural scientists, consultants, growers, and regulators to ensure relevant development and effective implementation of these increasingly more complex IPM systems. Crop consultants can play an important role in this process, but their role can become even more significant if the use of a broader array of pest controls is linked to their certification. A wide array of pests including insects, mites, weeds, nematodes, disease-causing organisms, and vertebrates lower the quality and the yield of agricultural products, affect the health of humans and other animals, invade structures and landscapes, and adversely affect natural ecosystems. Managing pests has always been a challenge. Before the introduction of synthetic organic pesticides in the 1940s, which allowed reduction of pest abundance and pest damage to levels that were not previously possible, farmers and others responsible for pest control typically employed multiple tactics, such as sanitation, crop rotation, crop diversity, bait trapping, and mechanical pest or host removal, which were applied preventatively based on knowledge of pest biology. Weeds were removed by hand hoeing and tillage; chemical herbicides were seldomly applied. They Handbook of Pesticide Toxicology
Volume 1. Principles
also used inorganic materials, such as copper, lead, antimony, and arsenic, or botanical compounds, such as nicotine and pyrethrum, which were available at the time. These materials were toxic and expensive to produce in quantity; therefore, availability was limited. Equipment for their application was relatively unsophisticated or lacking. Overall, pesticide use was low relative to contemporary levels. The chemical control paradigm was developed effectively by industry, government, and university researchers, and became widely implemented. Along with modem plant breeding, fertilization, and irrigation methods, the introduction of synthetic pesticides reduced on-farm labor requirements, facilitating the transition of agricultural production in developed countries to a highly mechanized system with relatively more concentrated production that is characterized by increased yields and reduced variability in production. Arguably, this transition has been beneficial in that fewer people must work on farms to produce the food and fiber products required to sustain an ever-growing population. The cost of food and fiber remains low as a proportion of income, and food supplies are relatively stable in developed countries. Unfortunately, in spite of an extensive regulatory system for registration, the increased use of pesticides has been accompanied by unintended social and environmental consequences. These consequences include documented cases of pest resistance and pesticide-induced pest outbreaks, environmental contamination, worker exposure, and public concern for residues on food. The only way to totally eliminate the risk of using pesticides is to prohibit their use, but at what cost? Pesticides are legally classified as economic poisons and are defined as substances used to control, prevent, destroy, or mitigate any pest. Pesticides include inorganic products like sulfur and naturally occurring botanical products like pyrethrum, both of which are acceptable for use by organic growers. Pesticides include vegetable and petroleum oils, fertilizers, and certain fatty acid soaps when they are used for pest control. Naturally occurring microbes, such as Bacillus thuringiensis and Trichoderma harzinium, are considered pesticides when they are pro-
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duced commercially and marketed as pest control agents. Many pesticides are very specific in their actions, acting as growth regulators, repellents, pheromones, desiccants, and defoliants. However, the general public seems most concerned with the use of certain synthetic pesticides, particularly those with broader activity and those with which they are unfamiliar. Agricultural uses in particular are not well understood by the public, so questions abound concerning the safety of these products and the need for their use.
9.1 INTEGRATED PEST MANAGEMENT In the 1950s, identification of pest resistance, pesticide-induced pest outbreaks, and the resurgence of pests that had been under controlled some researchers to call for "integrated control," that is, the combination of compatible biological and chemical control tactics. One of the first citations for integrated control in the literature involved walnut production, and it described the need for integration of pesticides so as to preserve parasites of the walnut aphid [Chromaphis juglandicola (Kaltenbach)] (Michelbacher and Bacon, 1952). Other researchers (e.g., Smith and AlIen, 1954; van den Bosch and Stem, 1962) supported the concept of integrated control to reconcile the use of insecticides and biological controls for insect pests. The concept was expanded to include economic thresholds by Stem et al. (1959), who called their approach integrated pest management or IPM. Economic thresholds are the pest densities at which the value of resulting damage exceeds the cost of applying a control. Their description of IPM added the requirements of pest monitoring and risk assessment before justifying the application of therapeutic measures such as pesticides. Environmental contamination by organochlorine insecticides was recognized in the 1960s, following the publication of the book Silent Spring by Carson (1962). Pesticide use became a political issue, and IPM was promoted as an acceptable approach for managing agricultural pests among some scientists and growers who were interested in applying "supervised control" rather than using strictly preventative pesticide treatments, which had become prevalent by that time. However, concerns about the slow rate of IPM adoption by farmers were raised by IPM researchers (e.g., van den Bosch, 1964). Funding for IPM research greatly increased during the 1970s and early 1980s, with increasing efforts to implement IPM practices through extension services, governmental agencies, and community-based programs. As the philosophy of integrated pest management matured, there grew an ever greater appreciation for integrating the management of weeds, pathogens, and nematodes as well as insects in a cropping systems context, recognizing that fundamental differences exist in the biology of these pests and, therefore, in the preventative and therapeutic measures that can be applied for their control. IPM strategies and tactics have gradually been adopted as alternatives to the conventional chemical control paradigm, and the breadth of institutions and organizations promoting IPM as the most effective way to reduce the risks of using pesticides has dramatically increased.
The United Nations Food and Agriculture Organization's (UNFAO) Panel of Experts on Integrated Pest Control (UNFAO, 1967) defined IPM as " ... a pest management system that, in the context of the associated environment and the population dynamics of the pest species, utilizes all suitable techniques and methods in as compatible a manner as possible and maintains the pest populations at levels below those causing economic injury." In the United States, several recent administrations have endorsed IPM. The United States Department of Agriculture (USDA) Council on Environmental Quality (1972) in its publication Integrated Pest Management wrote that IPM is " ... an approach that employs a combination of techniques to control the wide variety of potential pests that may threaten crops. It involves maximum reliance on natural pest population controls, along with a combination of techniques that may contribute to suppression--cultural methods, pest-specific diseases, resistant crop varieties, sterile insects, attractants, augmentation of parasites or predators, or chemical pesticides as needed." In urging IPM adoption in an environmental message, President Carter (1979) said that " ... IPM uses a systems approach to reduce pest damage to tolerable levels through a variety of techniques, including natural predators and parasites, genetically resistant hosts, environmental modifications and, when necessary and appropriate, chemical pesticides. IPM strategies generally rely first upon biological defenses against pests before chemically altering the environment." Attention to IPM in the United States increased again following the Clinton Administration's 1993 pledge to have 75% of cropland acreage under IPM by the year 2000. During the late 1980s and 1990s, several European countries set goals for reducing pesticide use by 50-75%, often suggesting IPM or integrated crop management as the preferred means to achieving the goals. In some countries, these goals have been met with documented use reductions. However, pesticide use reductions documented to date in these countries may be more the result of using products that are applied at reduced rates or by changes in application methodology than of using nonchemical approaches (Matteson, 1995).
9.2 WHAT IS IPM? Forty years after the term "integrated pest management" first appeared in the literature, a single definition has yet to be universally adopted. This is not unexpected because IPM can be as much a philosophy as a science. Because IPM has diverse proponents, the term has been adapted to support a variety of objectives and agendas. This adaptation has tended to permit narrow definitions of IPM to be proposed, in which it is mentioned primarily in terms of tactics, such as chemical controls or biological controls, which have particularly strong advocates. What was largely promoted as an ecologically based view of pest management by a relatively small group of academics and certain agricultural interests in the 1950s and 1960s has become a term for reaching consensus among government
9.3 The IPM Continuum
institutions, many mainstream agricultural and environmental organizations, agrichemical industry leaders, and sustainable agriculture advocates. Allowing diverse groups to reach common ground is indeed a strength of IPM. However, in many ways it has also become one of its major shortcomings. Depending on its interpretation, IPM can be used to justify current pest control practices, even those that are chemical intensive, without emphasizing reduced-risk alternatives or, more importantly, management of the pest species within an ecosystem framework. Cate and Hinkle (1993) stressed the ecological basis of IPM rather than the tactical emphasis of many IPM definitions in their report "Integrated Pest Management: The Path ofa Paradigm," by correctly stating that IPM is about the manner by which communities are managed. Perhaps the phrase itself has resulted in misinterpretation. Kogan (1988) identified "integrated" as the most ambiguous component of the term "integrated pest management." To many people, integrated refers to the use of multiple control tactics integrated into a single pest control strategy (e.g., Metcalf and Luckmann, 1982). This strategy most typically targets only one species of pest or a single class of pest and, in this sense, focuses upon control measures for the target species, prevention of natural enemy disruption and secondary pest outbreaks, and delaying development of pesticide resistance. A broader interpretation refers to management of the complex of pests that attack a crop, considering the combined effects of weeds, plant diseases, insects, and nematodes (e.g., Newsom, 1980). At its highest level, IPM incorporates interactions among pests, the crop, and the environment within the context of a social, political, and economic matrix. Prokopy (1994) likened the increasing levels of IPM complexity, from integration of control methods for a specific pest to its incorporation into a socioeconomic matrix, to the steps of a ladder, where progressing up the steps represents increased level of integration in a systems context. The word "management" as opposed to control also presents an important IPM concept. Flint and van den Bosch (1981) stated that the word "management" implies acceptance of pests as inherent components of an agricultural system. Indeed, some would say that acceptance of pests in an agricultural system is essential to permit their natural enemies to survive in an ecosystem. The IPM approach is to apply controls to suppress pest populations when necessary to reduce damage to an acceptable level, rather than to eradicate the pest.
9.3 THE IPM CONTINUUM Recently, IPM systems have been characterized as falling along a continuum (Sorenson, 1993) ranging from those that are more chemically intensive, where pesticides are applied based on scouting and the use of thresholds, to these that at biologically intensive, where reduced-risk pesticides may be applied, but in which biological controls and biologically based preventative approaches are predominant.
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The USDA formalized the continuum concept in quantifying IPM adoption by creating categories for no IPM use and three additional levels, which represent progressively greater use of biological or cultural practices instead of conventional pesticides (Vandeman et aI., 1994); other groups have accepted and modified this approach (e.g., Benbrook et aI., 1996; Hoppin et at., 1996; Kogan, 1998). When presented as a continuum, the minimum criteria that constitute the use of IPM are field scouting for both pests and natural enemies, and using action thresholds where they exist to make pesticide use decisions. When an action is warranted, those people who employ a minimum level of IPM would apply selective or the "least disruptive" pesticides available. Although IPM emphasizes a systems approach to management, it is impossible to discuss the practice of managing pests without mentioning tactical intervention. 9.3.1 PESTICIDES
Pesticides often represent the first line of defense in situations of pest outbreak or where a specific pest must be eradicated for quarantine or public health purposes. As mentioned previously, pesticides may be used in an IPM system when applied based on scouting and strict consideration of available action thresholds. However, where choices of pesticides exist, those which are least toxic and present the lowest potential for disruption should be selected for use. Wiggles worth (1950) pointed out that it is sometimes " ... through the activities of the entomologists themselves that entomological problems arise." He also stated that "the public loves the hospital, the doctor, and the bottle of physic; while the advances in preventative medicine which have transformed our lives are scarcely noticed. So too it creates a greater impression on the mind to destroy an infestation of insects that can be seen, than by some simple change in practice prevent any infestation from developing." Pest resistance to specific pesticides, and pest outbreaks that result from applications of broad spectrum pesticides can occur and are well documented. Pest resistance to a chemical can develop rapidly, particularly when the life cycle of a pest species is relatively short and the chemical is repeatedly applied. In a pest population, there are always some individuals that will be genetically resistant to a pesticide. Even when a high percentage of the population is killed, those few individuals that possess the resistant traits will survive and reproduce, passing their genes to the succeeding generation. Thus, a pest population develops that can be controlled only by higher chemical dosages. Eventually, the pesticide being applied can become ineffective against the pest. Thus, most pesticides have a finite effective life. Pesticide resistance has been documented in hundreds of species of insects and mites, plant pathogens, weeds, rodents, and nematodes (see Georghiou, 1986). The best way to manage pest resistance is to apply pesticides less frequently. IPM tactics such as scouting or utilizing nonchemical approaches that reduce the need to apply pesti-
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cides may be helpful. When pesticides must be used, alternating classes of pesticides applied to reduce selection pressure on the pest population can delay the development of resistance. Approaches for monitoring susceptibility of pest populations to specific pesticides have been developed for several key arthropod and disease species. Such technology is relatively common for research applications and when applied by the pesticide manufacturer, but commercial implementation of resistance monitoring by consultants is rare. Many plant-feeding insects do not significantly damage agricultural crops because they are kept under natural control by predators and parasites. However, these natural control agents can be inadvertently disrupted by chemical applications that target the bonafide pest species. This situation can result in the emergence of secondary pests, which have been released from natural control. For example, it is widely believed that spider mites have emerged as serious agricultural and forest pests primarily because their predators have been reduced in abundance by chemical sprays for primary pests.
Insect monitoring, which incorporates the use of bait or pheromone traps, is an approach that has become widely used (see Flint and Klonsky, 1989) for monitoring various pest species, providing information on the mobile adult stage. When used in conjunction with phenomenological models, monitoring can be used to predict pest development and, ultimately, to accurately time pesticide applications. Recent advances in technologies for monitoring temperature and leaf wetness have led to commercial implementation of risk assessment models for several key diseases, including late blight of potato (Phytophthora infestans) (e.g., Krause and Massie, 1975; Stevenson, 1983), and grape powdery mildew (Uncinula necator) (Gubler, 1991; SaIl, 1980). Model predictions are usually made by first predicting when conditions are met that are favorable to disease development, and then assessing the severity using a disease risk index. Commercial validation of risk assessment models has shown potential for reducing the number of applications, depending on year, geography, and disease pressure (e.g., Weber et a!., 1996).
9.3.2 FIELD SCOUTING
9.3.3 REDUCED-RISK PESTICIDES
9.3.2.1 Monitoring
When chemical tactics are deemed necessary in an IPM system, the choice of a selective pesticide that kills only the target species is desirable because it is the least disruptive to the crop ecosystem. Examples include some acaricides that are applied for spider mite control that do not affect beneficial predatory mites and microbial controls such as Bacillus thuringiensis, which target particular types of insects. Microbial antagonists have become available for certain pathogens, and selective herbicides are also available. Postemergence herbicides provide the opportunity for growers and consultants to use herbicides in an IPM system by first scouting for weeds to determine their composition and relative abundance before deciding on the control tactic to be employed.
Field scouting or monitoring includes proper identification of pests through surveys or scouting programs, and may incorporate trapping, weather monitoring, and soil testing where appropriate. It may be supported through the use of phenology or risk assessment models, or other types of decision support. In practice, monitoring can be done by either the grower or consultants who check the fields for growers, but there is a labor cost associated with monitoring that is not associated with the preventative use of pesticides. The challenge for researchers is to develop commercial monitoring plans that are economically implementable as opposed to sampling regimes developed for research purposes. Lack of practical monitoring procedures and use of those procedures results in poor timing of applications and an excessive use of pesticides. In many instances (e.g., National Research Council, 1989), pesticide use for controlling a given pest has been reduced 40% without affecting quality or yield simply by using quantitative monitoring procedures in combination with realistic control action thresholds. 9.3.2.2 Decision Support One focus of IPM research for many years has been the development of models that present a framework for integrating information from the various biological disciplines, meteorology, and the field monitoring of pest populations. These models have served to bring disciplines together in analyses of production systems and have yielded tools that can be implemented to support the monitoring or scouting process. IPM research has pioneered many applications for computer technology in agriculture and helped to bring about the early use of electronic instruments for field data gathering (Zalom and Strand, 1990).
9.3.3.1 Behavioral Chemicals Pheromones are highly specific chemicals released by insects to affect the behavior of members of their own species, usually as attractants for mating, but also as signals for aggregation, alarm, or feeding. Synthetically produced pheromones are frequently used in IPM programs as described earlier to monitor adult insect flights. The direct use of pheromones as control agents has also met with some success, usually when the chemical is released over the field from dispensers with the intent of confusing males and preventing mating by inhibiting their ability to locate females. This technique has been applied for control of such key pests as the oriental fruit moth [Grapholitha molesta (Busck)] in Australia and California (e.g., Rice and Kirsch, 1990), the tomato pinworm (Keiferia lycopersicella) in Mexico and the United States (e.g., Jimenez et a!., 1988), the codling moth (Cydia pomonella) (Brunner, 1994), and the pink bollworm (Pectinophora gossypiella) (Flint et a!., 1993). Recently, areawide programs have been established by the USDA
9.3 The rPM Continuum to implement mating disruption for certain key pests on extensive crop acreage in the United States. 9.3.3.2 Conventional Products and Risk Reduced-risk pesticides, as opposed to those that are selective, are usually considered to be safer than traditional pesticides in terms of toxicity to humans and the environment. As regulatory pressures increase the potential for eliminating older classes of pesticides, those pesticides that appear to have a reduced-risk profile are being favored by regulatory agencies as replacement products. Like selective pesticides, those pesticides that have less potential for harming humans or the environment are favored in IPM systems over those that are known to possess such characteristics. One type of risk occasionally ignored when promoting risk reduction in environmental and health terms is the financial risk associated with using less effective controls. Risk is probably the most important financial obstacle to IPM adoption. Growers value pesticides for reducing production risk as well as contributing to profit. For more biologically intensive IPM systems to be voluntarily adopted, it is very important that IPM can be shown to decrease this risk (see Antle and Park, 1986; Gruys, 1982; Way, 1977). In reality, IPM strategies such as monitoring can be tools for managing risk. The more growers learn about pests and their damage potential under an IPM scenario, the less is the uncertainty in their minds about the state of their crop and the more likely it becomes that they will not choose to make a preventative pesticide application. 9.3.4 CULTURAL AND PHYSICAL SUPPRESSION Cultural controls have been used historically to manage many pests, but were often abandoned in favor of pesticides that were less labor intensive. Such controls include a broad range of production practices that render the crop environment less favorable for the pest. Tillage and water management are effective cultural controls in the management of weeds. Furthermore, increased mortality in many insects that overwinter in the soil may result from particular tillage practices. Narrow row plant spacing or optimal in-row spacing can also suppress weeds under certain cropping systems. The destruction of crop residues is important in the management of many pests, such as navel orangeworm in almond, late blight of potato, stem rot of rice, and pink boIl worm and boIl weevil in cotton, for which there are compulsory plow-down dates in several regions. Physical suppression tactics may include cultivation or mowing for weed control, and temperature management or controlled atmospheres for postharvest pests. 9.3.5 PREVENTION Pests are managed in an IPM system in part by preventing their occurrence. Prevention includes those practices that keep
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pests from invading a crop or field and then becoming established. It includes such tactics as using pest-free seeds or transplants, excluding pests by screens or barriers, preventing weeds from reproducing by disking or mowing, and choosing plant cultivars with genetic resistance to insects, nematodes, or diseases, as well as benefits that result from transgenic gene insertion, irrigation scheduling to avoid situations conducive to disease development, cleaning tillage and harvesting equipment when moving between fields, using sanitation procedures to remove an incipient infestation, and eliminating alternate hosts or sites of pest organisms. Even applying fertilizer with the seed of annual crops or through drip irrigation systems can provide a measure of weed control, especially in contrast to broadcast application of fertilizers, which stimulates weed growth. 9.3.6 AVOIDANCE Avoidance is practiced when pest populations exist in a field or site, but the impact of the pest on the crop can be avoided through some cultural method. Examples of avoidance tactics include crop rotation to break the life cycles of pest species, using trap crops, choosing plant cultivars with maturity dates that may allow harvest before pest populations develop or that have a sufficiently short season to permit planting after conditions are conducive to infestation, fertilization programs to promote rapid crop development, and simply not planting certain fields or areas within fields where damaging pest populations are most likely to develop. 9.3.7 PESTICIDES AND BIOLOGICAL CONTROLS Biological control is the augmentation, conservation, and importation of natural enemies, including predators, parasites, and pathogens, to reduce a pest popUlation. This may involve either introduction of a natural enemy or augmentation of one that already exists in the crop ecosystem. Biological control is generally considered to be the cornerstone of any IPM program. Mass culture and release of predatory lacewings, various species of parasitic wasps, and insect pathogens such as Bacillus thuringiensis have been effective in certain insect pest management programs. The release of certain species of plantfeeding insects and pathogenic fungi has been successful in controlling some range land weed pests as well. Biological control has the advantage of generally being safe to nontarget organisms, although there is concern that biological control agents be specific so as not to disrupt native systems. Classical biological control, the release of an imported natural enemy to control a pest species, when successfully established remains more stable in the environment than other pest control tactics. Natural enemies used in augmentative releases generally do not persist and must be rereleased periodically. Conserving natural enemies by avoiding disruptive sprays has become an essential practice in IPM cropping systems.
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Prior to the 1940s, spider mites were considered to be sporadic pests in most perennial crops. Following the introduction and widespread use of broad spectrum pesticides, spider mites became annual pests in many crops. One of the best examples of conserving natural enemies through the careful use of pesticides involves apple production in the Pacific Northwest, where the spider mite Tetranychus mcdanieli (McGregor) can be a primary arthropod pest. Beginning in the mid 1960s, an effective mite management approach based upon the conservation of the western orchard predator mite, Galendromus occidentalis (Nesbitt), using selective insecticides for control of orchard pests was developed and implemented by Hoyt (1969). In this system, organophosphate cover sprays could be used for the codling moth (c. pomonella), the key pest of apples if an alternate prey, the apple rust mite [Aculus schlechtendali (Nalepa)], was encouraged to support populations of G. occidentalis. Implementation of this program reduced the average mite control cost for Washington state growers from $24 per hectare in 1967 to $8-12 per hectare in 1985 (Croft, 1990). As organophosphates and other older classes of insecticides are being replaced by newer materials, there exists a danger of disruption of IPM systems that have successfully integrated the use of chemical and biological control tactics. The increased use of pyrethroids in many cropping systems presents such a danger because they have been shown to be highly disruptive in orchards by killing predator mites (e.g., AliNiazee, 1984; Croft and Hoyt, 1978). Residues of pyrethroids have been shown to be persistent and remain biologically active against predatory mites long past their initial application (Zalom et al., 1998), and there is some indication that the residual effects of pyrethroids persist on orchard trees into the subsequent growing season (Bentley et aI., 1987). The release of natural enemy strains selected for resistance to disruptive insecticides allows the selected natural enemies to persist even when the disruptive materials are applied for control of key pests. A laboratory selected strain of the predator mite G. occidentalis that is resistant to carbaryl and organophosphates has been successfully used to manage spider mites in California almond orchards (Hoy et aI., 1984). Extensive research was conducted on economical mass-rearing (Hoy et aI., 1982), sampling (Wilson et aI., 1984; Zalom et aI., 1984), and applications of selective acaricides at lower than label rates to help adjust the ratio of spider mites to predator mites in favor of the predator mites, thus enabling the integration of this approach with other almond orchard practices.
9.4 PROGRESS ALONG THE CONTINUUM A USDA Economics Research Service study (Vandeman et aI., 1994) estimated that as many as 52% of U.S. vegetable growers were using some level of IPM to manage insect pests; fewer used any IPM practices for plant diseases (~42%) or weeds (~34%). The same study found that 50% of fruit and
nut acreage was under some level of IPM. By definition in this study, a minimum level of IPM use was making pest management decisions based upon sampling and thresholds. Higher level IPM required the use of multiple nonchemical control practices. A subsequent analysis by Benbrook et al. (1996), who used similar definitions for IPM use, produced similar estimates, but placed the use of more biologically intensive IPM approaches at only about 6% ofU.S. crop acreage. The National Agricultural Statistics Service (NASS) completed a national survey of producers of major crops and crop groupings to determine the percentage of crop acreage using various approaches to pest prevention, avoidance, monitoring, and suppression. Although a variety of practices were used to some extent, certain practices were more common. In the NASS survey (USDA NASS, 1998), alternating pesticides to delay resistance (a chemically intensive IPM tactic) reportedly was used on barley (41%), corn (43%), cotton (50%), soybeans (40%), wheat (30%), fruits and nuts (68%), and vegetables (68%). Scouting for pests was used on cotton (75%), corn (47%), soybeans (45%), wheat (30%), alfalfa (24%), fruits and nuts (80%), and vegetables (81%). Tillage or plowing (a cultural or physical IPM tactic) to control pests was used on barley (30%), corn (40%), cotton (63%), soybeans (41%), wheat (37%), alfalfa (21%), and fruits and nuts (74%). Crop rotation (a more strategic IPM approach) was a leading pest management practice to control pests of barley (59%), corn (69%), soybeans (69%), wheat (53%), alfalfa (32%), and vegetables (74%). All values in parentheses represent percentage of total acreage. Although many or even a majority of growers have adopted a minimum level of IPM, why has there not been more progress toward less chemically intensive IPM systems? A significant body of literature (e.g., Wearing, 1988; Zalom, 1993) discusses the technical, financial, educational, institutional, and social constraints to IPM use. The National IPM Forum held in Washington, DC, in 1992 (see Sorenson, 1994) asked participants representing a variety of public and private interests to identify and rate the constraints. Among the top issues identified were the lack of a national commitment to IPM, lack of funding for IPM research and extension activities, perceived problems with the regulatory process that affects registration of new technologies, and the shortage of well-trained, independent IPM consultants. Since the Forum, the national commitment to IPM increased with the Clinton Administration's 1993 pledge to have 75% of cropland acreage under IPM by the year 2000. However, increased funding for IPM research and extension activities has only marginally materialized through narrowly targeted partnership programs established by federal and state agencies. The U.S. Environmental Protection Agency has made significant changes in the way it approaches registrations of new technologies by establishing a fast track for certain compounds, notably insect pheromones. Little new attention has been paid to the issue of training, certification and promoting the use of IPM consultants.
9.6 Conclusion
9.5 ADVISORY SERVICES The development of a pest management consulting industry can be viewed as one of the most positive results of IPM implementation efforts, and consultants have become a major force in the delivery of pest management information to growers (Blair, 1986; Frisbie and McWhorter, 1986). As pest management becomes more complex and business requirements of the farming enterprise compete for their time, growers increasingly rely on the advice of third parties in the pest management decision process. Crop consultants play a significant role in implementing IPM systems, particularly those that are more biologically intensive and, therefore, require a greater degree of information about the status of pests and their natural enemies in the context of the cropping system. References to crop consultants providing pest management information to their grower clientele can be found in the scientific literature over the past 50 years (see Michelbacher, 1945). In the early 1970s, a program for licensing pest control advisers was initiated by the state of California, and over 4000 individuals are now so licensed. The law requires anyone who recommends pesticides or any other pest control method or device for agricultural use to be licensed. This law has had a far-reaching effect on increasing the number of growers who use a minimum level of IPM; for most crops, a higher proportion of acres are scouted in California than elsewhere in the United States. In spite of this, the potential effect of this program perhaps has not been realized because the licensing program does not distinguish private consultants from the majority of consultant who work for farm supply dealers or other chemical retailers (Wearing, 1988). Although most individuals (none in California) no longer receive commissions or bonuses based on their sales of farm chemicals, an incentive to consider alternative practices, including taking no action, is often lacking. In fact, there is a certain margin of safety in deciding to use a product if there is any question about a pest's damage potential. In human medicine, a regulatory system has been adopted whereby drugs, which (like pesticides) are chemicals intended to kill or neutralize the impact of target organisms or deleterious biological processes, are registered to insure their safe use. Some drugs are recognized as safer to apply than others, and these are available "over the counter" to consumers for selftreatment. Other drugs, the application of which have a greater risk of unintended consequences, are available only by prescriptive use as recommended by physicians who have met specific requirements of the profession. At present, most growers practice a form of self-treatment with pesticides, controlling the choice of chemicals and treatment schedules. As long as pesticides are used according to label restrictions, there are few additional restrictions on their availability or use. With few exceptions, there is no requirement that treatments be based on an accurate diagnosis of a problem or whether, in fact, a problem exists. There is no requirement that alternative treatments be considered or that knowledge of alternative treatments exist. This undoubtedly contributes to the public's negative attitude toward agricultural chemicals
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and continuing demands for new regulations. Increasingly more stringent regulations will likely lead to the loss of certain higher risk pesticides or uses of pesticides for which alternatives are less effective or more costly to growers. Prescriptive use of drugs has not eliminated cases of injury due to their application, the development of resistance to drugs, or other ancillary problems. However, there is public confidence in the regulatory system for medicine and drugs that does not exist in the regulatory system for pesticides. Would the prescriptive use of pesticides by licensed practitioners help to improve public confidence in the use of pesticides? Coble et al. (1998) addressed this issue by proposing a model similar to that used in the medical profession whereby relatively low-risk chemicals may be self-prescribed, but highrisk chemicals may be prescribed only by a trained and licensed professional. This proposal is one mechanism by which certain valuable pesticide uses could be maintained, while addressing the public's concern for safe use of those products. Already, pesticides are not treated equally in the registration process. For example, pesticides that present the greatest risk to human health or the environment have various restrictions placed on their use. Pesticides that are believed to be "safe" may be put on a fast track for registration.
9.6 CONCLUSION There are many challenges to the development and implementation of IPM systems, but an excellent framework exists in the scientific literature and in experiences with successful field implementation. The concept of integration began with the realization that the use of synthetic pesticides, which helped to make pest control more predictable and less labor intensive, brought about certain unintended consequences such as pest resistance, secondary pest outbreaks, and the resurgence of pests that previously had been under good control. Integrated control suggested that by utilizing pesticides in such a manner as to preserve naturally occurring biological control, more effective and, in the longer term, more economical pest control could be achieved. Integrated pest management incorporated the concept that pesticides should be used only when needed based upon careful assessment of the risk posed by specific pest densities and the potential for control of those pests by naturally occurring beneficial organisms or other factors in the environment. IPM became more interdisciplinary, incorporating an ecosystem approach. As concern about the impact of pesticides on the environment and on human health became elevated in society, IPM gained favor as an acceptable strategy for managing pests. With wide acceptance of the paradigm came a particular emphasis on IPM tactics within the range of practices that can be utilized in an IPM system. In fact, many practices can be and are utilized in an IPM system to prevent, avoid, and suppress the range of pests that threaten crops, human and animal health, or other elements of
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the landscape. Some observers refer to the range of IPM practices as falling along a continuum from those that are more chemically intensive to those that are more biologically intensive. As older chemistries are replaced with newer classes of pesticides and materials that are biologically derived, users of pesticides can expect to enjoy more possibilities for using products that are selective and are presumed to present less risk to human health and the environment, while providing the opportunity for development and implementation of economically sound IPM systems.
REFERENCES AliNiazee, M. T. (1984). Effect of two synthetic pyrethroids on the predatory mite, Typhlodromus arboreus, in the apple orchards of western Oregon. In "Acarology VI" (D. A Griffiths and C. E. Bowman, eds.), pp. 655-658. Wiley Interscience, New York. Antle, J. M., and Park, S. K (1986). The economic ofIPM in processing tomatoes. Calif. Agric. 40(3/4), 31-32. Benbrook, C. M., Groth, E., Halloran, J. M., Hansen, M. K., and Marqnartdt, S. (1996). "Pest Management at the Crossroads." Consumers Union, Yonkers, NY. Bentley, W J., Zalom, F. G., Barnett, W W, and Sanderson, J. P. (1987). Population densities of Tetranychus spp. (Acari: Tetranychidae) after treatment with insecticides for Amyelois transitella (Lepidoptera: Pyralidae). J. Econ. Entomol. 80, 193-200. Blair, B. D., and Parochetti, J. V. (1982). Extension implementation of pest management systems. Weed Sci. 30, 48-53. Brunner, J. F. (1994). Integrated pest management in tree fruit crops. Food Rev. Internat. 10, 135-157. Carson, R (1962). "Silent Spring." Houghton Mifflin, Boston, MA. Cate, J., and Hinkle, M. (1993). "Integrated Pest Management: The Path of a Paradigm." National Audubon Society, Washington, DC. Cob le, H. D., Bonanno, A R, McGaughey, B., Purvis, G. A., and Zalom, F. G. (1998). "Feasibility of Prescription Pesticide Use in the United States." Issue Paper 9, Council Agric. Sci. Tech. Council on Environmental Quality (1972). "Integrated Pest Management." Council on Environmental Quality, Washington, DC. Croft, B. A. (1990). "Arthropod Biological Control Agents and Pesticides." Wiley, New York. Croft, B. A., and Hoyt, S. C. (I978). Considerations for the use of pyrethroid insecticides for deciduous fruit pest control in the U .S.A Environ. Entomol. 7,627-630. Flint, H. M., Yamamoto, A K, Parks, N. J., and Nyomura, K. (1993). Aerial concentrations of gossyplure, the sex pheromone of the pink bollworm (Lepidoptera: Gelechiidae) within and above cotton fields treated with longlasting dispensers. Environ. Entomol. 22, 43-48. Flint, M. L., and KIonsky, K. (1989). IPM information delivery to pest control advisors. Calif. Agric. 43(1), 18-20. Flint, M. L., and van den Bosch, R. (1981). "Introduction to Integrated Pest Management." Plenum, New York. Fl1sbie, R. E., and McWhorter, G. M. (1986). Implementing a statewide pest management program for Texas, USA. In "Advisory Work in Crop Pest and Disease Management" (J. Palti, and R Ausher, eds.), pp. 234-262. Springer-Verlag, Berlin. Georghiou, G. P. (1986). The magnitude of the resistance problem. In "Pesticide Resistance: Strategies and Tactics for Management." National Academy Press, Washington, DC. Gruys, P. (1982). Hits and misses. The ecological approach to pest control in orchards. Entomol. Exp. Appl. 31, 70-87. Gubler, W. D. (1991). Powdery mildew: Epidemiology and Control. In "Proceedings Nelson J. Shaulis Viticultural Symposium," pp. 44-47. New York State Agric. Exp. Station, Geneva, NY.
Hoppin, P., Liroff, R A, and Miller, M. M. (1996). "Reducing Reliance on Pesticides in Great Lakes Basin Agriculture." International Policy Program, World Wildlife Fund, Washington, DC. Hoy, M. A., Barnett, W W, Hendricks, L. c., Castro, D., Cahn, D., and Bentley, W. J. (1984). Managing spider mites in almonds with pesticide-resistant predators. Calif. Agric. 38(7/8), 18-20. Hoy, M. A., Barnett, W W, Reil, WO., Castro, D., Cahn, D., Hendricks, L. c., Coviello, R, and Bentley, W. J. (1982). Large scale releases of pesticideresistant spider mite predators. Calif. Agric. 36(1/2), 8-10. Hoyt, S. C. (1969). Integrated chemical control of insects and biological control of mites on apples in Washington. J. Econ. Entomol. 62, 74-86. Jimenez, M. J., Toscano, N. C., Flaherty, D. L., Ilic, P., Zalom, F. G., and Kido, K (1988). Controlling tomato pinworm by mating disruption. Calif. Agric. 42(11/12), 10-12. Kogan, M. (1988). Integrated pest management theory and practice. Ann. Rev. Entomol. 49, 559-570. Kogan, M. (1998). Integrated pest management: Historical perspectives and contemporary developments. Ann. Rev. Entomol. 43, 243-270. Krause, R A., and Massie, L. B. (1975). Predictive systems: Modem approach to disease control. Ann. Rev. Phytopath. 13,31-47. Matteson, P. C. (1995). The "50% pesticide cuts" in Europe: A glimpse of our future? Amer. Entomol. 41(4), 210-220. Metcalf, R. L., and Luckmann, W. H. (1982). "Introduction to Insect Pest Management." Wiley, New York. Michelbacher, A. E. (1945). The importance of ecology in insect control. J. Econ. Entomol. 38,129-130. Michelbacher, A. E., and Bacon, O. G. (1952). Walnut insect control in northern California. J. Econ. Entomol. 45, 1020-1027. National Research Council (1989). "Alternative Agriculture." National Academy Press, Washington, DC. Newsom, L. D. (1980). The next rung up the integrated pest management ladder. Bull Entomol. Soc. Amer. 26, 369-374. Prokopy, R. J. (1994). Integration in orchard pest and habitat management: A review. Agric. Ecosyst. Environ. 50, 1-10. Rice, R E., and Kirsch, P. A (1990). Mating disruption of oriental fruit moth in the United States. In "Behavior-Modifying Chemicals for Insect Management" (R L. Ridgway, R M. Silverstein, and M. N. Inscoe, eds.), pp. 193-211. Dekker, New York. SaIl, M. A. (1980). Epidemiology of grape powdery mildew: A model. Phytopath. 70, 338-342. Smith, R F., and AlIen, W W (1954). Insect control and the balance of nature. Sc;' Amer. 190(6), 38-92. Sorenson, A. A. (1993). "Regional Producer Workshops: Constraints to the Adoption of Integrated Pest Management." National Foundation for Integrated Pest Management Education, Austin, TX. Sorenson, A. A (1994). "Proceedings of the National Integrated Pest Management Forum." Center for Agriculture in the Environment, American Farmland Trust, De Kalb, IL. Stem, v., Smith, R. F., van den Bosch, R. F., and Hagen, K. S. (1959). The integrated control concept. Hilgardia 29, 81-97. Stevenson, W. R (1983). An integrated program for managing potato late blight. Plant Dis. 67, 1047-1048. United Nations Food and Agriculture Organization (UNFAO) (1967). "Report of FAO Panel of Experts on Integrated Pest Control." United Nations Food and Agriculture Organization, New York. USDA National Agricultural Statistics Service (USDA NASS) (1998). "1997 Pest Management Practices." Special Circular 1(98), U.S. Dept. Agric., Washington, DC. van den Bosch, R (1964). Practical application of the integrated control concept in California. In "Proc. Intern. Congo Entomol." Vol. 12, pp. 595-597. van den Bosch, R, and Stem, V. M. (1962). The integration of chemical and biological control of arthropod pests. Ann. Rev. Entomol. 7, 367-386. Vandeman, A, Fernandez-Cornejo, J., Jans, S., and Lin, B. H. (1994). "Adoption of Integrated Pest Management in U.S. Agriculture." Agricultural Information Bulletin 707, United States Department of Agriculture, Economic Research Service, Washington, DC.
References
Way, M. J. (1977). Integrated control-Practical realities. Outlook Agric. 9, 127-135. Wearing, C. H. (1988). Evaluating the IPM implementation process. Ann. Rev. Entomol. 33, 17-38. Weber, E., Gubler, W. D., and Derr, A. (1996). Powdery mildew controlled with fewer fungicide applications. Winegrowing Jan/Feb, 13-16. Wigglesworth, V. B. (1950). The science and practice of entomology. Adv. Sci. 7, 154-161. Wilson, L. T., Hoy, M. A., Zalom, F. G., and Smilanick, J. M. (1984). The within-tree distribution and clumping pattern of mites in almond orchards: Comments on predator-prey interactions. Hilgardia 52(7), 1-13. Zalom, F. G. (1993). Reorganizing to facilitate the development and use of integrated pest management. Agric. Ecosystems Environ. 46, 245-256.
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Zalom, F. G., Walsh, D., Stimmann, M. W, PickeI, c., Krueger, W, Buchner, K, and Brazzle, J. (1998). Impact of pyrethroids on beneficial mite predators. In "Proceedings of the California Plant and Soil Conference," pp. 62-67. Agronomy Society of America, California Chapter, Sacramento, CA. Zalom, F. G., Hoy, M. A., Wilson, L. T., and Bamett, W W (1984). Presenceabsence sequential sampling for web-spinning mites in almonds. Hilgardia 52(7), 14-24. Zalom, F. G., and Strand, J. F. (1990). Expectations for computer decision aids in rPM. AI Appl. Nat. Res. Manag. 4(1), 53-58.
CHAPTER
10 Toxicity Testing Donald J. Ecobichon Queen's University
10.1 INTRODUCTION Although governmental concerns about the safety of chemical pesticides were evident in the early 1900s in many agricultural nations, these concerns did not prompt legislation and regulations until the late 1940s when organochlorine insecticides were introduced into agricultural use. Important changes in pesticide legislation, particularly concerning regulations about testing, were promulgated in the late 1950s and early 1960s when the much more toxic organophosphorus and carbamate ester insecticides appeared and concomitant increases in the number of poisonings were reported among agricultural workers, bystanders, and produce consumers. An examination of the development of pesticide testing requirements in different nations reveals a parallel with the evolution of tests required for drugs, industrial chemicals, and home and health care products. Most industrialized countries have developed legislation/regulations for pesticide testing, registration, and use. Now companies that wish to register a pesticide product are required to submit a toxicity data base that is comparable to that required by the U.S. Environmental Protection Agency (EPA) under the Federal Insecticide, Fungicide and Rodenticide Act (FIFRA) as is shown in Table 10.1. In addition to these specific test procedures, a considerable amount of environmental testing is required-a topic beyond the scope of this chapter and this book. Although distinct national requirements have evolved over year of test development, recent emphasis has been on international harmonization of testing requirements for pesticides, thereby permitting a chemical registered in one country to be registered in others, given that the toxicity data base meets common standards. Such a change would simplify the regulatory procedures for industries, which currently must meet variable, individual, and national regulatory requirements, a:1d thereby reduce the efforts, cost, and time for multinational registrations. If the flow of toxicological information is coordinated internationally, it would be extremely advantageous to emerging nations, some of whom have shortages of trained scientific, technical, and legal professionals to develop their own legislation. The availability of this information would allow developing countries to adopt a common regulatory framework for pesticides. Such harmonization is being encouraged by the European Economic Handbook of Pesticide Toxicology
Volume I. Principles
Community (EEC), the Organization for Economic and Cooperative Development (OECD), the World Health Organization (WHO) via the International Programme on Chemical Safety (IPCS), and the United Nations Food and Agricultural Organization (FAO).
10.2 TESTING STRATEGIES As defined by Hayes (1975), toxicology is "the qualitative and especially the quantitative study of the injurious effects of chemical and physical agents, as observed in alterations in structure and response in living systems; it includes the application of the findings of these studies to the evaluation of safety and to the prevention of injury to humans and all useful forms of life." The aim of any of the studies listed in Table 10.1 is to develop a quantitative dose-response (effect) relationship between level of exposure and measured biological effects, hopefully, thereby, permitting extrapolation of the study data from animal surrogate models to humans. Within reason, extrapolations and predictions are possible for all classes of insecticides, as well as some rodenticides and fungicides, because target sites and mechanisms of action in most, if not all, mammalian systems and species are comparable. However, for many other classes of pesticides, the toxicity encountered has less to do with the active agent and more involvement with low levels of contaminants and by-products that are capable of causing unique, adverse, biological effects. The extent of the testing required for any new pesticide must address any and all possible toxic events, both predictable and idiosyncratic, as well as the concerns of product safety raised by workers, bystander populations, and the general public-a tall order indeed! As is found with any chemical, all pesticides possess some degree of inherent toxicity-a property of the agent that is as distinctive as the physical and chemical properties of the molecule. This is a basic tenet of toxicology. Thus, anyone member of a class of chemicals should behave qualitatively in the same manner as any other homolog, although the concentration required to elicit a given biological effect may vary considerably. A second basic tenet is exposure, that is, the
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Copyright © 2001 by Academic Press All rights of reproduction in any form reserved.
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CHAPTER 10 Toxicity Testing
Table 10.1 Federal Toxicity Testing Requirements for a Pesticide
Table 10.2 Toxicity Data Base Development for Chemicals-Progression of Studies
Acute orallethality
First wave
Acute dermallethality Acute inhalation lethality
Acute (lethality) studies, via routes of exposure anticipated for the human
Primary dermal irritation
Irritation studies (ocular, dermal)
Dermal sensitization
Dermal sensitization studies
Primary ocular irritation
Mutagenicity studies with in vitro microbial and
Acute-delayed neurotoxicity
mammalian cell lines
21-Day dermal exposure
Teratogenicity studies (mouse, rat, rabbit) with agent
90-Day dermal exposure
administered to normal, pregnant animals during
90-Day feeding study 90-Day inhalation exposure 90-Day neurotoxicity assessment Chronic feeding study Oncogenicity study Teratogenicity study Reproduction study Gene mutation Chromosomal aberration studies
organogenesis Second wave Subchronic studies, 21-90-day feeding studies in rodent and nonrodent species; may use other routes of exposure Chronic/oncogenicity studies of 6-month (rodent) duration for the former and 24-month (or life-span) for the latter in rodents Reproductive studies, generally in rodents: effects are studied in both males and females
Other genotoxic effects General metabolism
Source: Ecobichon (l997a-d).
Domestic animal safety Source: V.S. Code of Federal Regulations, Title 40, Part 158 (40CFR Part 158).
amount of agent necessary to elicit an adverse biological effect. Exposure consists of two components: the concentration of the agent and the duration (time) of exposure, which reflects the possibility of both acute, high-level, short duration as well as low-level, long duration exposures. The third basic tenet of toxicology arises from the other two, that is, the previously mentioned dose-response relationship, where the degree of adverse response is proportional to the level of exposure. This evaluation entails a knowledge of the levels at which adverse biological effects begin to appear. A threshold toxicant is presumed to pose no risk below some experimentally determined concentration. In contrast, a nonthreshold toxicant is presumed to pose some risk at all dosages above zero. The exorbitant costs of conducting extensive toxicological assessments of pesticides makes it imperative that any potential toxicity be discovered as early as possible in the development program. Toxicity studies are carried out in two waves (Table 10.2). The first wave consists of studies that can be completed in a relatively short time period (approximately two months) and at minimal cost, giving good indications of what hazards might be encountered following acute exposure and providing clear options whether to abandon the agent or proceed. The second-wave package consists of the more costly, time-consuming, longer-term studies required to examine chronic effects, including fertility, reproductive outcomes, and carcinogenicity, on various target organs. The results of these studies will be in hand only after some 3-4 years.
10.2.1 FIRST-WAVE STUDIES
Given the distinct possibility that, at some time in the production and use of the pesticide, some individual(s) (factory workers, handlers, mixer-loaders, sprayers, bystanders, general public) will be exposed accidentally or by intent to toxic levels of the pesticide, studies that involve exposure to high concentrations of the agent are essential to determine its potency relative to other, known agents tested in the same species. The potency will be assessed in three ways: (1) acute lethality of the agent along with other selected endpoints of acute toxicity; (2) irritancy to both skin and eyes; (3) dermal sensitization. This information is important for the occupational health and safety of workers and will be incorporated into the material safety data sheet (MSDS), which, by law, must be prepared and made available for every chemical manufactured and used. In addition, the potential of a chemical to cause mutations in the nucleic acid of microbial and mammalian cell systems will be examined at this time, because an extension of this biological activity is the occurrence of adverse reproductive outcomes (including birth defects) and cancer. Teratogenicity studies are generally conducted at this stage, providing possible confirmation of positive mutagenicity results. 10.2.1.1 Acute Lethality
The potency of a chemical is frequently expressed in terms of the median lethal dose (LD50). This index is defined as a "statistical estimate of the acute lethality of an agent administered under defined and controlled conditions to a certain sex, age and strain of species of animal." The route of administration (oral, dermal, inhalation) is selectively based on that route by
10.2 Testing Strategies which humans are expected to be exposed. Standardized protocols are plentiful for designing and conducting this test (Chan and Hayes, 1994; Ecobichon, 1997a). Unless it is indicated, the LDso is assumed to represent the median lethal dose for deaths that occur in the first 24 hours after treatment. In the case of inhalation and dermal studies, the LDso is actually an LCso, that is, the concentration in the breathing air or applied to the skin rather than the known amount of agent taken up by the body via the lungs or through the skin. The value of the LDso rests in its use as an index of relative toxicity of the unknown chemical compared to the toxicity of known chemicals administered by the same route to the same species, strain, age, and sex of test animals. Within a certain class of chemicals that have a similar mechanism(s) of action, direct comparisons between various members can be made concerning relative potency. Such agent-to-agent comparisons should not be done for chemicals of diverse structure and mechanisms of action, that is, insecticides, herbicides, fungicides, avicides, molluscicides, etc. Although such comparisons are commonly made, it is far more complex than comparing apples, oranges, and bananas, and is a serious limitation of the LDso value obtained experimentally. However, it is important to emphasize that acute toxicity studies are not limited to lethality alone. Classic oral LDso determinations are conducted in an appropriate surrogate animal model (mice, rats, etc.) to ascertain the total adverse biological effects during a finite period of time following the administration of the test agent, by an appropriate route, as single, frequently large, doses to various groups (n = 10 per treatment) of preselected animals. Lethality, expressed as the LDso (mglkg body weight), is one endpoint determined at 24 hours posttreatment. The appearance and behavior of the animals is closely observed during the 24-hour period to detect possible mechanisms of action of the test agent. All dead animals and those moribund animals that are euthanized to prevent suffering will be subjected to gross anatomical examination and tissues are preserved for microscopic study. The animals that survive the toxic insult will be monitored for 14 more days. These surviving animals provide a vast amount of information on behavior, growth and development, recovery, persistence of signs and symptoms, secondary target organ toxicity and delayed toxicity as well as biochemical parameters from blood and urine. On day 14 posttreatment, the surviving animals are euthanized and undergo gross examination. Terminal samples of biological fluids are collected for analysis and tissues are secured for preservation and microscopic examination. Dermal exposure to pesticides is a constant hazard and concern. National and international regulatory agencies require that estimates of lethality via the dermal route be determined for active ingredients as well as for concentrated solutions and end-use formulations (Ecobichon, 1997a; Patrick and Maibach, 1994). In such studies, the test agent is applied to the closely clipped dorsal skin, one dose per group, to several groups of male and female rodents (mice, rats) and nonrodents (guinea pigs, rabbits). The test agent, as a powder, paste, or liquid (neat
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or dissolved in a vehicle), is spread uniformly over the test site, which should be equivalent to approximately 10% of the total body surface (3.0 x 4.0 cm for rats; 5.0 x 6.0 cm for rabbits; Ecobichon, 1997a). A number of doses are used, and the range is spaced to elicit toxic effects and lethality. The test agent is left in place for 24 hours with the site occluded by a gauze pad taped securely to the skin and the entire region covered by a rubber sheet or a form-fitting elastic cloth sleeve. After 24 hours of exposure, the unabsorbed material is washed off with warm water and soap, avoiding vigorous scrubbing. The acute lethality is expressed as the LCso (median lethal concentration, mglkg). If chemical-related toxicity and/or mortality are not observed on reaching a dose of 2000 mglkg body weight, further study at higher doses is unnecessary. The likelihood of exposure to vapors, aerosols, and/or particulates of pesticide formulations is high among agricultural workers both during and after spraying; thus determination of the acute lethality by the inhalation route is necessary. Groups of experimental animals-rodents and nonrodents, males and females of each-will be exposed to the test substance for a defined period, usually 4 hours, at a graduated range of three concentrations, one concentration per group. The animals may be accommodated within an inhalation chamber (whole body exposure) or in a head-only or nose-only apparatus. The desired concentration of aerosolized active ingredient or complete formulation is generated via a nebulizer system. At the end of the test period, the animals are removed from the test apparatus and are returned to their cages for further monitoring as well as determining the LCso, the concentration in the breathing air of the animal that resulted in acute lethality in 24 hours. The surviving animals are monitored biochemically and physiologically for 14 more days, as previously described. The multitude of problems inherent in acute lethality studies by inhalation have been discussed (Ecobichon, 1997a). Most national and international regulatory agencies no longer demand determination of a precise LDso, but do require an estimation of the acutely lethal dose that can be obtained using a limited number (n = 6 or 8) of animals rather than the 50-60 animals required in the preceding test. What is desired is a rough estimate of the toxicity of the agent so that suitable ranges of doses can be defined, thereby avoiding excess early mortality in future subchronic and chronic studies. The up-anddown or staircase method uses only one animal per dose; a second animal will receive a higher or lower dose, depending on whether the first treated animal dies (Ecobichon, 1997a). Although it is used mostly for oral LDso determination, the same approach would be valid for dermal and inhalation routes of exposure. The fixed dose approach avoids the death of the animals as an endpoint, relying on clear observations of toxicity at one or another of a set of preselected or fixed doses (5.0, 50, 500, and 2000 mglkg), where the results permit classification or ranking of the chemical according to the system used by the EEC as shown in Table 10.3. Oral and dermal LDso values have been utilized by the WHO to prepare a classification of pesticides that distinguishes between the more and the less hazardous forms of each pesticide.
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Table 10.3 Ranking System Used by the European Economic Community for Acute Oral Toxicity of Chemicals Category
LDSO (mg/kg)
Very toxic
25 25-200 200--2000 2000
Toxic Harmful Unclassified Source: European Commission Directive 83/467IEEC (1983).
Such a distinction pennits formulations to be classified according to percentage of active ingredient and physical state (Table 10.4; Copplestone, 1988). It must be emphasized that the quoted LD50 values are not the median values, but are the lower confidence limit values for the most sensitive sex, thereby ensuring that a large safety factor has been built into the classification. The use of a single value avoids confusion or uncertainties in the case of products that have confidence limits that lie across a class boundary. The WHO is of the opinion that this classification scheme works well in practice, reflecting the idea that acute toxicity is the most effective parameter by which to judge the hazards of pesticides to human health. 10.2.1.2 Dermal Irritation Any chemical that may come into contact with the skin must undergo primary irritation studies using suitable animal models. Pesticides are no exception, because occupational dennatitis is a major problem among pesticide users. Several standard protocols have been described for such studies; most are only slight modifications of the original Draize test (Draize et aI., 1944; Ecobichon, 1997a; Patrick and Maibach, 1994). All tests require closeclipping of the animal's dorsal hair to provide a suitable test site, preparing several areas of intact as well as Table 10.4 World Health Organization Recommended Classification of Pesticides by Hazard LDsO for the rat (mg/kg body weight) Dermal
Oral Class la
Extremely
Solids
Liquids
Solids
Liquids
:sS
:s20
:s10
:s40
5-50 50--500
20--200 200--2000
10--100 100--1000
40--400 400--4000
>500 >2000
>2000 >3000
> 1000
>4000
hazardous lb
Highly hazardous
II
Moderately hazardous
III
Slightly hazardous
III
Unlikely to present hazard in normal use
Source: Copplestone (1988).
abraded epidennis, and applying suitable amounts of the test agent (not exceeding 1.0-2.0 ml/kg neat if liquid or dissolved in a vehicle; not exceeding 2000 mg/kg if a powder or paste) uniformly over the test sites (3 x 4 cm for rats; 5 x 6 cm for rabbits). The site is usually occluded to prevent the animals from removing the test substance during the 24-hour exposure period. After 24 hours, the residual, unabsorbed material is washed off using warm soapy water and the dermal reaction (erythema or redness, edema or puffiness) is assessed immediately and at 48 and 72 hours, using a standardized, subjective scoring system that has numerical values from 0 (no effect) to 4.0 (severe effects; McCreesh and Steinberg, 1983). Thus, for any pesticide, it can be detennined whether the components of a formulation (co-solvents, emulsifiers, adjuvants, etc.) play any role in the dennal penetration of the active ingredient or exert any effect on toxicity. Two scenarios could occur: a formulation component could either enhance or reduce the potency of the test agent. Valid concerns about the infliction of pain and suffering on animals in primary irritation studies has led to substitution of in vitro studies that use established mammalian cell lines and/or epidennal surrogates and examine a number of biological endpoints that measure cell penneability, cytotoxicity, and inflammation (Ecobichon, 1997a). Many of the assay systems listed in Table 10.5 have been validated, and good agreement has been obtained between in vitro results and previously determined in vivo effects. In the future, greater use will be made of test batteries of in vitro assays to screen new chemicals to eliminate highly to moderately toxic agents, leaving those of apparent low toxicity to be evaluated in animals. 10.2.1.3 Dermal Sensitization Repeated dermal exposure to a number of chemicals results, over time, in a hypersentitive, allergy-like reaction similar to that seen for certain foods, pollens, plants, venoms, etc. The initial exposure elicits little or no reaction, but repeated exposure may cause dennal swelling and redness at the site of exposure. Each time exposure occurs, the reaction occurs more quickly and persists for a longer period of time. This mechanism appears to be similar to the classical allergic reaction, where the agent that acts as a hapten binds with proteins to form an antigenic complex that is recognized as foreign to the body and triggers cell-mediated production of IgE antibodies. A number of test procedures have been described, usually using the guinea pig as the surrogate model (Chan and Hayes, 1994; Ecobichon, 1997a; Seabaugh, 1994). All tests are similar in that an attempt is made to develop an immunological "awareness" by daily application of the test agent to the shaved skin of the animal over a period of 24 days, followed by a rest period of 10-14 days. A challenge dose, usually smaller than that used to raise the sensitivity, is administered to a fresh, untreated dermal site, and the severity of the redness and swelling is scored at 24, 48 and 72 hours posttreatment, the same as mentioned for dermal irritation. Depending upon the severity of the immunological reaction, the test agent can be classified as a weak-to-strong sensitizer.
10.2 Testing Strategies Table 10.5 Alternative Test Systems to Assess Dermal Irritation Using Cell Cultures
Table 10.6 Test Methods for Skin Sensitization Acceptable to the EEC
Assessment
Test
Test endpoint
Draize Test
Cytotoxicity
Trypan blue
Dye inclusion in damaged
Neutral red
Dye sequestered in lyso-
cells somal membranes of viable cells Cell viability
Kenacid blue
Total protein content
Coomassie blue
Buehler Test Guinea Pig Maximization Test Mauer Optimization Test Freund's Complete Adjuvant Test (FCAT) Open Epicutaneous Test Split Adjuvant Test Information sources: OECD Test Guideline 406 (1981); Patrick and Maibach (1994); Seabaugh (1994); Ecobichon (1997a).
Lowry protein Rhodamine or
289
Total protein content
Nile blue Hexosaminidase
Cell activity metabolism
Leakage of cytosolic
Lactic acid
enzymes through damaged
dehydrogenase 14C-Ieucine
cell membranes Transport and accumulation
3H-uridine
intracellularly and in-
3H-thymidine
corporation into proteins or nucleic acids
Mitochondrial
Reduction of yellow tetra-
function
zolium salt (MTT) to blue-black, insoluble formazan
Cell metabolism
Microtox™
Luminescent bacterium that emits light during metabolism (reduction of light)
Irritation
Skintex™
Test agent induced alterations of macromolecular matrix of keratin-collagendye membrane, with release
shaved, tape-stripped site for 3 consecutive days; (2) a challenge dose on day 10, applied to one ear and vehicle applied to the other; (3) measurement of the ear thickness by micrometer at 24 and 48 hours posttreatment. The detection of weak haptens can be enhanced by using abraded skin (Botham et aI., 1991). A more promising test, the murine local lymph node assay (LLNA), has shown remarkable responses to sensitizing agents (Kimber and Basketter, 1992). The protocol, using 8-week-old mice, involves daily, topical treatment with agent in solution (25 J.!l) on the dorsum of both ears for 3 consecutive days and, 5 days later, an injection (IV) of 20 J.!Ci of 3H-methylthymidine or 3H-thymidine in saline. After an additional 5 days, the mice are euthanized and the auricular lymph node cells are recovered. After incubating overnight in suspension, the lymph node cells are subjected to ,B-scintillation counting to determine a treated/control ratio. Validation of this test has demonstrated the correct identification of human sensitizers (Basketter et aI., 1994).
of dye Inflammatory response
Cell culture
Measurement of eicosanoids
(chemical
and cytokines from keratino-
mediators)
cytes incubated with test agent
Currently, seven tests, all using the guinea pig, are accepted by the OECD and EEC (Table 10.6). Guinea pigs are used as the test animal because of their recognized susceptibility to a wide variety of chemical sensitizers. All of these tests involve repeated, intermittent, topical or intradermal application of agent, with or without Freund's adjuvant, examination of the site of application for erythema and edema after a 2-3-week "rest" period, and administration of a challenge dose of the test agent. Details of these individual tests were given by Patrick and Maibach (1994) or the references listed in the table may be consulted. The mouse ear-swelling test (MEST) was developed as an accurate, sensitive, alternative test to evaluate sensitization (Gad et aI., 1986). The test includes (1) an induction phase with topical application of the agent in liquid form to an abdominal,
10.2.1.4 Ocular Irritation The inadvertent, accidental splashing of a diluted or concentrated pesticide formulation during handling (owing to the frequent absence or lack of use of appropriate protective equipment) makes it mandatory that the potential of an agent to cause ocular irritation be tested. The Draize eye test has been the standard method used to assess product safety (Draize et aI., 1944). Few modifications have been made to the basic methodology with the exception of the low-volume test, but additional, quantifiable parameters have been introduced, including erythema, thickness of the eyelids and nictitating membranes, edema, discharge, corneal opacity, capillary damage, and pannus (vascularization) of the cornea (Buehler, 1964; Conquest et aI., 1977). The standard ocular irritation test is no longer in favor because of perceived trauma and pain to the eyes of test animals. In reality, if a thoroughly conducted dermal evaluation reveals irritational or corrosive properties for the test agent, there is no need to conduct the eye test. However, given the fact that the corneal membrane and the epidermis are quite different in structure, false negative results with the skin could mask serious ocular problems. A number of in vitro tests that use mammalian cell lines, isolated, intact eyes from chickens, or isolated bovine
290
CHAPTER 10 Toxicity Testing
corneas have been designed and introduced to reduce and replace the use of animals and to refine the ocular irritation test (Ecobichon, 1997a). These tests assess (1) cellular cytotoxicity, which affects the uptake or exclusion of dyes by cell cultures, (2) opacity, that is, light transmission through the isolated bovine or rabbit corneas, and (3) inflammation, which measures the release of inflammatory mediators (histamine, serotonin, leukotrienes, prostaglandins) from the bovine corneal epithelium (Ecobichon, 1997a). Although validation of these tests points to their utility, it is likely that these in vitro assay systems will be used to screen moderate to severe irritants, reserving the animal test for those agents being contemplated for use in formulations. All final products will continue to be tested by the Draize test to assure the consumer that damage (reversible or irreversible) to the eye cannot occur. In the standard ocular irritation test, 0.1 ml of liquid or 100 mg of solid is instilled into the pouched, lower conjunctival sac of one eye of each of six rabbits and the eyelids are held together for 1.0 second and then released. The treated eye is not washed; rather, the animal's own tear secretions are allowed to flush the eye. The untreated eye serves as a control. Generally, the eyes are irrigated with 0.9% physiological saline after 24 hours. Both eyes are examined at 1, 24,48, and 72 hours following treatment. Such studies are usually terminated at 72 hours, but if residual injury is present, examination of the eyes may be prolonged at the discretion of the investigator. Damage to the cornea, conjunctiva, and iris is scored subjectively according to the numerical system described by Draize et al. (1944). To improve the reproducibility of the scoring, slit-lamp examination and staining of the eye with fluorescein dye are frequently employed. A low-volume test (LVET), using 0.01 ml of liquid or 0.01 g of solid has been introduced, wherein the agent is applied directly to the cornea and the eyelids are not held shut following treatment. Using only three rabbits, good correlations were obtained between the LVET results and the human response. The LVET predicted human ocular exposure more accurately than the standard test (Bruner et aI., 1992; Freeberg et aI., 1986). In contrast to the preceding study design, more recent investigations have demonstrated that high-level accuracy can be obtained with fewer rabbits. Two- or three-rabbit tests have been shown to be accurate in classifying the irritational potential of known agents to levels of 88-91 and 93-94%, respectively (De Sousa et aI., 1984; Talsma et aI., 1988).
10.2.1.5 Mutagenicity Many chemical mutagens interact with cellular deoxyribonucleic acid (DNA) by alkylating reactions. The results of interaction are manifested as cellular lethality (severe cases) or as chromosomal breaks, dysfunctions, deletions, fragments, exchanges, point mutations, etc., that are replicated faithfully through descendent cells. It is accepted dogma that the consequences of subtle alterations in chromosomal material may be expressed as infertility, embryolethality, embryotoxicity, spontaneous abortion, congenital anomalies, altered resistance to infections, reduced life-span, and even carcinogenesis. To attempt meaningful "mutagenicity" studies in surrogate animal
models and to measure the above endpoints collectively is possible, but a lot of time may be consumed, the costs will be excessive, and the results will often be clouded by confounding factors. These problems led to the search for relatively simple, short-term, inexpensive, sensitive, and reproducible biological tests that were capable of predicting the mutagenic potential of chemicals. The search has been successful to a degree: a plethora of rapid, sensitive assays currently are available to explore a variety of mutational end points and even to give insight into the mechanism(s) by which these events are produced. The utility of these tests will be to screen the vast array of untested chemicals for mutagenic properties. A variety of microbial systems using prokaryotic, auxotrophic bacteria (Salmonella typhimurium, Escherichia coli), yeasts (Saccharomyces cervisiae), and fungi (Neurospora crassa) have been developed based on extensive knowledge of the genetics of various strains of the organisms. The experiments are conducted rapidly and reproducibly, and the organisms are responsive to the agents being tested (Ecobichon, 1997b). The prototype assay was the Ames test (Ames et aI., 1973, 1975), which used genetically defective strains of S. typhimurium as the indicator organism. These strains depend on provision of an essential nutrient for growth (histidine for S. typhimurium; other organisms require sugars, amino acids, purines, pyrimidines, etc.). When a standard mixture of organisms, top agar, and a small amount of essential nutrient was incubated with a range of concentrations of the suspect mutagen and then grown for 24 hours on agar plates that were deficient in the essential nutrient, a quantitative, dose-related increase in the number of revertant colonies was visible that indicates mutagenic reversion (reverting to prototypic or wild-type organisms capable of synthesizing the essential nutrient). At high concentrations, reverse mutation would be obscured because of excessive lethal mutations that result in cell death. With this simple assay, a quantitative relationship can be developed between the concentration of suspect mutagen and the extent of reverse mutation of the organism, at minimal cost and in a relatively short time (Ames et aI., 1975). Many mutagenic substances exist as promutagens, which are inactive per se and require biotransformation into reactive intermediates before any mutagenic activity is exhibited. Microorganisms are incapable of carrying out such reactions rapidly. Hence modifications in the microbial assay to include an active metabolizing enzyme system are needed. The usual modification involves adding an aliquot of a 9000-g (S-9 fraction) supernatant obtained from homogenized rodent (mouse, rat) liver from a donor animal that was pretreated with a suitable enzyme-inducing agent (polychlorinated biphenyls, phenobarbital, ,B-naphthoflavone, etc.) to provide maximally induced, hepatic, microsomal Phase I enzymes. Before the test mixture is spread on the minimal agar, the S-9 aliquot plus suitable cofactors are incubated with the microorganism and test substance to ensure conversion of the promutagen to a reactive, possibly mutagenic, intermediate that is capable of causing a mutagenic effect in the nucleic acids of the microorganism. Such an "activity system" can be used with any of the prokaryotic organisms
10.2 Testing Strategies
and also with eukaryotic, mammalian cells as described subsequently. The apparent supersensitivity of the Ames test (almost everything was positive) and the prudence of not relying on anyone particular bioassay, resulted in the development of other in vitro assays that use a variety of immortalized, standardized, eukaryotic, mammalian cell lines [Chinese hamster ovarian cells (CHO), hamster pulmonary cells (V79), mouse lymphoma cells, fibroblasts, leukocytes] to examine various endpoints such as clastogenicity (chromosomal aberrations, breaks, fragments, chromatid exchanges, micronucleus formation) and gene mutations (forward or reverse) that result in unscheduled DNA synthesis, the loss or appearance of specific enzymatic functions, etc. Mutagenic assay systems have also included plants (onion root tip, allium, tradescantia) and insects (Drosophila melanogaster), but such tests take somewhat longer to complete and yield results that are more difficult to interpret beyond the observation that a mutation has occurred, and growth and development are affected. The overall assessment of the mutagenic potential of a chemical is based on the spectrum of positive and negative results from five or six assays selected from a test battery such as that shown in Table 10.7. All the tests in Table 10.7 are representative, reproducible, and well-established assays that examine different facets of genetic toxicology and allow exploration of the mechanism(s) of action of the test substance. Such a group of assays can be completed within 30 days at moderate cost.
Table 10.7 Test Battery for Mutagenic Evaluations of Chemicals Type Bacterial Mammalian (in vitro)
Test system/assay
Function
Ames salmonella-liver S9
Reverse mutation
Escherichia coli
Reverse mutation
Mouse lymphoma L51778 (TK + / - )
Forward mutation
Chinese hamster ovary HGPRT
Forward mutation
Pulmonary V79 cells HGPRT
Forward mutation
Mammalian cell lines
Chromosome aberrations Sister chromatid exchange
Mammalian
Drosophila melanogaster
Sex-linked recessive
Rodent bone marrow stem cells
Chromosomal
(in vivo)
lethal assay aberrations Micronucleus formation Rodent (mouse, rat) dominant
Lethal mutations
lethal assay Source: Data from National Research Council, Toxicity Testing, Strategies to Determine Needs and Priorities, National Academy Press, Washington, DC (1984).
291
10.2.1.6 Teratogenicity Teratology may be defined as the study of the generation, causes, and manifestation of structural (anatomical) and functional (metabolic or physiological dysfunction, psychological, behavioral) alterations in development. The rapid growth and development of embryonic and fetal tissues significantly increase the susceptibility of cellular DNA to toxicant-induced changes in gene expression or in the retardation and/or arrest of cellular development. It is important to appreciate that biological effects can be elicited in embryonic tissues by chemicals at concentrations far below those that cause target organ toxicity in adults. Such chemicals, called teratogens, are a major concern to the public, to industry, and to regulatory agencies because of the low levels required to initiate cellular damage (in some cases, at levels on the order of those detected in various environments). The fundamental aim of screening studies used to evaluate the teratogenic potential of chemicals is to predict the absence of a teratological hazard for humans. Toxicity may be elicited in the pre- and postimplantation embryo, causing death (embryolethality) or mild to severe dysmorphogenesis in one or more organ systems, resulting in structural malformations and physiological or biochemical dysfunction, as well as psychological, behavioral, and cognitive deficits in the offspring at birth or in a defined postnatal period (Schardein, 1985; Ecobichon, 1997c). Timing is critical in the design of teratogenic studies because exposure must occur during the period of organogenesis, beginning after the implantation of the fertilized ovum. This event varies between animal species, as is shown in Table 10.8 (Ecobichon, 1997c). In the human, organogenesis occurs in the first 8 weeks of pregnancy. Different organ systems develop at different but sometimes overlapping rates and time periods. Frequently, the critical "toxic window" of time, when an agent can elicit an effect, is only a few days duration. Agents acquired before and after the "window" show no effect on one particular target organ but, perhaps, cause some adverse effect on another developing organ system. The concept of the toxic window becomes crucial when chemicals are tested in surrogate animal models because of the shortened gestation time and windows as narrow as 24 hours in which the appropriate concentration of agent must be in the right place to elicit an effect (Ecobichon, 1997c). Upon completion of organogenesis, further chemical exposure will not cause teratogenicity except in the still developing central nervous system, which, in most species continues to develop throughout gestation and well into the postnatal period. Frequently, behavioral and cognitive deficits are detected. The reader is referred to the chapter by MacPhail (Volume 1, Chapter 12) on neurobehavioral toxicity for further details. Assessment of teratogenicity is routinely conducted in two species, a rodent (mouse, rat or hamster) and a nonrodent (rabbit), and involves the daily administration of a range of three appropriate dosages to different groups of timed-pregnant animals throughout the period of organogenesis (e.g., days 6-15 for rats and days 6-18 for rabbits). The pregnant animals are euthanized 24 hours prior to the calculated day of parturition
292
CHAPTER 10 Toxicity Testing
Table 10.8 Species Differences in Gestational Endpoints (in days) Species
Implantation
Organogenesis
Gestation
10--56
270
Human
7-8
Rat
5.5-6
6-15
21-22
Mouse
4.5-6
6-15
19-21
Hamster
4.5-5
4--14
16
Rabbit
7
6-18
32
Guinea pig
6
6-20
67
Monkey (Rhesus)
9
9-40
165
7-35
114
Dog
13-14
Pig
10--12
63
Source: Ecobichon (1997 c).
and undergo a complete autopsy. The numbers of dead (late fetal deaths) and live pups are determined, the uterine muscle is examined for reabsorption sites indicative of embryonic deaths, and the ovaries are examined for the number of corpora lutea to determine the number of ova released prior to fertilization. The latter provides an index of fertility in the female. The position of each pup in the uterine horn is recorded, each is weighed, the sex is determined, and each fetus is examined for external abnormalities prior to dissection to detect internal malformations. Whole fetuses may be fixed in special fluids for histological examination or for the detection of skeletal anomalies. The design and conduction of teratogenicity studies has been described in detail elsewhere (Ecobichon, 1997c; Manson and Kang, 1994). Concerns about subtle changes in behavior and cognitive functions have led to a modification of teratological study protocols wherein some of the treated females are allowed to give birth and to rear their young for 6 weeks or longer after parturition, at which time the pups can be examined by testing strategies designed to assess behavioral and cognitive development. In such protocols, treatment may begin on the last day of implantation and continue throughout the gestational period and even into the postpartum lactational period up to weaning. With this modification in dosage regimen, the developing young receive the agent both transplacentally and via the milk throughout the longer period of neurological development-an important consideration with pesticides and their known covert actions On neuronal membranes, receptors, and neurotransmitters. 10.2.1.7 Acute Neurotoxicity Unfortunate experiences with a few of the early organophosphorus ester insecticides, principally the agent leptophos (0-4bromo-2,5-dichlorophenyl O-methyl phenylphosphonothioate, M PHOSVEC ), led to the mandatory requirement by the U.S. EPA that all pesticides undergo an acute delayed neurotoxicity assessment using the domestic chicken (Callus domesticus) as the preferred, experimental animal (Abou-Donia, 1981; U.S. EPA, 1978). The study is "acute" in the sense that, frequently, a single dose of the test substance is administered to the animal,
but "delayed" in aspect because the animals must be studied for some 7-21 days before peripheral and central neurological effects can be detected and quantified both physiologically and morphologically (Durham and Ecobichon, 1984, 1986; Slott and Ecobichon, 1984). Chemicals used as positive control standards and known from extensive study to produce the characteristic effects of leptophos include such organophosphorus esters as mipafox (N, N-di-isopropylphosphorodiamidic fluoride), DFP (O,O-di-isopropylphosphorofluoridate), and TOTP (tri-o-tolyl phosphate). The subject of organophosphate-induced delayed polyneuropathy (OPIDP) has been extensively reviewed and discussed (Abou-Donia and Lapadula, 1990; Cranmer and Hixon, 1984; Ecobichon, 1996) and is the subject of a lengthy chapter in this text by Ehrich and Jortner (Volume 2, Chapter 52). The discussion in this section will be restricted to the format of the test protocol. Adult hens of 2.0-2.5 kg body weight are housed in individual stainless steel, mesh-bottomed cages and acclimatized for 3--4 days in an environment that has a 12-hour light/dark photoperiod cycle. The birds are assessed visually for neurological competence (locomotion, posture, equilibrium-coordination, and walking strength). Groups of birds are fasted for 18 hours prior to receiving single doses of agent, usually by oral gavage using a fine rubber catheter attached to a syringe. Appropriate doses of test agent and a positive control chemical are prepared in a suitable vegetable oil solution. In situations where signs of acute poisoning are observed following treatment, atropine sulfate (30 mg/kg body weight) can be administered subcutaneously as required to alleviate the signs/symptoms. A physical examination of the control and treated birds is conducted every 24 hours by removing the hens from their cages, placing them On the floor, and observing any changes in behavior and walking ability. The responses of the hens are scored independently by two individuals, who have nO knowledge of the treatment regimen, using the scoring systems of Cavanagh (1954) and Sprague et at. (1980). At selected posttreatment time intervals over a period of 18-21 days, subgroups of hens are euthanized by anesthesia with an intravenous injection of 5.0% chloralose in 5.0% sodium borate followed by exsanguination by cardiac puncture. At dissection, sections of brain, spinal cord, and sciatic nerve are either fixed in 4.0% neutral-buffered formaldehyde prior to histological preparation and staining or frozen at -20°C in tightly sealed glass bottles for enzymatic analysis of acetylcholinesterase (AChE) and neuropathic target esterase (NTE). The morphological, physiological, and biochemical changes caused by an OPIDP-inducing organophosphorus ester are beyond the scope of this chapter and the reader is referred to the relevant chapters by Ehrich and Jortner (Volume 2, Chapter 52), Johnson (Volume 2, Chapter 50), and Wilson (Volume 2, Chapter 51). The classic, positive, morphological effects are a typical, dying back, Wallerian peripheral and central axonal degeneration with a concomitant myelinopathy (AbouDonia, 1981). The birds will show ataxia, a high stepping gait, toe dragging, an inability to balance even with the aid of the wings, and, in the most severe state, an inability to move, al-
10.2 Testing Strategies
ways remaining in a squatting position. Biochemically, an initial inhibition of nervous tissue AChE will be observed, with a slow recovery over the next 18 days. In contrast, inhibition of NTE (if the agent has the capability to do so) will be observed up to 24 hours posttreatment, but quickly recovers to normal levels of activity within 2 or 3 days. The main drawback of this test of neurotoxic potential is that it behaves properly only with a few organophosphorus ester insecticides, none of which, on the basis of structure-activity relationships, would ever be marketed. Theoretically, all phosphate esters and those phosphorothioate esters that can be metabolically desulfurated to form phosphate analogs have the potential to cause the foregoing effects. The fact that they do not do so is perhaps due to (1) high toxicity and lethality, (2) rapid detoxification, or (3) slow desulfuration to the phosphate, all of which suggest that insufficient levels of the toxic agent were acquired in vivo to elicit the effects. However, this does not mean that other classes of insecticides (organochlorines, carbamates, and pyrethroid esters) and, indeed, other classes of pesticides do not cause neurotoxicity. They do, but are quite different in their actions, a topic that was discussed in depth by Ecobichon and Joy (1994). When some of these agents are in formulations, some thought must be given to effects caused by co-solvents, emulsifiers, or adjuvants, rather than by the pesticide. Completion of the test battery listed in the first-wave studies (Table 10.2) will provide adequate information concerning the acute toxicity of a test chemical in occupationally exposed individuals, in those accidentally exposed, and intentional, suicidal attempts, all situations that reflect high-level, short-duration exposures. Some indication of the relative potency of the unknown chemical in comparison to other agents in the same chemical class can be obtained as can possible mechanisms of action. A measure of the mutagenic potential can be obtained from a number of assay systems, and identified mutagenicity possibly can be confirmed by teratogenic effects in two sensitive animal models. These studies, completed at relatively low cost, are the basis upon which decisions will be made either to shelve the potential pesticide, halt further development, or proceed to the more expensive and time-consuming studies that comprise the second-wave list. 10.2.2 SECOND-WAVE STUDIES Considerably more humans are exposed to pesticides at relatively low levels over a much longer period, even a lifetime. The origins of such pesticide exposure are air- and water-borne remnants of agricultural or forestry operations or residues in foods above, at, or below tolerances established by regulatory agencies. The simulation of such exposures requires the development of other testing strategies, that is, short-term (subacute, subchronic) and long-term (chronic) studies (Ecobichon, 1997d). The endpoints of toxicity in such studies may differ considerably from those of acute intoxication and they include: 1. Reproductive studies in which target organ toxicity in the various phases of male and female reproduction are
293
explored. Many of these phases are as susceptible to chemical-induced damage and/or dysfunction as would be expected in the developing fetus. 2. Cancer, a threat to the well-being of all humanity and a "disease" that has been linked to pesticide residues as part of the "environmental contamination/exposure" scenario. A degree of flexibility must be maintained in these studies because since it is impossible to predict if and when toxicity will appear. Although anticipated target organ toxicity can be predicted from physiological properties of the test agent and experience with similar chemicals, serendipitous observations are the rule rather than the exception. Studies should be kept as open-ended as possible to permit the exploration of results, which necessitates the inclusion of sufficient numbers of animals at the "front end" to (1) demonstrate a dose-related toxicity without too high a mortality rate, (2) identify suspicious toxicological events, and (3) allow enough animals in each treatment group to study the possible permanence or reversibility of the toxicity. As shown in Table 10.9, there are specific goals to be achieved in addition to detecting toxicity. The experimental design of such studies has been discussed in detail by Arnold et al. (1990), and by Ecobichon (1997c), among others, and will only be summarized here. Subacute toxicity studies, usually of 2-4 weeks duration, are conducted as range-finding studies to choose dosage levels to be used in longer-term, subchronic (up to 90 days) and chronic (6 months to 2 years) studies. However, a recent examination of newly introduced OECD guidelines indicates that the subacute/subchronic terminology is being phased out; particular tests designate the duration of the study and leave no doubts in investigators' minds as to how long the study should last (e.g., OECD Guideline 407-Repeated Dose 28-Day Oral Toxicity Study in Rodents. There has also been a reduction in the duration of chronic studies from 2 years to 6 months for rodents, although chronic studies in dogs are still conducted for 2 years.
Table 10.9 Objectives to Be Achieved in Subchronic and Chronic Toxicity Studies 1.
Examine the biological nature of the toxic effects elicited from low dosages, monitoring a range of biological parameters
2.
Ascertain the variation in species response(s) to repeated exposure to the agent, looking for commonality of responses and/or distinct species differences
3.
Assess possible cumulative effects of the repeated exposure as body burdens of the agent are acquired with time
4.
Determine the nature of macroscopic and microscopic organ or tissue damage as it develops
5.
Identify the approximate dosage at which the altered physiological, biochemical, and morphological changes might occur
6.
Predict the long-range adverse health effects in the species arising from intermittent, repeated, or chronic exposure to the agent
294
CHAPTER 10
Toxicity Testing
10.2.2.1 Subchronic Studies
Such studies are designed to explore possible mechanisms of action in animals over a longer time period and at a lower range of doses than those reported to be lethal. Both sexes of the same age, strain, and species are used. At least three dosage levels (low, moderate, and high) as well as control (untreated or vehicle-treated) are included. The chosen dosage levels, generally derived from acute toxicity results, are usually fractional dosages, such as 1.0,5.0, and 10.0% of the LDso, and are administered via the route by which humans would be expected to acquire the agent, for example, in drinking water or in the food. To ensure that the entire dose is received, oral gavage with a feeding needle may be chosen. Dermal application may be required to satisfy occupational exposure concerns. Although a wide variety of physiological and biochemical parameters can be used to monitor the general well-being of the animals and to detect any adverse biological effect(s), it is imperative to determine when the first signs of toxicity appear at each dose level. A plan should be made to euthanize subgroups of each treatment group at 30, 60, and 90 days during treatment and carry out a full morphological and biochemical study to detect the appearance and the progression of a lesion rather than waiting for the treated animals to become ill. Frequently, a time- and a dose-effect relationship can be established. All dead and severely moribund animals, as well as those selected for euthanasia at fixed time points in the study, should undergo careful necropsy and examination. Tissues should be preserved for embedding, sectioning, staining, and microscopic examination. A large number of body organs are removed from each animal at necropsy for examination or are held for special study (Ecobichon, 1997d). Frequently, the initial tissue study involves only control and high dose animals' tissues to establish that there is a difference. Having done this, the low and moderate dose treated animals' tissues will be examined to establish a sUbjective dose-effect relationship. Available techniques allow microscopic quantification of tissue cell size and volume subcellular organelle volume, etc. Frequently seen cytologicai reactions to a chemical insult appear to adhere to a dose-effect relationship.
conducted with longer-lived species would require an adjustment (increase) in the length of the study if the same three-dose range was to be used. Such a study was a costly endeavour, considering that a 2-year study is only 20% of the life-span of a dog and 13% of the life-span of a monkey. The alternative approach, taken by regulatory agencies, is to adjust (increase) the dosage administered to longer-lived species so that the animals acquire a "life-span dosage" in 2 years. The pitfalls of this approach are obvious: Exposure levels are so high that the animals are unable to efficiently biotransform and excrete the test agent. Currently, most national and international regulatory agencies will accept a 6-month chronic study in rodents, but require a 24-month study in nonrodents and primates. A controversial point is the perspective that no new chronic toxicity has been detected in 24-month studies that was not seen in "appropriately designed" 6- or 12-month studies. All chronic studies, regardless of duration, suffer from the same design defect-an insufficient number of animals (usually 50 per sex per treatment group) to detect the effects associated with an agent that has a low incidence of toxicity in humans. An incidence rate of 1.0 in humans would require some 299 animals to be certain that the effect was not overlooked (Zbinden, 1973). An incidence rate of 0.1, would require some 2995 animals in the study. At best, the observations or lack thereof are "guesstimates" of toxicity/safety. As with the subchronic studies, groups of animals (both sexes, same strain and species) should receive one of a range of three dosages (low, moderate, and high) via a suitable route of exposure. These dosage levels should be chosen on the basis of the results of the subchronic studies (Ecobichon, 1997d). Once again, the investigator-selected subgroups should be euthanized at suitable intervals during the treatment period to detect the appearance and progression of lesions. The general well-being of the animals should be monitored by suitable parameters of
Table 10.10 Duration of Studies in Experimental Animals and Time Equivalents in the Human Duration of study (in months)
10.2.2.2 Chronic Studies
Species
A major challenge in designing long-term toxicologic experiments is to calibrate exposure levels to permit a reasonable, normal laboratory life (health, appearance, growth, and development) for the animals, while guaranteeing obvious evidence of chronic toxicity over and above that typically seen in aged animals (Huff et aI., 1991). As with the subchronic studies the objectives of chronic studies are to characterize the mech~ anism(s) by which an agent induces some toxic effect(s) when administered over a considerable portion of the life-span of the test animal. The greatest controversy centers on just how long that study should be. Originally, in regulatory parlance, a chronic mammalian study had a duration of 2 years, which represents the approximate life-span of a laboratory rodent (Table 10.10). Studies
Percent of life-span
6
3
24
12
Rat
4.1
12.0
2S.0
49.0
99.0
Rabbit
I.S
4.S
9.0
18.0
36.0
Dog
0.82
2.5
4.9
9.8
20.0
Pig
0.82
2.5
4.9
9.8
20.0
Monkey
0.55
1.6
3.3
6.6
13.0
Human equivalents (in months) Rat
34
101
202
404
808
Rabbit
12
36
72
145
289
Dog
6.S
20
40
81
162
Pig
6.5
20
40
81
162
Monkey
4.5
13
27
61
107
Source: Ecobichon (1997d).
10.2 Testing Strategies
blood chemistry and urinalysis. The required minimum number of animals should be expanded to allow for possibly high, toxic doses and (unexpected) observations of toxicity. In addition, sufficient numbers of test subjects permit a recovery period study following termination of treatment. Extensive morphological examination of a variety of tissues should be conducted (Ecobichon,1997d). 10.2.2.3 Carcinogenicity Studies The design of a carcinogenicity study, which usually is conducted in rodent species, must permit detection of one (or more) of the properties in the operational definition of a carcinogen (Table 10.11; Ecobichon, 1997c). Generally, carcinogenic chemicals cause an effect at a low incidence rate, creating difficulties with the numbers of animals that must be used costeffectively in the study. Prolonged exposure (in excess of 12 months) of the animals is required to demonstrate the carcinogenic potential of most test agents. The normal 6- or 12-month chronic toxicity study would not be of sufficient duration to demonstrate this properly. The exorbitant costs of running parallel chronic toxicity and carcinogenicity studies has resulted in their combination, where high exposure levels are used to identify chronic toxicity and additional groups of animals receive much lower daily doses of the test agent for a much longer time to detect carcinogenicity (Ecobichon, 1997b). Typically, carcinogenicity studies are of 24-month duration and the highest dose of agent selected will be one that neither produces clinical signs of toxicity nor affects the long-term health or the normal longevity of the animals. Appropriate designs for such studies have been discussed (Ecobichon, 1997b; Robens et aI., 1994). There has been considerable controversy over the selection of appropriate dosage levels for carcinogenicity studies. The dosage, defined in the previous paragraph as the highest level, is called the maximum tolerated dose (MTD) and is usually chosen as a consequence of the results obtained in subchronic studies. It elicits no adverse biological effects other than cancer. Normally, two additional dose levels, lower than the MTD, are used in an attempt to establish a dose-related incidence of tumors, decreased latency period, etc. Butterworth et al. (1991) provided a rational discussion of this controversy. Along with the routine biochemical monitoring of general health, particular attention is focused at necropsy on detection of "lumps and bumps" and, at later microscopic examination, on the detection and identification of cell hyperplasia, preneoTable 10.11 Operational Definition of a Carcinogen
An agent that has the ability to induce tumors as evidenced by: I.
An increased incidence of tumor types found in controls
2.
Occurrence of tumors earlier than in controls
3.
Development of tumor types not seen in controls
4.
An increased multiplicity of tumors in individual animals
Source: Ecobichon (1997b).
295
plastic nodules, and tumors (benign and malignant) to satisfy the criteria listed in Table 10.11 for a carcinogen. 10.2.2.4 Short-Term Carcinogenicity Studies The discovery and promotion of several strains of mice and rats that have predispositions toward developing high incidences of certain tumor types has been of considerable assistance in short-term studies for carcinogenicity. Chemical exposure for 90-120 days is frequently sufficient to elicit an increased incidence and/or earlier appearance of tumors. Although this type of testing does not replace the full-scale studies, these animal models can be used to screen chemicals at considerably lower cost, which saves funds for application to extended studies of those chemicals that do cause positive results. 10.2.2.5 Reproductive Studies Reproductive toxicology is the study of the occurrence, causes, manifestations, and sequelae of the adverse effects of exogenous agents on reproduction (John son, 1986). Reproductive hazards encompass adverse health effects to the prospective mother and father (loss of libido, infertility, sterility) as well as to the developing offspring (abortion, fetal and/or perinatal death, teratogenesis). Many of these events are considered to be associated with cellular mutations. Although mutagenicity and teratogenicity were assessed by the test protocols in the first-wave studies, toxicological evaluation of any chemical must encompass the entire breadth of the life cycle. To date, there are no tests to examine the direct effect(s) of the agent on gametogenesis in either adult male or female mammals. The gametes (spermatozoa, ova) as well as the fertilized ovum, and the pre- and postimplantation embryo are exquisitely susceptible to physical and chemical insult, primarily because of the frequency of cell division and replication of cellular DNA in repeated mitotic and meiotic division. These activities are sensitive to exceptionally low concentrations of agents. In the past, most emphasis was placed on studies that examined the effects of agents on the female in her role as the vehicle for the susceptible, developing embryo. Essentially, these studies are the Segment 11 or teratogenic studies already discussed. More recently, attention has focused on pretreating either male or female animals before mating to examine any effects on gametes, because any damage to the gamete may play a vital role in the viability of the developing embryo. Experimental paradigms have been described to examine the mechanisms of toxicity in either or both of the sexes (Ecobichon, 1997c; Manson and Kang, 1994; Zenick et aI., 1994). In a classical Segment I (fertility, reproduction) study, groups of either male or female test animals are treated prior to mating with a range of dosages (usually three levels) for one gametogenic cycle (60-days for the male or 15 days for the female). If the treated female animals are mated with normal, untreated males, the pregnant females continue to receive the same daily treatment for either the duration of the pregnancy or beyond parturition if there is concern about effects in the lacta-
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Toxicity Testing
tional time period. If the postnatal development period is being studied, the pregnant animal will be allowed to give birth and to rear the young past the weaning stage, thereby permitting study of postnatal viability and behavioral/cognitive development. The pretreated male animals are mated with healthy, untreated, virgin females throughout the next gametogenic cycle. Essentially, a group of new females (n = 4 or 5) are placed with each male every 5 days. Those females that are confirmed to be pregnant are removed and placed in individual cages until the day before parturition, when they are euthanized so that the effects on the developing embryo can be studied. The focus of the study is on whether the test agent can produce an effect on spermatogenesis, which is expressed as alterations in a number of indices (i.e., fertility, mating, fecundity, gestational time, live births, dead fetuses) that are calculated at the termination of the pregnancy (Ecobichon, 1997c). In this bioassay, the reproductive biology of the female is assumed to be normal; only that of the male is presumed to be affected. A 60-day treatment with an active mutagenic agent might produce disastrous effects on spermatogenesis at several different stages, although the results are unclear except in the cases of sterility or reduced fecundity (Ecobichon, 1997c). A variant of the previously discussed test is the dominant lethal test in which selected doses of test agent are administered either as a pulse dose or single dose for a short term (5 to 7 days) to adult, breeding-age males. The dosage(s) selected is estimated to cause severe chromosomal damage or lethal germ-cell mutations, resulting in fetallethality. Whereas gamete production is a continuous synthesis from stem cells in the testes, a pulse dose may elicit limited damage at one or only a few stages. A breeding program is established where, for the next cycle after treatment, females are bred with the treated male every 4-5 days. These females are then followed to the end of their pregnancies and the fetuses are examined on the day before parturition. Some 50 females will be bred with each male over the 60-day period following the pulse dose of test agent. The main advantage of this assay is the ability to test the chemical sensitivity of the germ cells in vivo at different (premeiotic, meiotic, postmeiotic) stages of spermatogenesis. The disadvantage is that if the chemical is only a weak mutagen, it will not produce a significant amount of chromosomal damage, thereby "escaping" the screening assay (Ecobichon, 1997c). The classical reproductive, two-generation study has, in the past, been one in which groups of young, healthy, female animals (Fo generation) are treated with the test agent (three different exposure levels) from the time of conception (mated with normal males), throughout the gestation period, the rearing of the postnatal progeny (F lA generation), and at least through one repeated breeding cycle (FIB generation) (Ecobichon, 1997c). The FIA progeny are euthanized for gross and microscopic study and, following a 2-week rest period, the female is bred again to produce an FIB generation, the female offspring of which are the source of animals for the next generation. The selected FIB females receive the same level of exposure to the test agent as their dams from weaning throughout their breeding cycle to produce F2 generation pups. Normally, a two-generation
study is conducted under regulatory requirements, but the study can be extended to any number of generations, keeping in mind that the progeny receive the test agent transplacentally and via the milk until weaned, at which time they must begin to receive the agent via their food or drinking water. It must be emphasized that in both teratological and reproductive studies, the toxicity observed may be related to maternal and/or nutritional toxicity, especially if high concentrations of a rather nontoxic chemical are being administered via the diet or the test agent has an unpleasant odor or taste. In such scenarios, the female may reject the food, thus depriving the developing fetuses of nutrients essential for normal growth and development or affecting fertility, fecundity, and/or the ability of the female to maintain the pregnancy. All too frequently, adverse reproductive outcomes are falsely attributed to the test chemical, when, in effect, the defect is nutrition related.
REFERENCES Abou-Donia, M. B. (1981). Organophosphorus ester-induced delayed neurotoxicity. Am. Rev. Pharmacal. Taxical. 21, 511-548. Abou-Donia, M. B., and Lapadula, D. M. (1990). Mechanism of organophosphorus ester-induced delayed neurotoxicity: Type I and Type n. Ann. Rev. Pharmacal. Taxical. 30, 405-440. Ames, B. N., Durston, W. E., Yamasaki, E., and Lee, E. D. (1973). Carcinogens are mutagens: A simple test system combining liver homogenates for activation and bacteria for detection. Proc. Natl. Acad. Sci. U.S.A. 70, 22812285. Ames, B. N., McCann, J., and Yamasaki, E. (1975). Methods for detecting carcinogens and mutagens with the salmonella/mammalian microsome mutagenicity test. Mutat. Res. 31, 347-364. Arnold, D. L., Grice, H. C., and Krewski, D. R., eds. (1990). "Handbook of In Vivo Toxicity Testing." Academic Press, San Diego. Basketter, D. A., Scholes, E. W., and Kimber, 1. (1994). The performance of the local lymph node assay with chemicals identified as contact allergens in the human maximization test. Faad Chem. Taxical. 32,543-547. Botham, P. A., Basketter, D. A., Maurer, T., Mueller, D., Potokar, M., and Bontinck, W. J. (1991). Skin sensitization-A critical review of predictive test methods in animals and man. Faad Chem. Taxicol. 29, 275-286. Bruner, L. H., Parker, R. D., and Bruce, R. D. (1992). Reducing the number of rabbits in the low-volume eye test. Fundam. Appl. Taxica!. 19, 330-335. Buehler, E.V. (1964). A new method for detecting potential sensitizers using the guinea pig. Taxical. Appl. Pharmacal. 6, 341. Butterworth, B. E., Goldsworthy, T. L., Popp, J. A., and McClellan, R. O. (1991). The rodent cancer test: An assay under seige. ClIT Activities 11, 1-8. Cavanagh, J. B. (1954). The toxic effects of tri-a-cresylphosphate on the nervous system: An experimental study in the hen. J. Neural. Neurosurg. Psychiatry 17, 163-172. Chan, P. K., and Hayes, A. W. (1994). Acute toxicity and eye irritancy. In "Principles and Methods of Toxicology" (A. W. Hayes, ed.), 3rd ed., pp. 579648. Raven Press, New York. Conquest, P., Durand, G., Laillier, J., and Blazonnet, B. (1977). Evaluation of ocular irritation in the rabbit: Objective vs subjective assessment. Taxicol. Appl. Pharmacal. 39, 129-139. Copplestone, J. F. (1988). The development of the WHO Recommended Classification of Pesticides by hazard. Bull. WHO 66, 545-551. Cranmer, J., and Hixon, J., eds. (1984). "Delayed Neurotoxicity." Intox Press, Little Rock, AR. De Sousa, D. J., Rouse, A. A., and Smolon, W. J. (1984). Statistical consequences of reducing the numbers of rabbits utilized in eye irritation testing: Data on 67 petrochemicals. Taxical. Appl. Pharmacal. 76, 234-242.
References
Draize, J. N., Woodward, G., and Calvery, H. O. (1944). Methods for the study of irritation and toxicity of substances applied topically to the skin and the mucous membranes. J. Pharmacol. Exp. Ther. 82, 377-389. Durham, H. D., and Ecobichon, D. J. (1984). The function of motor nerves innervating slow tonic skeletal muscle in hens with delayed neuropathy induced by TOTP. Can. 1. Physiol. Pharmacol. 62, 1268-1273. Durham, H. D., and Ecobichon, D. J. (1986). An assessment of the neurotoxic potential offenitrothion in the hen. Toxicology 41, 319-332. Ecobichon, D. J. (1996). Toxic effects of pesticides. In "Casarett and Doull's Toxicology. The Basic Science of Poisons," 5th ed., (c. D. Klaassen, ed.), Chap. 22, pp. 643-689. McGraw-Hill, New York. Ecobichon, D. J. (1997a). Acute toxicity studies. In "The Basis of Toxicity Testing," 2nd ed., Chap. 3, pp. 43-86. CRC Press, Boca Raton, FL. Ecobichon, D. J. (1997b). Mutagenesis-{:arcinogenesis. In "The Basis of Toxicity Testing," 2nd ed., Chap. 6, pp. 157-190. CRC Press, Boca Raton, FL. Ecobichon, D. J. (1997c). Reproductive toxicology. In "The Basis of Toxicity Testing," 2nd ed., Chap. 5, pp. 117-156. CRC Press, Boca Raton, FL. Ecobichon, D. J. (1997d). Subchronic and chronic studies. In "The Basis of Toxicity Testing," 2nd ed., Chap. 4, pp. 87-116. CRC Press, Boca Raton, FL. Ecobichon, D. J., and Joy, R. M. (1994). "Pesticides and Neurological Diseases," 2nd ed., CRC Press, Boca Raton, FL. European Commission Directive 83/4671EEC (1983). Freeberg, F. E., Nixon, G. A., Reer, P. J., Weaver, J. E., Bruce, R. D., Griffith, J. F., and Sanders, Ill, L. W (1986). Human and rabbit eye responses to chemical insult. Fundam. Appl. Toxicol. 7, 626-634. Gad, S. c., Dunn, B. J., Dobbs, D. W, Reilly, c., and Walsh, R. D. (1986). Development and validation of an alternative dennal sensitization test: the mouse ear swelling test (MEST). Toxical. Appl. Pharmacol. 84, 93-114. Hayes, Jr., W J. (1975). "Toxicology of Pesticides," p. I. Williams and Wilkins, Baltimore. Huff, J., Haseman, J., and Rail, D. (1991). Scientific concepts, values and significance of chemical carcinogenesis studies. Ann. Rev. Pharmacal. Toxicol. 31,621-652. Johnson, E. M. (1986). Perspectives on reproductive and developmental toxicology. Toxicol. Indus. Health 2, 453-482. Kimber, I., and Basketter, D. A. (1992). The murine local lymph node assay: A commentary on collaborative studies and new directions. Food Chem. Toxicol.30, 165-169.
297
Manson, J. M., and Kang, Y. J. (1994). Test methods for assessing female reproductive and developmental toxicology. In "Principles and Methods of Toxicology," 3rd ed. (A. W Hayes, ed.), pp. 989-1038. Raven Press, New York. McCreesh, A. H., and Steinberg, M. (1983). Skin irritation testing in animals. In "Dennatotoxicology," 2nd ed., (F. N. Marzulli and H. I. Maibach, eds.), pp. 147-166. Hemisphere Publishing, New York. National Research Council, Toxicity Testing, Strategies to Detennine Needs and Priorities, National Academy Press, Washington, DC (1984). Patrick, E., and Maibach, H. I. (1994). Dennatotoxicity. In "Principles and Methods of Toxicology," 3rd ed., (A. W Hayes, ed.), pp. 767-803. Raven Press, New York. Robens, J. F., Calabrese, E. J., Piegorsch, W. W., Schuler, R. L., and Hayes, A. W (1994). Principles of testing for carcinogenicity. In "Principles and Methods of Toxicology," 3rd ed., (A. W Hayes, ed.), pp. 697-728. Raven Press, New York. Schardein, J. L. (1985). "Chemically Induced Birth Defects." Dekker, New York. Seabaugh, V. M. (1994). EPA's requirements for dennal irritation and sensitization testing. Food Chem. Toxicol. 32, 93-95. Slott, v., and Ecobichon, D. J. (1984). An acute and subacute neurotoxicity assessment of trichlorfon. Can. J. Physiol. Pharmacol. 62, 513-518. Sprague, G. L., Sandvik, L. L., Bickford, A. A., and Castles, T. R. (1980). Evaluation of a sensitive grading system for assessing acute and subchronic delayed neurotoxicity in hens. Life Sci. 27, 2523-2528. Talsma, D. M., Leach, C. L., Hatoum, N. S., Gibbons, R. D., Roger, J. c., and Garvin, P. J. (1988). Reducing the number ofrabbits in the Draize eye irritancy test: A statistical analysis of 155 studies conducted over 6 years. Fundam. Appl. Toxicol. 10, 146-153. U.S. Code of Federal Regulations, Title 40, Part 158 (40CFR Part 158). U.S. Environmental Protection Agency (EPA) (1978). Acute delayed neurotoxicity study and subchronic neurotoxicity studies. Fed. Register 43, 37,36237,363,37,374-37,375. Zbinden, G. (1973). "Progress in Toxicology. Special Topics," Vol. I, pp. 1719. Springer-Verlag, New York. Zenick, H., Clegg, E. D., Perreault, S. D., Klinefelter, G. R., and Gray, L. E. (1994). Assessment of male reproductive toxicity: A risk assessment approach. In "Principles and Methods of Toxicology," 3rd ed., (A. W. Hayes, ed.), pp. 937-988. Raven Press, New York.
CHAPTER
11 Regulatory Evaluation of the Skin Effects of Pesticides MichaelO'Malley University of California, Davis
11.1 INTRODUCTION 11.1.1 BASIC PATTERNS OF SKIN REACTION Clinical effects of pesticides on the skin include both systemic and topical reactions. Systemic effects, such as urticaria, chloracne, and porphyria cutanea tarda, may occur following ingestion, inhalation, or topical exposure. Direct topical effects include acute irritation and corrosion, subacute (gradual-onset) irritation, and delayed-onset allergies. Any of the injuries described above may damage the pigment-producing basal layer of the skin, resulting in either an increase or a decrease in epidermal melanin production. Typically, both injuries and residual effects occur in a pattern that coincides with the site of contact. Depending upon the time interval between exposure and the onset of lesions, recognizing the source of the skin injury may be simple or complex. Protocol for clinical patch testing
In the regulatory arena, evaluation of the capacity of individual compounds to cause irritant or allergic reactions depends upon animal testing as well as analysis of human use experience. Protocols for evaluating allergy and irritation in experimental animals are discussed.
11.1.2 TESTING REQUIREMENTS AND TEST PROTOCOLS The requirements for skin testing of pesticides for the purposes of federal and state registration vary with the government jurisdiction. In the United States, primary dermal irritation studies and sensitization studies are required for each manufacturinguse product and each end-use product. The tests performed in this manner are considered part of the regulatory database and are not available in the public literature. The irritation testing requirements are similar to the standard Draize tests (Inset 2). Dermal sensitization tests may be done according to one of several standard protocols (Inset 3).
Application of previously identified nonirritating concentration of test substance for 48 hours, followed by removal of patch and initial reading. Follow up reading at 96 hours. Simplified scoring system for grading patch tests: O--no visible reaction 1+ -erythema 2+ -erythema and blistering 3+ -necrotic reaction
Dermal irritation tests
Distinguishing between allergic and irritant effects is a primary goal of both clinical and regulatory evaluation of the skin effect of pesticides. Clinically, irritant reactions tend to develop within a short time of exposure, whereas skin allergies are typically delayed in onset. Exceptions to this simple rule occur-some irritant reactions are cumulative and some allergic reactions occur within minutes of contact with the offending allergen (urticaria). A summary of the clinical protocol for provocation tests (or patch tests) is shown (Inset 1). Handbook of Pesticide Toxicology
Volume 1. Principles
Irritation test using albino rabbits (Draize test) A single dose of the technical material with detailed characterization of contaminants, or an end-use product, is applied to the skin of one or several experimental animals (depending upon the possibility of a corrosive reaction) for 4 hours. The irritation is scored at intervals until the irritation has resolved or is considered permanent. Based upon scores at 72 hours and persistence of irritation for more than 14 days, materials are categorized as corrosive (category 1-72 hr dermal irritation score >7), severe irritants (category II-72 hr dermal irritation score 5-7), moderate irritants (category III-72 hr dermal irritation score 2-5), or minimally irritant (category IV-72 hr dermal irritation score 0-2) -may be divided into compounds that produce no irritation (nonirritants) and those that produce transient, mild irritation.
A newer technique, not yet approved by the EPA, but probably more reproducible than any of the currently approved methods for testing dermal sensitization, is the regional lymph
299
Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
300
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides
Allergy testing protocols Sensitization test protocols Following initial exposure to a test substance, tbe animals are challenged, to establish whether the animal has developed hypersensitivity. This is evaluated by comparing scores during the induction period with those during the challenge period and witb those of control animals who received the challenge witbout initial exposure. For ambiguous results on challenge, a rechallenge phase is used. Epicutaneous test methods: Buehler test-A closed patch is applied for six hours, weekly, during a 3-week induction period; tbe test concentration for induction is chosen to be approximately lO-fold higher than the expected human exposure concentration. Challenge, with a nonirritating concentration, takes place weeks 5,6,7. Open epicutaneous test-After establishing the concentration that produces minimal irritation and no irritation threshold, induction is begun at the latter concentration. Applications are repeated daily for tbree weeks or five times weekly for four weeks, always on the same skin site. The challenge is conducted on day 21 using the minimal irritant and some lower concentrations-skin reactions are read after 24, 48, and/or 72 hours. Rechallenge, if necessary, is done on day 35. Intradermal methods: Guinea-pig maximization test (GPMT)-Induction is begun on day 0 with 0.1 ml test material intradermal (by injection) togetber with 0.1 cc Freund's Complete Adjuvant (FCA). Control animals receive only tbe injection of 0.1 ml FCA. On day 7 induction is boosted by occluding the test material against the skin for 48 hours. On day 21 tbe challenge is performed on a shaved 4-cm2 area on the left flank, using a nonirritating concentration of the test material. Controls are treated with occluded vehicle only. If challenge reactions are ambiguous, animals are rechallenged on day 28. Freund's complete adjuvant (FCA) test-Induction day 1,5,9: 0.1 ml test material in FCA in shoulder of animals of the control group treated witb FCA only. Day 21 + 36 challenge and rechallenge: A: a minimum irritating concentration; B: max nonirritating concentration both tested in test animals and controls. Days 22-24; 36-38 skin sites read 24, 28, 72 h after challenge and rechallenge. The test is simple to perform and involves low material and operational expenses.
node assay (Ashby et aI., 1995; Ikarashi et aI., 1994, 1996). The test involves use of a 3-day induction period, followed by monitoring of the uptake of H3 -labeled thymidine in regional lymph nodes (excised and placed in cell culture after sensitization) as a marker of sensitization. In addition to H3 -labeled thymidine, sensitization can be monitored using cell number or levels of interleukin-2 produced in cell culture (Hatao et aI., 1995). The results compare favorably to the more cumbersome, 3- to 7-week-long in vivo assays (Ikarashi et aI., 1994), but pesticides have not been systematically studied with this technique. For interested readers, additional details of predictive skin testing procedures have recently been described by Bashir and Maibach (Bashir and Maibach, 2000).
11.1.3 REGULATORY DECISIONS The principal regulatory decision dependent on the results of preregistration animal testing is the content of precautionary statements on the pesticide product label. Materials showing corrosive effects or reversible skin irritation are labeled as such.
Labels for products not found to be corrosive or irritating generally carry statements advising the user to minimize the degree of skin contact. Materials judged as sensitizers in animal tests or reported as sensitizers in the public domain scientific literature are required to indicate the possibility of skin sensitization on the product label.
11.1.4 USE OF ILLNESS-SURVEILLANCE DATA WITH EXPERIMENTAL DERMAL IRRITATION AND SENSITIZATION TESTS Apart from tests required for registration, additional regulatory information is obtained from postregistration surveillance of pesticide illness reports and from public literature reports on adverse skin effects of pesticides. A summary of information on irritation, sensitization, and postregistration illness-surveillance information derived from California Department of Pesticide Regulation (CDPR) data for 178 active ingredients is shown in Table 11.1. The data are discussed below by use and structural category. The sensitization data are based on a review of study summary memoranda for 427 formulations or active ingredients reviewed between January, 1989 and July, 1997. The irritation data derive from a review of the publicly available product label database.' Probable or definite skin illness or injury cases involving the same products are also shown in the table. Because the pesticide illness database (1982-1995) contained more than 10,000 reports of skin injury or illness from both agricultural and nonagricultural use, the review was limited to 991 cases involving probable or definite skin effects to pesticide applicators from a single active ingredient. None of the 991 cases had patch testing performed. For purposes of this review, these cases are referred to in the text below as the "pesticide handler database" or "handler database." Further details regarding the cases in the database are included in a recent review (O'Malley, 2000).
11.2 REVIEW OF USE CATEGORIES 11.2.1 ANTIMICROBIAL AGENTS Many of the antimicrobial agents registered as pesticides are corrosive or markedly irritant. This biological property may be correlated with underlying chemical reactivity linked in many cases to an ability to provoke sensitization.
11.2.1.1 Isothiazolins (Kathon® Compounds) In animal irritation studies, several of the isothiazolin compounds (octhilinone, 2682-20-4, 82633-79-2, 26172-55-4, 226530-20-1) are acutely irritant or corrosive in their concentrated forms. Although no dermal sensitization studies were reviewed for these compounds, all are labeled as sensitizers because of numerous reports in the public literature documenting their capacity to sensitize (Bruze and Gruvberger, 1988; I http://www.cdpr.ca.gov.
under database resources.
11.2 Review of Use Categories
301
Table 11.1
Compound
CAS #
Predictive tests in animals
Data from pesticide handler data base 1982-1995
Draize irritation test
# of cases
Sensitivity
Case examples
Antimicrobials
Isothiazolin (Kathon®) compounds 1,2-benzisothiazolin3-one
2634-33-5
Corrosive
Sensitizer per public domain literature
2-methyl-4-isothiazolin3-one
2682-20-4
Labeled as a corrosive
Sensitizer per public domain literature
5-chloro-2-methyl-4isothiazilone-3-one and 2-methyl-4,5-trimethylene4-isothiazolin-3-one
82633-79-2 26172-55-4
Corrosive
Sensitizer per public domain literature
2-n-octyl-4-isothiazolin3-one
26530-20-1
Labeled as a corrosive
Sensitizer per public domain literature
13
89-1312: A worker sanitizing cooling towers with a Kathon® compound accidentally spilled some of the solution on himself and suffered a chemical bum. 90-552: An employee adding concentrate (of Kathon®) to washing solution spilled some of the material onto trouser leg of work pants, causing a chemical bum.
Quarternary ammonium compounds alkyl dimethyl benzyl ammonium chloride (multiple compounds alkyl groups = cl4,cl6 etc.)
112-18-9
Corrosive at 50% concentration
Nonsensitizer in modified Maguire method (variant of the FCA test)
dioctyl dimethyl ammonium chloride
5538-94-3
Corrosive at 50% concentration
Sensitizer by Buehler test
trimethyl ammonium chloride (didecyl dimethyl ammonium chloride)
7173-51-5
Corrosive
Nonsensitizer in Buehler test; sensitizer per public domain literature
107
88-1893: A worker splashed a sanitizer containing quaternary ammonium compounds onto his face. Four hours after exposure, he developed 6-10 macular lesions at the site where the material splashed on him.
Chlorine compounds and chlorine stabilizer---Cases reported represent irritant reactions in end-users of sanitizers/disinfectants cyanuric acid
108-80-5
No data
No data
16
hexahydro-l,3,5-triethyls-triazine
108-74-7
Corrosive
Weak sensitizer by Buehler test
0
sodium hypochlorite
7681-52-9
Labeled as an irritant
Sensitizer per public domain literature
132
91-2327: A custodial employee splashed material on her right arm while cleaning a toilet and subsequently developed itchy, red, and swollen area at the site of contact.
87-1468: A pet store employee developed a severe, painful rash on her hands from using a 12.5% sodium hypochlorite product to clean kennels without wearing gloves. The product proved to be a swimming pool sanitizer used in violation of the product label.
Phenolic compounds: 128 phenolic compounds registered for use as disinfectants; typical examples include ortho-phenylphenol, p-tert-butyl phenol o-phenylphenol
90-43-7
11 % formulation is a category II irritant
No data
p-tert-butylphenol
98-54-4
No data
Sensitizer per public domain literature
Phenol
108-95-2
Undiluted phenol is an irritant
No data
tributyltin oxide
56-35-9
Labeled as a minimal irritant
No data except on mixtures
tributyltin methacrylate and tributyltin fluoride
91745-52-7
Corrosive
Equivocal reaction in Buehler test
29
88-909: A hospital janitor got disinfectant on the hand through a hole in disposable glove, caused burning of the skin, diagnosed as irritant contact dermatitis.
Organotins 90-728: An employee added a mildewcide to a can of paint. When she pounded the lid back on the paint can, some of the material splashed on her, resulting in a rash on the face and neck. 0
(continues)
302
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides
Table 11.1 (continued)
Compound
CAS #
Predictive tests in animals
Data from pesticide handler data base 1982-1995
Draize irritation test
Sensitivity
# of cases
Case examples
Miscellaneous
2,2-dibromo-3-nitrilpropionamide
10222-01-2
Category II irritant
Inadequate data
0
2-(hydroxymethylamino)ethanol
34375-28-5
Labeled as a corrosive
N onsensitizer
0
Bronopol 2-bromo-2-nitro-1 ,3propanediol
52-51-7
Labeled as a corrosive
Sensitizer per public literature
0
1,2-dibromo-2,4-dicyano butane
35691-65-7
Labeled as an irritant
Sensitizer per public domain literature
0
iodine
7553-56-2
Technical material corrosive
Nonsensitizer
acephate
30560-19-1
Technical material causes transient irritation
Nonsensitizer
3
89-2500: An applicator developed a rash, described as urticaria and contact dermatitis, on both arms soon after acephate. The symptoms disappeared soon after he showered, but reappeared when he next applied acephate. 83-2409: An applicator spraying acephate on trees was exposed to liquid insecticide soaking through his clothes from a leaking fitting which allowed material to soak through his clothes; developed contact dermatitis. 86-1084: A structural pest control operator was trying to attach a crack and crevice injector to a spray can containing acephate. He sprayed his face and hands, and contaminated the respirator he was wearing and subsequently developed erythemaous papules and vesicles on forearms, hands, and ears.
chlorpyrifos
2921-88-2
Technical material causes transient irritation
41 % agricultural formulation labeled as a sensitizer; animal studies negative
11
diazinon
333-41-5
Minimal irritant
Sensitizer
5
85-343: An applicator treating a large carpet area for fleas came using a backpack, noticed that some of the spray material (chlorpyrifos) had leaked, soaking his lower back and upper legs. He had burning at the site of contact, but did not develop overt dermatitis. 87-2537: An applicator developed a rash on his arms and chest after a hose ruptured. He changed his shirt, but did not shower.
dichlorvos
62-73-7
Labeled as a ninimal irritant
No data
dimethoate
60-51-5
Technical material causes transient, minimal irritation
Nonsensitizer
2
83-1880: Even though protective clothing was worn, applicator developed a rash after application of dimethoate on grapes. He also had nonspecific symptoms of systemic poisoning (nausea and headache).
malathion
121-75-5
Technical material causes transient, minimal irritation
Nonsensitizer in animal studies: 20% product labeled as sensitizer
5
94-401: An employee of a small central valley city was pumping up a spray tank containing malathion when a hose coupling broke, spraying the material on his face and neck. Despite washing immediately, he developed a mild erythema in the exposed areas.
methamidophos
10265-92-6
Labeled as minimal irritant
Nonsensitizer in Buehler test
methidathion
950-37-8
25% formulation nonirritant
Nonsensitizer
87-2288: While disinfecting with an iodine product, an employee developed pruritic rash on arm.
Insecticides Organophosphates
83-1870: An apartment manager was spraying dichlorvos for roach control when the hose broke on his hand sprayer, splashing material on his right arm and left ear. On exam he had severe dermatitis of the right elbow and forearm and left ear, complicated by a possible secondary infection.
84-129: A mixerlloader splashed mixture of methamidophos and buffer on himself while transferring material and developed blisters in the exposed area. 0
(continues)
11.2 Review of Use Categories
303
Table 11.1 (continued)
Predictive tests in animals
Data from pesticide handler data base 1982-1995
Draize irritation test
Sensitivity
# of cases
2
Compound
CAS #
naled
300-76-5
Labeled as a corrosive
Sensitizer
oxydemeton-methyl
302-12-2
53% formulation causes transient irritation
Labeled as sensitizer
83-304: A worker was cleaning the tip of a spray rig when
70% formulation causes transient irritation
Nonsensitizer in Buehler test
86-318: A San Diego pet shop employee developed a rash on
3689-24-5
Labeled as a minimal irritant
Nonsensitizer in Buehler test
22781-23-3
Labeled as a minimal irritant
Nonsensitizer in Buehler test
phosmet
sulfotep
732-11-6
Case examples 88-942: A worker hand poured naled (Dibrom®) for application on strawberries, without wearing rubber boots, gloves, respirator, or eye protection. After spilling the material on his leather boots, he wore them the rest of the day. He developed severe blister on foot, which did not improve with home treatment and eventually required medical attention. 88-2330: A worker was using naled for fly control, when a hose broke, spraying him in the face. He developed a rash on the ears despite wearing coveralls, gloves, respirator, and goggles. At the time of treatment, he was noted to have a chemical contact dermatitis with a secondary infection.
oxydemeton-methyl (Metasystox®) splashed into face underneath the shield he was wearing. He developed erythema at the site of contact. her hands after she began using a phosmet flea dip.
o
Carbamate bendiocarb
82-1278: While treating for cockroaches with bendiocarb, a
hotel employee got his fingers into the material. He later stuck his fingers in his mouth, causing a condition described as a mild allergic reaction to the lips and tongue.
carbaryl
63-25-2
Technical material nonirritant
Nonsensitizer in Buehler test
2
fenoxycarb
72490-01-8
Minimal irritant effects ascertained from 21-day dermal study
Nonsensitizer in Buehler test
o
methiocarb
2032-65-7
75% formulation nonirritant
o
methomyl
16752-77-5
Technical material nonirritant
Nonsensitizer in Buehler test Nonsensitizer in Buehler test
propoxur
114-26-1
Technical material nonirritant
Nonsensitizer in Buehler test
4
121-21-1
57% technical material causes transient irritation
Positive Buehler tests for some formulations
10
2
82-2634: A turkey farm employee developed dermatitis after a hose broke during an application of carbaryl. 82-2703: An applicator applying carbaryl dust developed dermatitis after getting the material on his hands and arms.
84-1512: While spraying methomyl on corn, an applicator cleaned clogged nozzles on his equipment with his bare hands and developed a bad rash. 88-297: A structural pest control worker was spraying propoxur, and wiped his hands on shirt. He then developed rash on chest where he wiped his hands.
Pyrethrins
pyrethrins
90-2621: A fairgrounds employee suffered chemical bum to
right leg while applying a pyrethrin insecticide to livestock bams. "The fogger" machine he was using had a loose cap on the reservoir tank causing insecticide concentrate to come in contact with his leg. 86-385: A kitchen employee set off a fogger and remained in the treated area for 10 minutes in violation of and thus was cited (nov) for conflict with labeling "leave area immediately after setting off fogger." He developed skin irritation on his face. (continues)
304
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides
Table 11.1 (continued)
Compound
CAS #
Predictive tests in animals
Data from pesticide handler data base 1982-1995
Draize irritation test
# of Sensitivity
cases
Case examples
Pyrethrin synergists piperonyl butoxide
51-03-6
92% formulation category II irritant
Inadequate-studies on mixtures with pyrethrins
o
Listed in registry data only as a component of mixtures.
n-octylbicycloheptenedicarboximide
113-48-4
20% concentration, mixed with pyrethrins causes transient irritation
Studies with mixtures only
o
Listed in registry data only as a component of mixtures.
allethrin
584-79-2 42534-61-2 (cis/trans allethrin)
Technical material causes transient irritation
Nonsensitizer
bifenthrin
82657-04-3
Technical material nonirritant
Nonsensitizer
2
94-1052: An applicator bumped his right arm against a spray nozzle and got some of the bifenthrin spray solution on the arm and then developed numbness and tingling in the right arm. 95-1210: A mixerlloader, employed by a professional agricultural pest control company to treat cotton, splashed bifenthrin on his arms, face and eyes while transferring product from a closed system holding tank into a I-gallon container. The container overfilled and the pressure created forced the product out. He develped redness, and a burning sensation on the face, chest, and shoulder.
permethrin
52645-53-1
Technical material category III irritant
sensitizer
3
92-1381: Worker was mixing material and small amounts kept getting under gloves and shirt. A dermatitis problem developed as a result. Pain, swelling, and blisters on hands and forearms.
phenothrin
26002-80-2
Technical material nonirritant
mixture with tetramethrin sensitizer in Buehler study
o
resmethrin
10453-86-8
85% technical material is category III irritant
3% formulation weak sensitizer in Buehler study
o
tetramethrin
2117279
21 % mixed with 21 % resmethrin causes transient irritation
mixture with phenothrin is a weak sensitizer in Buehler test
o
cyfluthrin
68359-27-5
20% formulation nonirritant
Nonsensitizer in Buehler test
o
cyhalothrin
91465-08-6 68085-85-8
Technical material causes transient irritation
Inadequate data
o
cypermethrin
52315-07-5
27% formulation causes transient irritation
Nonsensitizer in animal tests; 40% wettable poweder labeled as a sensitizer
Type I pyrethroids 94-138: A worker spilled a mixture of pip butoxide and allethrin on the outside of a backpack sprayer and his arms while mixing a tank load and wiped off the sprayer with a paper towel. After applying the material, he noticed itching, redness, and swelling of the lower back and arms.
Type II pyrethroids
88-2388: A mixerlloader handling cypermethrin developed burning in the groin area, shortly after going to the bathroom without thoroughly washing his hands.
(continues)
11.2 Review of Use Categories
305
Table 11.1 (continued)
Predictive tests in animals
Data from pesticide handler data base 1982-1995
Draize irritation test
# of cases
Sensitivity
Compound
CAS #
esfenvalerate
66230-04-4
35% formulaton causes transient, mild irritation
Nonsensitizer by Buehler method
0
fenpropathrin
395151-41-8
Technical material nonirritant based upon observations in acute dermal toxicity study
Weak sensitizer in human repeated insult patch test
0
fenvalerate
51630-58-1
Category II irritant-persistent erythema at day 7
3.5% formulation nonsensitizer in Buehler test; 25% formulation labeled as a sensitizer
dicofol
115-32-2
50% formulation causes transient irritation
Sensitizer
3
84-954: An employee was mixing dicofol for an aerial application on corn, wearing gloves and face shield, when some material splashed up on his neck. He saw doctor 3 days later when the burning and itching on the front of neck did not improve after initial treatment with first aid ointment. The condition was recorded as a second degree burn. 84-1454: A mixerlloader/applicator splashed dicofol on the arms that soaked through his protective clothing. Erythema of the forearm was noted when he sought treatment 3 days later.
lindane
58-89-9
No data
Borerlleaf miner formulation is a sensitizer in the Buehler assay.
2
86-309: A hose split during an application under a house and the material sprayed onto the applicator's hands. He made repairs without gloves, washed off, but noticed a rash later in the day.
methoxychlor
72-43-5
25% formulation labeled as minimal irritant
No data
0
borax
1303-96-4
99% formulation labeled as nonirritant
Inadequate data
disodium octaborate tetrahydrate
3791278
Labeled as a nonirritant
Nonsensitizer
Case examples
86-1191: Concentrated pydrin® spilled from a measuring cup onto a mixerlloader's arm, but he did not wash or change clothes. A contact burn developed at the site, as recorded by the treating physician.
Organochlorines
Borates 84-197: A restaurant employee applying boric acid as a crack and crevice treatment, suffered a reported mild skin reaction.
Bacillus insecticides & other biologicals Azadiracthin-from Neem extract
1141-17-6
Nonirritant
Nonsensitizer
0
avermectin (abamectin)-mixture of avermectin Bla and BIb
71751-41-2
Nonirritant
Labeled as sensitizer
2
Bacillus thuringiensis
No CAS #
Nonirritant
Nonsensitizer in Buehler test
0
Capsicum oleoresin
404-86-4
Nonirritant
Labeled as a sensitizer
0
92-520: An employee developed skin problem on arm after spraying roses with Abamectin. He was wearing a rubber rainsuit, but felt wetness on his arm, and did not wash the affected area immediately. Examination showed ulcerative lesions with mild surrounding erythema on right proximal forearm. 92-2243: A worker developed a rash on his right hand while applying abamectin to roses. He developed a similar rash the previous year after spraying the same pesticide.
(continues)
306
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides
Table 11.1 (continued)
Compound
CAS #
Predictive tests in animals
Data from pesticide handler data base 1982-1995
Draize irritation test
Sensitivity
# of cases
Case examples
Miscellaneous chemical structure amitraz
33089-61-1
50% wettable powder labeled as corrosive
50% EC sensitizer in maximization study
o
benzyl benzoate
120-51-4
Technical material causes transient irritation
Labeled as sensitizer
o
butoxy polypropylene glycol
9003-13-8
Category III irritant
Inadequate
o
Diethyl toluamide (DEET)
134-62-3
Labeled as minimal irritant
Labeled as a sensitizer
3
hydramethylnon (aminohydrazine) bait
67485-29-4
Minimal irritant
Nonsensitizer
o
Imidacloprid
105827-78-9
Minimal irritant
Nonsensitizer
k salts of fatty acids (1596)
61790-44-1
49% formulation Irritant/corrosive
Inadequate
o o
oxythioquinox
2439-01-2
Technical material causes transient irritation
40% fomulation sensitizer in Buehler test
propargite
2312-35-8
Corrosive
Nonsensitizer
82-1871: Worker had an allergic reaction (hives) after treating himself with an insect repellent according to the label directions. 93-1422: Worker sprayed an insect repellent on her exposed skin before collecting a lab sample from treated sewage water. She suffered an apparent allergic reaction to the repellent a short time later.
Details of case not specified
55
82-1667: A mixerlloader splashed a few drops of concentrated propargite on his neck while opening a can for closed system loading. Over the following day, he developed a rash, which persisted for a week until he got it treated. 85-1642: During a mixlload operation, the nurse tank overflowed, dousing worker's arm with material and a chemical burn subsequently developed at the site of contact.
sulfluramid (bait)
o
4151-50-2
Nonirritant
Labeled as sensitizer
Captafol
2939-80-2
Labeled as irritant
Sensitizer per public domain literature
captan
133-06-2
Technical material nonirritant
Sensitizer per public domain literature
6
folpet
133-07-3
Technical material causes transient irritation
Sensitizer in the maximization test
o
17804-35-2
Labeled as minimal irritant
Labeled as a sensitizer
6
Fungicides
Phthalimido compounds 83-1783: An applicator developed a rash after using difolatan on tomatoes. He did not seek treatment until after the rash became infected. 84-1668: A worker was loading captan dust into a helicopter for application on grapes when he developed contact dermatitis. 87-260: After spraying captan using adequate protective equipment, an orchard applicator developed a rash on face, neck, and arms, thought to be an an allergic reaction. 87-694: A worker removed lids from 5 gal containers of captan, wearing gloves and goggles. He developed a rash, but did not get any of the material on his skin via splash or spill.
Carbamates benomyl
85-448: An applicator splashed benomyl onto his face and neck while spraying pruning cuts in vineyard. He suffered burning, erythema, irritation, and swelling of eyes and face. The condition was described as a first degree chemical burn.
(continues)
11.2 Review of Use Categories
307
Table 11.1 (continued) Predictive tests in animals
Data from pesticide handler data base 1982-1995
Draize irritation test
Sensitivity
# of cases
Compound
CAS #
thiophanate methyl
7912735
85% formulation causes transient irritation
50% formulation labeled as sensitizer; others not
mancozeb
8018-01-7
80% formulation nonirritant
Sensitizer in Buehler study in combination with thiophanate
88-1764: During hot weather, mixerlloader developed a rash after he started handling Dithane®. Examination showed a rash on the arms and abdomen characterized as contact dermatitis.
maneb
12427-38-2
No data
Sensitizer
84-811: An employee developed an allergic rash after spraying dithane on grapes
thiram
137-26-8
Technical material category IV
Sensitizer per public domain literature
84-1488: A worker treating seeds with thiram dust developed lesions around the respirator line on the day of the application.
zineb
12122-67-7
Inadequate
Labeled as a sensitizer
ziram
137-30-4
96% industrial formulation labeled as an irritant
Labeled as a sensitizer
2
copper
7440-50-8
85% formulation category III irritant
Labeled as sensitizer
2
copper ammonium carbonate
33113-08-5
24% formulation minimal irritant
Labeled as sensitizer
0
copper hydroxide
20427-59-2
90% formulation minimal irritant
Labeled as sensitizer; Buehler study on CuOH negative
3
94-561: A worker applied copper hydroxide to walnuts using a high-volume sprayer. His face began to itch and bum a few hours after he finished the application. The affected area was on the unprotected portion of his face. 90-1368: A worker applying fungicide to nut orchard, wearing all protective gear, got wet from rain blowing in around his hood and down his gloves. He then began itching in areas that had gotten wet.
copper naphthenate
1338-02-9
80% formulation category II irritant
68% CuNaphthenate Nonsensitizer in Buehler test
2
87-1724: A wood worker was painting a copper naphthenate wood preservative to the cut end of wood and developed a chemical bum to his arms. 90-1127: A student employee wearing rubber boots, gloves, goggles, respirator, and ransuit treated wooden benches with preservative. He accidentiy rubbed his neck and face while wearing the rubber gloves and developed contact dermatitis.
copper oxide
1317-39-1
97% technical material nonirritant
Animal tests do not show clear evidence of sensitization
0
copper oxychloride
1332-40-7
Technical material nonirritant
Labeled as sensitizer
0
copper sulfate
7758-98-7 7768-98-7
99% technical material minimal irritant
Sensitizer per public domain literature
4
Case examples
0
Thiocarbamates
83-298: A worker applied ziram with no hand or face protection and developed contact dermatitis. 84-518: An applicator spraying almonds with ziram developed a rash after a hose broke on his spray equipment.
Copper compounds
90-2391: After adding copper sulfate to water, a worker developed a rash on his forearms and itching all over. 90-2588: While an employee mixed copper sulfate, some powder got inside the glove causing the irritation to his right forearm. The resulting dermatitis was subsequently complicated by an infection. 93-1839: A worker applied copper sulfate granules to canal water by a piece of equipment he called the "sandblaster"-that air blasts the material on the canal water. After copper sulfate dust landed on his neck and chest, he developed large pruritic, erythematous patches on the neck and chest.
(continues)
308
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides
Table 11.1 (continued) Predictive tests in animals
Data from pesticide handler data base 1982-1995
Sensitivity
# of cases
Compound
CAS #
Draize irritation test
copper triethanolamine complex
68027-59-6 82027-59-6
3.5% formulation minimal irritant
Nonsensitizer
o
cupric oxide
1317-39-1
97% technical minimal irritant
Buehler method nonsensitizer
o
cuprous oxide
1317-38-0
85.7% category III irritant
Buehler method nonsensitizer
o
Anilazine
101-05-3
No data
Labeled as a sensitizer
2
carboxin
5234-68-4
Technical material is nonirritant
Mixture carboxin and other compoundsnonsensitizers
0
chloroneb
2675-77-6
30% formulation with 3.5% metalaxyl category III irritant
Negative test on mixture
0
fIusilazole
85509-19-9
60.7% formulation category III irritant
Buehler test on 20% formulation negative
0
fosetyl-al
39148-24-8
80% formulation causes transient irritation
Buehler test on 80% formulation negative
0
imazaIiI
3554-44-0
Technical material 98.5% causes transient irritation
Buehler test on 13.5% formulation negative; technical material is labeled as sensitizer
0
iprodione
36734-19-7
50% material nonirritant
Nonsensitizer in the Buehler assay
methylene bis(thiocyanate)
6317- 18-6
10% material corrosive
Labeled as sensitizer
mycIobutanil
88671-89-0
84.5% technical minimal irritant
Nonsensitizer
Pentachloronitrobenzene (PCNB)
82-68-8
No data
No data
0
sulfur
7704-34-89
98% formulation category IV
Nonsensitizer by Buehler method
28
TCMTB (thiocyanomethylthiobenzothiazole)
21564-17-0
30% formulation corrosive
10% material labeled as sensitizer
0
Case examples
Miscellaneous compounds 92-732: A mixer-loader for an aerial application developed rash on exposed skin areas while dumping wettable anilazine powder into mix tank. At examination, he had a generalized rash on face, neck, and arms thought to be allergic in nature.
84-897: A worker sprayed iprodione on grapes and developed a rash 3 days afterwards. He also worked in a thinning crew after app. No other crew members developed a similar rash. 0 92-1379: A worker overfilled a spray tank and spilled pesticide solution (mycIobutanil + adjuvant) on his feet. He rinsed his feet and shoes off with water, but did not remove his shoes. Examination showed pruritic dermatitis, scaling, and crusting of the bottoms of both feet.
82-1211: A worker's coveralls became complete while he applied sulfur and he developed dermatitis. 83-1252: An employee tore pants while applying sulfur to cotton, and got a rash in the area of the tear. 85-2263: A worker developed a rash on body, especially on neck, while dusting with sulfur. 87-174: A worker complained of rash after mixing, loading and applying Kolospray@ and had a 2 year history of sensitivity to the formulation.
(continues)
11.2 Review of Use Categories
309
Table 11.1 (continued) Predictive tests in animals
Data from pesticide handler data base 1982-1995
Draize irritation test
# of cases
Sensitivity
Case examples
Compound
CAS #
thiabendazole
148-79-8
Transient irritation with 98.5% formulation
Tests on mixtures only
triadimefon
4312173-43-3
50% formulation causes transient irritation
95% technical material is a sensitizer in the Buehler test
vincIozolin
50471-44-8
Transient irritation
Labeled as sensitizer
0
1,3 dichloropropene
542-75-6
Corrosive
Labeled as sensitizer
11
aluminum phosphide
20859-73-8
No data
No data
84-2184: A rash developed on torso after application of Phostoxin® under tarp for rice in warehouse. There was no history of direct exposure.
dazomet
533-74-4
Technical material is a minimal irritant
Labeled as sensitizer
90-2448: Applying dusty granular form of pesticide, a worker developed a rash at the belt-line as weII as front of legs, abdomen, and arms.
ethylene oxide
75-21-8
No data required as minimal dermal contact expected
Some products labeled as sensitizers
5
fosthiozate (2325)
not specified
Nonirritant
Technical material is sensitizer in the maximization test
0
metam-sodium
137-42-8
Corrosive
? Animal skin sensitizer
34
84-20: A worker handled vapam while wearing leather boots; the material soaked through the boots and irritated his feet. 85-1697: After making an application, he noticed pressure remaining in the hosel; the material splashed on him and skin began to bum.
methyl bromide
74-83-9
Corrosive
No data
59
82-32: Methyl bromide leaked into employee's boots when he changed cylinders on a tractor. His feet became inflamed, ulcerated, and infected over several days before employer noticed problem and sent the worker for treatment.
diquat
85-00-7
Irritant
No data
15
86-1498: An applicator wet his shoes with diquat and did not change them. He developed a rash on the top of one foot. By the time he saw a doctor, 8 days later, his foot had become infected.
paraquat
1910-42-5 50-2
Irritant
Labeled as sensitizer
17
83-480: A gust of wind blew material onto arms which had previous abrasions and his condition was aggravated by contact with paraquat.
0
82-1412: While applying BayJeton® to grapes, the wind blew spray back on to the applicator. He suffered a reported aIIergic reaction.
Fumigants 88-2091: An employee was uncIogging one of the tubes through which the telone is injected on his spray rig when material spiIIed on his foot. He continued to work and the next day he had blistering and sweIIing offoot.
87-2720: A hospital worker stuck her hand in a sterilizer, before it had aerated to get rid of the ethylene oxide and suffered a chemical bum on her hand.
Herbicides Bipiridyls
Chloracetanilide alachlor
15972-60-8
metolachlor
51218-45-2
Sens per precaution
85% formulation category 3 irritant
Mixed formulations negative in Buehler test
84·537: An applicator developed a fine rash on trunk, arms, and legs two years in a row after handling alachlor. Condition was reported as suspected aIIergic dermatitis. 0
(continues)
310
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides
Table 11.1 (continued)
Compound
CAS #
Predictive tests in animals
Data from pesticide handler data base 1982-1995
Draize irritation test
# of cases
Sensitivity
Case examples
Nitroaniline compounds
benefin
1861-40-1
Technical material cat 2 irritant
Mixtures with other herbicides labeled as sensitizer
ethalfluralin
55283-68-6
Technical material cat 2 irritant
Labeled as sensitizer
oryzalin
19044-88-3
85% technical material minimal irritant
75% formulation labeled as sensitizer
5
84-51: The wind blew material into an applicator's face, resulting in a rash on neck and face. 84-272: An applicator wiped his face with a wet glove, and developed a rash immediately.
pendimethalin
40487-42-1
2.5% formulation minimal irritant
Nonsensitizer
2
82-467: A mixerlloader/applicator developed a rash while applying pendiemethalin. Examination showed a rash on both arms up to elbow.
trifluralin
578064
80% formulation caused minimal, transient irritation
Sensitizer per public domain literature
7
93-340: An applicator was loading his tractor with trifluralin and some of the material leaked out and was blown onto his face. He failed to wash the exposed area right away and developed developed an itching, burning, red rash on face.
bensulide
741-58-2
Nonirritant
Inadequate
glyphosate
1071-83-6
Minimal irritant
Nonsensitizer
70
sulfosate
81591-81-3
62% technical material minimal irritant
Sensitizer
0
tribufos
78-48-8
70% formulation labeled as corrosive; technical material category III
No data
0
94-75-7
Technical material nonirritant
Labeled as sensitizer
0
dicamba
1918-00-9
Minimal irritant
Nonsensitizer
0
MCPP
7085-19-01
Minimal irritant
Nonsensitizer
0
fenoxaprop ethyl
66441-23-4
Irritant
Inadequate data
0
dithiopyr
97886-45-8
91.5% technical
12.7% material sensitizer in Buehler test
0
imazethapyr
81335-77-5
90% formulation category 4 in Draize
22.9% formulation negative in Buehler study
0
triclopyr
5721-4069-1
61.6% formulation causes transient irritation in the Draize test
Buehler study on 17% formulation was negative. However, the 44% formulation is labeled as sensitizer
2
0
90-1832: An applicator disconnected a filter valve, was sprayed in the face with herbicide, and developed pruritus.
Organophosphates 83-1770: A hose ruptured and sprayed his arm. A rash developed, reported to cause 20 days lost-work-time. 83-220: The wind blew spray mist onto a worker's forehead while he was applying a glyphosate formulation. He experienced a rash and itching at the site of contact that lasted for for several days. 83-917: As a worker was treating vineyard weeds with a glyphosate formulation, a hose burst on his backpack sprayer, covering his back with the material and he subsequently developed a rash at the site of contact.
Phenoxy herbicides
2,4-D
Pyridine derivatives
88-2832: The worker was spraying weeds along the roadside when a big rig drove by causing a shift in the wind direction and the spray blew back in her face. She developed mild erythema on the face.
(continues)
11.2 Review of Use Categories
311
Table 11.1 (continued)
Compound
CAS #
Predictive tests in animals
Data from pesticide handler data base 1982-1995
Draize irritation test
# of cases
Sensitivity
Case examples
Thiocarbamates and carbamates
molinate
2212-67-1
thiobencarb
15% granular flake formulation was nonirritating in Draize test, but caused transient irritation when moistened prior to application
10% formulation was a nonsensitizer in the Buehler assay
84% material caused very minimal and transient irritation in Draize test
84% material is a nonsensitizer in Buehler test
87-937: An employee was loading Ordram® bags into the bucket of the loader truck for an aerial application. Some of the material got down his protective clothing, contacting his legs and feet and he subsequently developed a rash in the corresponding areas.
Triazines
atrazine
1912-24-9
No data
No data
0
cyanazine
21725-46-2
Technical material caused transient erythema in the Draize test
A Buehler study on a mixture of cyanzine and metolachlor was negative
0
prometryn
7287-19-6
Technical nonirritant in Draize
45% material minimal reaction in Buehler test
0
simazine
122-34-9
90% formulation minimal irritant
6.3% simazine negative in max. test
314-40-9
80% formulation category 4 Draize
Nonsensitizer in Buehler test 40% form, mixed with 40% diuron
0
83-2394: Mixlloading material, apparently urinated during operation, depositing material on penis. Developed rash.
Urea herbicides
bromacil
chlorsulfuron
64902-72-3
75% minimal irritant
No data
0
diuron
330-54-1
81 % formulation category III irritant
Nonsensitizer in Buehler test
0
thidiazuron
51707-55-2
12% formulation with 6% diuron corrosive
Ginstar® sensitizer in Buehler test
0
dichlobenil
1194-65-6
Technical material category 4
0.55% root control sensitizer in maximization test
0
flumetsulam
98967-40-9
Mixture with metolachlor category 2 irritant
Mixture labeled as sensitizer
0
isoxaben
82558-50-7
Nonirritant
Nonsensitizer in Buehler assay
0
sethoxydim
74051-80-2
13% formulation category 11 irritant
13% formulation nonsensitizer in maximization test
polymerized pinene
None given
Summary of study describes material as slightly irritating
Nonsensitizer in human repeated insult patch test
0
Stephan C-65
Not given
Irritant
Nonsensitizer
0
Miscellaneous
88-1253: A worker was pumping up sprayer when leaky gasket caused solution to spray in face and he developed a dermatitis characterized as a chemical bum.
Adjuvants
312
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides
Sodium hypochlorite
Emmett et al., 1989; Foussereau et al., 1984; Mathias et al., 1983; Menne, 1991; Menne et al., 1991; Pilger et al., 1986; Thormann, 1982). In the pesticide handler database, the isothiazolin compounds appear to be relatively frequent causes of irritant skin reactions or frank chemical bums following accidental direct contact (e.g., cases 89-1312 and 90-552). The data are limited, however, because the individual isothiazolin compounds involved in each case are not identified.
JN~ Ld Octhilinone
li
N~N
O~9~OH Cyanuric acid
o
6
N
",,","""';=1;00" hexahydro-1,3,5-triethyl-s-triazine
N -
Cl
OH
Na-OCI
Kat~on DP, LX and other trade names
©IN
11.2.1.4 Phenolic Compounds Promexal X50 preservative
1,2-benzisothiazol-3(2H)-one Proxel
11.2.1.2 Quarternary Amines The quartemary amines have a capacity akin to isothiazolin compounds to cause both irritation and sensitization in animal test models. They are also often used in mixtures and are grouped together in the handler database. The 107 cases associated with quartemary amines accounted for 11.8% of the total number reported. Typical cases (88-1893) occurred on direct contact.
A review of the DPR label database showed 128 separate phenol compounds with DPR chemical codes, including sodium phenate salts of several phenol derivatives. O-phenyl phenol is a typical compound (Table 11.1), identified as an irritant in the Draize test, but no sensitization test was available for review. Its sodium salt has been identified as a cause of contact urticaria (Tuer et al., 1986). P-tert-butyl phenol is notable for being identified as a human sensitizer and as an occasional cause of leukoderma (Mancuso et al., 1996; O'Malley et al., 1988).
OH
~CH3)2-N-(C8HlO)~ + Cl Dioctyl dimethyl ammonium chloride
ortho-phenylphenol (CH3 h - N Cl
~.5H3 HO~CH3 CH 3 p-tert-butyl phenol
tXHx+2 x> 14
Alkyl dimethyl benzyl ammonium chloride
QCH3)2-N-(ClOH12~ +CI didecyl dimethyl ammonium chloride
11.2.1.3 Chlorine Compounds and Triazine Chlorine Stabilizers The animal data on sodium hypochlorite show irritant effects (Table 11.1), and sensitization has been reported in the form of contact urticaria (Hostynek et al., 1989). Nevertheless, most sodium hypochlorite-containing antimicrobials are not labeled as sensitizers. The cases reported in the handler database occurred principally in end users of sanitizers and disinfectants (e.g., 87-1468). Cyanuric acid was associated with several cases in the handler database. The other chlorine stabilizer hexahydro-l ,3,5-triethyl-s-triazine ([7779-27 -3]) was corrosive in the Draize test and a weak sensitizer in the Buehler assay.
11.2.1.5 Other Antimicrobial Compounds Among the miscellaneous antimicrobial compounds, only iodine was associated with a case in the handler database. Nevertheless, many of the compounds, (2,2-dibromo-3-nitrilpropionamide), 2-(hydroxymethy lamino )-ethanol, 2-bromo-2nitro-1,3-propanediol, and sodium pyrithione, appear to be corrosive or severely irritant. Bronopol has long been recognized as a sensitizer (Camarasa, 1986; Frosch et al., 1990), related to its capacity for releasing formaldehyde (Kranke et al., 1996); 1,2-dibromo-2,4-dicyanobutane (Tosti et a!., 1995; Vigan et a!., 1996) and iodine (Ancona et al., 1985; Erdmann et al., 1999), have also been associated with cases of contact sensitization. Organotin compounds have been reported as irritants (Gammeltoft, 1978). Allergenic effects of tributyl tin compounds generally are probably not significant, but tributyltin methacrylate is an equivocal sensitizer in the Buehler test (Table 11.1).
11.2 Review of Use Categories
o 11
CHBr2-C-NH-CN
313
dler database appeared to be instances of mild irritation (e.g., 85-343), consistent with the transient irritation occurring in animal tests of technical material.
2,2-dibromo-3-nitril-propionamide OH 1
N-CH 3 1
CH 2-CH 2-OH
2-(hydroxymethylamino)-ethanol Br 1
HO-C-C-C-OH 1
N02 2-Bromo-2-nitropropane-1,3-diol Bronopol
Naled (Dibrom®) is labeled as a corrosive and also as a sensitizer (Table 11.1), presumably because of the cases of contact sensitivity (Edmundson and Davies, 1967) and irritation (Mick et aI., 1970) reported in the public domain literature. Cases of irritant dermatitis similar to those described by Mick were also found in the handler database (88-2330 and 88-942).
Sodium 1-hydroxypyridine-2-thione Omadine
1,2-Dibromo-2,4-dicyanobutane
11.2.2 INSECTICIDES AND INSECT REPELLANTS 11.2.2.1 Organophosphates The organophosphates are generally thought to cause minimal irritation (Rycroft, 1977) and are nonsensitizers in animal models (Table 11.1). Nevertheless, some organophosphates, such as parathion and malathion, have previously been reported as causes of contact sensitization or other dermatoses. Details of the animal studies and pesticide applicator cases are discussed for compounds which have been possibly associated with significant skin effects.
Concentrated forms of chlorpyrifos (e.g., 40.7% formulation of Lorsban) are irritants in animal studies, but all of the numerous formulated products containing chlorpyrifos are negative in the animal sensitization studies conducted by the Buehler method. Paradoxically, the technical formulation with 98% chlorpyrifos causes less irritation than the 40.7% formulation. The 11 cases associated with chlorpyrifos in the pesticide han-
Although a report by Milby identified malathion as a sensitizer, animal sensitization studies submitted for pesticide registration did not employ adequate negative controls to verify that the dermal reactions observed were due to sensitization rather than irritation (Milby and Epstein, 1964). Nevertheless, the irritation studies conducted, per se, showed that technical malathion causes minimal irritation. Cases associated with malathion in the handler database are consistent with mild, transient irritation. Although diazinon appears to be minimally irritant, in animal tests, cases in the handler database are consistent with an irritant mechanism (87-2537). No sensitization studies on diazinon were reviewed, but most of the diazinon products reviewed were labeled as sensitizers, as required in the V.S. EPA's diazinon reregistration standard. A case report from Australia identified an isomer of diazinon as a cause of porphyria cutanea tarda (Collins et aI., 1982).
314
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides
11.2.2.2 Carbamates None of the carbamate insecticides appeared to be markedly irritant or to be sensitizers in animal test models. One carbaryl formulation containing pyrethrins was found to be a sensitizer, but none of the formulations containing only carbaryl produced similar responses. Cases associated with carbamates in the handler database appeared to be principally transient irritation following direct contact (e.g., 88-297, 84-512). The transient irritation reported in the applicator cases may be due to nonpesticidal ingredients: Draize tests on carbaryl, methiocarb, methomyl, and propoxur did not show even transient irritation. Cases of dermatitis associated with application of carbamates are also reported in the public domain scientific literature (Bruynzeel, 1991; Vandekar, 1965).
o 0-8-NH-CH 3
00 Carbaryl
CH3-C
a category 11 irritant in the Draize test, but the concentration in ready-to-use formulations is less than 1%. The only dermal sensitization studies reported involve mixtures with pyrethroids. Cases reported in the handler database also had simultaneous exposure to pyrethrins or pyrethroids.
N-octyl-bicycloheptene dicarboximide (NOBD) is also used as a pyrethrin synergist in hundreds of formulations. The only irritation study available for review involved a mixture with diethyl toluamide, isochromyl cinchonerate. This mixture caused only transient irritation and was scored as category IV. There is no study available on the capacity of NOBD to cause sensitization. Cases reported to the handler database all involved mixtures with pyrethrins and other compounds.
0
=N-O-~-NH-CH3
~-CH3
Methomyl
11.2.2.3 Pyrethrins
n-octylbicyclJ"heptenedicarboximide
The California Pesticide Label Database shows more than 1000 pyrethrin formulations registered (2/98). The cases associated with pyrethrins reflect their broad-scale use, nearly always in mixtures with piperonyl butoxide, and often with pyrethroids, carbamates, organophosphates, and other materials. Many formulations containing pyrethrins are positive in the Buehler test, and this is reflected in the product labeling. Dermal irritation studies with technical pyrethrins (57% concentration) show only transient erythema, disappearing by 72 hours. Irritation suffered by users (Table 11.1) may be due to petroleum distillates or piperonyl butoxide (92% formulation is a grade 11 irritant in the Draize irritation test) contained in formulated products, as well as the pyrethrinlpiperonyl butoxide combination. C~
Jc-8-o--Q "'" C~
'C=CH CH!
C~ Pyrethrin I
0
C~CH=CHCH=C~
C~ 0
11.2.2.4 Pyrethrin Synergists Piperonyl butoxide is a synergist used with formulations of pyrethrins and pyrethroids, but is not chemically related to either group. The technical formulation (92%) was classified as
11.2.2.5 Synthetic Pyrethroids Effects of synthetic pyrethroids on the sodium channels of cutaneous nerve endings may cause paresthesias at levels of exposure that do not provoke visible erythema (Lisi, 1992). The standard Draize irritation study may be a poor means for evaluating such purely symptomatic endpoints. An alternative animal test developed by Cagen evaluates the sensory effect of pyrethroids through observations of grooming behavior focused on the site(s) of applied test material. The behavioral test demonstrated direct effects on grooming behavior for 4 hours after pyrethroid application and increased response to other chemical irritants (oil of mustard in the test model) for 24 hours after application (Cagen et aI., 1984).
Type I Pyrethroids These contain two cisltrans isomeric sites and may have as many as four isomers with ability to stimulate cutaneous nerves in the human epidermis (Flannigan and Tucker, 1985; Flanniganet aI., 1985a, b; Gammon, 1985; Gammon and Casida, 1983; Tucker et aI., 1984). Allethrin Technical d-allethrin is a cis-trans mixture that shows minimal irritation in the Draize test; a Buehler sensitization study was negative on a dilute end-use product (a mixture
11.2 Review of Use Categories of allethrin, cypermethrin, piperonyl butoxide, and petroleum distillates). Case 94-138 in the handler database (a formulation of allethrin and piperonyl butoxide) involved irritation on direct accidental contact.
315
Type 11 Pyrethroids These contain as many as three isomeric sites and most contain a cyano group attached near the ester linkage. They are relatively more potent systemic toxins than the type I pyrethroids and cause a greater degree of paresthesia in experimental studies on human volunteers. Cyfluthrin and Cyhalothrin Cyfluthrin and cyhalothrin both caused minimal irritation in the Draize test and were nonsensitizers in the Buehler assay. Neither was associated with cases in the handler database.
Permethrin Technical permethrin shows minimal irritation in the Draize test, with only very slight erythema persisting 72 hours after initial application. Several formulations appeared to be nonsensitizers in the Buehler assay, but a mixture of 1% permethrin and 1% piperonyl butoxide showed mild sensitization reaction on the rechallenge portion of the assay. Case 92-1381 in the handler database, associated with permethrin, appeared to be due to cumulative irritation.
Phenothrin and Tetramethrin A mixture of 5% phenothrin and 5% tetramethrin proved corrosive in Draize tests, according to the product labeling. Nevertheless, technical phenothrin caused no irritation in the same assay. Sensitization tests were negative except for a mixture of phenothrin and tetramethrin, which showed marginal reaction on rechallenge in the Buehler assay. No cases were included in the handler database.
Resmethrin Technical resmethrin caused mild persistent irritation in the Draize assay. A 3% formulation was also a sensitizer in the Buehler assay. No cases were included in the handler database.
--cr°
C=Ck-0 cH{ 1 C-o-CHz CfiJ
CfiJ
C~ Cl
~
CN
~CH-C~2cH-t-o-6H~
~~ 0
CH 3/ 'CH 3
0
A-cyhalothrin
Cypermethrin Technical cypermethrin causes minimal irritation in the Draize test. The 40% wettable powder is labeled as a sensitizer, but the Buehler assay on the 18% formulation of cypermethrin showed no evidence of sensitization. A case of dermatitis in the handler database was associated with accidental transfer of cypermethrin from the hands to the genitalia (88-2388).
o
CN
0
:~CH-C~~CH-~-O-CH~ ~ CHI 'CH3 Cypermethrin
Fenvalerate The 24% formulation is an irritant (category 11) in the Draize test and also a sensitizer in the Buehler assay. Case 86-1191 in the handler database was described as a bum in a patient who spilled fenvalerate on himself and did not promptly decontaminate his clothing.
Tetramethrin
CfiJ,
Cyfluthrin
CHz
~
Resmethrin
Fenvalerate
11.2.2.6 Organochlorines
-
Dichloro-diphenyl-trichloro-ethane (DDT) is the model organochlorine compound, but is not used legally anywhere in the world, except in public health vector control campaigns. Organochlorines still registered in the United States are those with the shortest environmental and biological half-lives.
316
CHAPTER 11 Regulatory Evaluation of the Skin Effects of Pesticides C--cl3
cl---IQ\-6~1 ~o~~
CCIs
CH °-Q-Z=Q-°CH 3
3
Methoxychlor
Dicofol Dicofol is an agricultural insecticide with 11 current product labels registered in the United States. It is structurally similar to DDT, but contains a central hydroxyl group that allows more rapid environmental and metabolic degradation. A 50% formulation of dicofol causes minimal skin irritation and is also labeled as a sensitizer. The handler database contains a report of irritant contact dermatitis following contact with dicofol occluded against the skin for a prolonged period of time (84-954 and 84-1454).
Endosulfan There are 14 products containing endosulfan currently registered in California. No data were reviewed for either dermal sensitization or dermal irritation, but the product is not labeled as a sensitizer or as an irritant. No cases associated with endosulfan were included in the handler database.
Cl Cl
°8=0 \
d
C--cl3
CI-©-6~1 Oicofol
bH
11.2.2.7 Borates Lindane Lindane is used for control of ectoparasites on cattle and for control of insects on a variety of commercial crops. A 20% formulation used for control of borers and leaf miners is labeled as a skin irritant. It is also a sensitizer in the Buehler assay. The handler database contained two cases of contact dermatitis following direct accidental exposure to lindane, both consistent with irritant reaction. Dermatitis has also been reported among workers in lindane manufacturing operations, but the reported cases were possibly attributable to precursors and by-products not typically found in commercial formulations of lindane (Smith, 1991a). Although the agricultural products may contain as much as 40% lindane, post-treatment dermatitis has also occasionally occurred in patients treated for scabies with 1% formulations of lindane. The extensive series reported by Farkas also contained cases reacting to a 20% scabicidal formulation of sulfur (Farkas, 1983).
Cl
*:
Borates (including borax and hydrated octaborates) are nonsensitizers and nonirritants in animal tests. Case 84-197 in the handler database described a transient skin reaction associated with handling borates in animal tests.
11.2.2.8 Biological Insecticides and Repellants Two products with active and l32 products with active registrations in California contain Bacillus thuringiensis (BT) variant Berliner. Related products include BT variant San Diego and BT derived endotoxins. Animal studies demonstrated that a 15% formulation of BT variant Berliner, subspecies Kurstaki, strain EG 2348 caused no dermal irritation. A formulation of BT variant San Diego showed a slight reaction in the Buehler sensitization assay (score = 0.2 at 48 hours), but the sensitization score for the compound was markedly less than the score for 0.05% dinitrochlorobenzene (DNCB), used in the study as a positive control. The product label for a formulation of BT endotoxin mixed with killed Pseudomonas ftuorescens identifies it as a dermal sensitizer. A 2.5% formulation of Capsicum oleoresin (pepper plant extracts) was surprisingly non-irritant in the Draize tests and was also negative in the Buehler assay.
Lindane
Methoxychlor There are two products containing methoxychlor currently registered in California (mixtures with captan and malathion) and 134 inactive labels listed in the product label database (California Department of Pesticide Regulation, 1998). Worldwide, the relative use of the product is reported to be increasing. Its range of applications is similar to that for DDT, but it has a markedly shorter environmental and biological half-life (Smith, 1991a). No animal data are on file for either sensitization or irritation studies. There were no cases associated with methoxychlor included in the handler database.
Azadirachtin is a triterpenoid derived from Neem oil. The 10% technical material was completely nonirritating in the Draize test. A dermal sensitization conducted by an unspecified method was reported for a product containing 0.25% azadirachtin, but was not adequate to allow judgment of its sensitization status. None of the azadirachtin products listed in the
11.2 Review of Use Categories
label database are reported as dermal sensitizers and there were no cases listed in the handler database. Avermectin is a mixture of avermectin Bla and BIb used as an ant and cockroach bait and sometimes for nursery and agricultural pest control. The 1.9% formulation used as an agricultural miticide is labeled as a dermal sensitizer, but the bait formulations are not. Avermectin produced no irritation in the Draize test, but the case reported in the handler database (92520) was compatible with an irritant mechanism.
H)5 CH,
0
)5
CH 3
CH 3 CH 3
0
'H CH 3
Avermectin 1b
11.2.2.9 Miscellaneous Insecticides and Repellants
The aniline derivative amitraz is used for ectoparasites in veterinary practice and for pear psylla control on pears and for whitefly on cotton. There were no dermal irritation data on file, but a formulation of 50% wettable was labeled as corrosive. Amitraz also proved to be a sensitizer in the guinea pig maximization test. There were no cases associated with isolated exposure to amitraz in the handler database. CH 3
CH 3
CH3-o-N=CHN-CH=N~H3 tH3
317
CH3 I
C,.H9o-(CH2CHO)n-CH2CHOH I
CH3 Butoxy poly propylene glycol Diethyltoluamide (DEET) is an insect repellant sold in many over-the-counter formulations that range from 20% to 100% active ingredient. It causes transient irritation in the Draize test and is a non-sensitizer in both the Buehler and maximization tests. Nevertheless, some formulations are reported as sensitizers on the product label because of published reports indicating that DEET can cause contact urticaria (Maibach and Johnson, 1975; von Mayenburg and Rakoski, 1983; Wantke et aI., 1996). Three cases listed in Table 11.1 were not classified as involving pesticide handlers and were not included in the handler database. However, two (82-1871 and 93-1422) involved apparent allergic reactions similar to those reported by Maibach.
Hydramethylnon is an insecticide used in roach, ant, and termite baits. It is a nonsensitizer in the Buehler assay. Dermal irritation studies were not reported. Imidacloprid is a soil, seed, or foliar insecticide formulated as a wettable powder, flowable, and granular. It caused only transient irritation in the Draize test and was a nonsensitizer in the Buehler assay. There were no cases associated with imidacloprid in the handler database.
CX
H ,
HNrNH
Amitraz
Benzyl benzoate is a benzoic acid derivative used for control of mange in dogs. The technical material produced transient erythema in the Draize test. No adequate dermal sensitization study was reviewed. The technical material, but none of the derivative formulations, is labeled as a sensitizer. Q - C H 2 --©-COOH Benzyl benzoic acid
Butoxypolypropylene glycol is a fly repellant used for dogs and cattle. A mixture with piperonyl butoxide and pyrethrins (both nonirritants) caused mild irritation in the Draize test. There was no dermal sensitization study available for review. No cases were reported in the handler database.
~ F3C- Q - C H = C H - t - C H = C H - Q - C F3 Hydramethylnon
Insecticidal soaps (potassium salts of fatty acids) are used to control aphids and spider mites on plants and vegetables in gardens and nurseries. A 49% formulation was reported as corrosive in the Draize test. Dermal sensitization data were not reported and there were no cases associated with insecticidal soaps in the handler database.
aNy Imidacloprid
'NO 2
318
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides
Oxythioquinox is a miticide, insecticide, and fungicide used for control of mite eggs and mildew on deciduous fruit. It is a minimal irritant in the Draize test. The 40% formulation is a sensitizer in the Buehler test. One case was reported to the handler database, but the details were not specified.
Oxythioquinox
Propargite is a miticide with broad use on many California crops. The technical material is corrosive in the Draize assay. A study compared the dermal sensitization potential of propargite with that of iprodione following a California field worker dermatitis that involved exposure to both compounds. It used a modified version of the Buehler method. While the study had some technical deficiencies (i.e., lack of a positive control group and the use of the same animals to test both products), a number of the findings were of significance. In the range-finding portion of the study, it was determined that iprodione could be applied during the challenge tests at the maximum concentration allowed by the protocol (5%); propargite could only be applied at concentrations of 0.1 %. Both materials produced less reaction during the challenge portion of the study than during the induction phase, indicating neither material was a sensitizer under the conditions tested (O'Malley et aI., 1990). There were 55 cases associated with propargite in the handler database. Typical cases shown in Table ILl (82-1667, 85-11642) both involved chemical burns during application accidents. Many additional cases occurred following initial registration of propargite, prior to control of the hazard of mixing powdered forms of propargite by using water-soluble bags. The separate issues involved in regulating the hazards of propargite residues on crops harvested, or cultivated, with manual labor are discussed in a separate section.
o
(CH 3b-
c--O-o/\O-~-CH2-C=CH ~
Propargite
Sulfturamid is a sulfonamide derivative used as a cockroach and ant control bait. It is a nonirritant and a nonsensitizer. There were no cases associated with its use listed in the handler database.
11.2.3 FUNGICIDES
Compounds used to control fungi on plants overlap to some degree with the broader category of antimicrobial compounds. There are numerous chemical classes represented.
Sulfluramid
11.2.3.1 Phthalimido Compounds Captan is a fungicide commonly used on grapes, apples, almonds, and other crops. It is formulated as a wettable powder, as a dust, and as ftowable powders-alone or in combination with other fungicides and insecticides. In addition to use as a pesticide, it has been used successfully as a treatment for pityriasis versicolor (Simeray, 1966). A technical formulation of captan caused no irritation in the Draize test. No animal sensitization study was reviewed, but product labels indicate that captan is a sensitizer (Table 11.1). Jordan and King found a 5% sensitization rate to captan using a modified Draize test on volunteer subjects, and a 10% sensitization rate on volunteers using captan in the human maximization test. Women appeared to become sensitized more frequently than men (Jordan and King, 1977). Captan has also been reported to cause dermatitis in association with apple spraying in Scandinavia (Fregert, 1968); this has also been a relatively frequently reported problem in California. In Japan, a series of 178 patients at the Nagoya City University Medical School were routinely tested between 1977 and 1980 with the North American Contact Dermatitis Research Group of standard allergens, and 5.6% had significant positive reactions to captan. No clinical details were given in the report, but the surprisingly high percentage reacting to captan, presumably an uncommon exposure, raises the possibility that the material cross-reacts with other allergens in the standard series (Hirano and Yoshikawa, 1982). Rudner observed a similar high percentage of captan reactors in the North American Contact Dermatitis Group results in 1976 and speculated that results might be due to cross-reaction with thiurams (Rudner, 1977).
o
~N-S-CCI3
o
o
Captan
CaptafoI has a chemical structure nearly identical to that of captan, and has many similar uses as a fungicide. Captafol
11.2 Review of Use Categories
o ( y \ - S - C C I 2-CCI 2 H
~ Captafol
accounted for 62 (28.7%) of a series of 274 cases of pesticideassociated contact dermatitis seen in Japan between 1968 and 1970 (Matsushita et aI., 1980). In a similar series, 22 (18.2%) of 121 Korean farmers likewise reacted to the material (Lee et aI., 1981). Cottel observed several cases of San Joaquin Valley orchard farmers with positive patch test responses to a 0.1% aqueous preparation of captafol (Cottel, 1972). Irritant and allergic contact dermatitis was also seen in 23% of 133 New Zealand timber workers tested with the material by Stoke (Stoke, 1979). Camarasa found 4 of 7 ill workers from a captafol packaging plant had 3+ patch test responses to 1% captafo1 (Camarasa, 1975). An outbreak of dermatitis due to captafo1 sensitivity was also seen among a group of 36 workers on a Kenyan coffee plantation (Verhagen, 1974). Urticaria and asthma were part of the clinical picture reported affecting 7 (17.1 %) of 41 workers in a captafol packing operation in a chemical shed (Camarasa, 1975). The similar occurrence of asthma and contact dermatitis in a welder employed by a maintenance firm which serviced captafol distribution plants was reported by Groundwater (Groundwater, 1977). Captafol thus apparently is capable of causing both delayed and immediate types of hypersensitivity, as well as irritant dermatitis.
319
Plondrel@ on roses and subsequently developed dermatitis. All four reacted to 0.1 % Plondrel@ in petrolatum, but no reactions occurred in twenty control subjects tested with the same material (van Ketel, 1975). Van Kete1 subsequently reported a third case of hand eczema in a 21-year-old florist who had a 3+ reaction to 0.1 % plondrel (van Ketel, 1977). 11.2.3.2 Carbamates Benomyl is a benzimidazole compound with a carbamate moiety, but has no activity as a cholinesterase inhibitor. It is used in the control of many diseases of fruits, nuts, vegetables, and ornamental plants. Guinea pig tests of benomyl for irritancy conducted by the manufacturer at 12.5% and 25% aqueous dilutions were reported to be negative (Matsushita and Aoyama, 1981). However, the maximization test conducted in the same study showed 2% benomyl to be a potent experimental allergen. The first report implicating benomyl as a contact allergen appeared in 1972. Seven Japanese women employed in a greenhouse by a carnation grower developed dermatitis of exposed skin after benomyl was sprayed there on two occasions. No cases occurred until two weeks after the second spraying. The seven patients had 2+ reactions to a 1:10 dilution of benomyl in olive oil; three control subjects were negative (Savitt, 1972). Van Ketel also reported a case of be no my I sensitivity, confirmed by patch testing (with a 1% preparation which elicited no reaction from 10 controls), in a begonia grower (van Ketel, 1976). A second report from the Netherlands also highlighted the occurrence of benomyl hypersensitivity in nursery workers and florists (van Joost et aI., 1983).
o 0-S-CC'3
~FolPet
Folpet is a fungicide intended for protection of fruits, berries, and ornamentals. It currently has limited use. The 88% technical material is a minimal skin irritant in the Draize test, but is a sensitizer in the maximization assay. There were no cases associated with its use in the handler database.
Plondrel (ditalimifos) Plondrel is a fungicide that structurally resembles captan and captafol, that is sometimes classed as an organophosphate because of its phosphothioate group. In 1975, van Ketel reported the cases of two culturists and two florists who sprayed
The above cases illustrate the capacity of foliar residues of benomyl to cause allergic contact dermatitis in nursery workers. Zweig et al. (19??) demonstrated that exposure up to 5.4 mg/person-hour to benomyl is also a potential problem in strawberry harvesting. Everhart studied benomyl applicators and noted a maximum total exposure of 26 mg of benomy1 in mixing/loading operations, and markedly lower total exposures associated with field residue exposure (12 mg), and home use of the material ( < 1 mg) (Everhart and Holt, 1982). Hargreave noted the possibility of exposure from handling treated commodities; he demonstrated persistent benomyl residues on litchi nuts up to 15 days after postharvest treatment in a dipping process: 20 ppm of benomyl in the skin and 1.3 ppm in the flesh of the nut (Hargreave, 1983). Thiophanate methyl is a systemic fungicide used on vegetables, beans, nuts, and turf. An 85% formulation is a minimal irritant in the Draize assay. A formulation containing both mancozeb (63%) and thiophanate methyl (15%) was a sensitizer in the Buehler assay, but a formulation containing only thiophanate methyl (2.5%) did not provoke a sensitization response.
320
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides
There were no cases associated with thiophanate methyl in the handler database.
U
S
0
!
NH-8-NH-8-OCH3
CH 2- NH
~NH-C-NH-C-OCH3 ~
(Matsushita et ai., 1976). No data on irritation or sensitization studies submitted for registration were available for review.
-!-s-l
ltH2-NH-?i-S-Zj
&
Zineb
S
N
N>1
Thiophanate methyl
11.2.3.3 Thiocarbamates The thiocarbamate group of fungicides structurally resembles the rubber accelerator disulfiram (Antabuse®, tetraethylthiuram disulfide-CAS # 97-77-8), a common sensitizer present in both the European and the North American standard patch test series (Adams and Fischer, 1990). The prototype thiocarbamate fungicide, thiram (thiuram) is simply the methyl analog of disulfiram, and experimentally (Freundt and Netz, 1977) has a similar effect on the metabolism of alcohol. Technical formulations of thiram do not cause irritation in the Draize test, but the thiram case reported to the handler database was consistent with an irritant mechanism (84-1488). Most thiram formulations are labeled as potential sensitizers, because of reported cases of sensitization in the clinical literature (Cronin, 1980; Schultz and Hermann, 1958; Shelley, 1964).
Cases of allergic reactions to maneb documented with provocation (patch) testing have been reported in the clinical literature. Typical cases described from the Netherlands included two office workers who had purchased maneb spray to care for the plants in their office, and a 51-year-old woman who worked as an assistant in a flower shop (Nater et ai., 1979). Similar cases have been reported from the United States (Adams and Manchester, 1982), Italy (Peluso et ai., 1991), and Germany (Koch, 1996). A case reported to the handler database (Table 11.1) involved a possible allergic reaction after spraying a maneb-containing formulation of dithane® (84-811). Mancozeb is a polymer similar in structure to zineb and maneb, containing both zinc and manganese. The Draize irritation study reported to the registration database for the 80% formulation of mancozeb showed a minimal irritant reaction, and a combined formulation of mancozeb and thiophanate methyl was a sensitizer in the Buehler assay. A mancozeb case (88-1764) reported to the handler database was compatible with a simple irritant mechanism, but a recent case reported from Japan identified mancozeb as a cause of allergic contact dermatitis and photodermatitis (Higo et ai., 1996).
S
CH2-NH-~-S,-...----Mn/Zn
Thiram
t
The structure of ziram is very similar to that of thiram, but the compound contains a zinc atom between the two atoms of sulfur. It is also similar to zineb, which is a zinc/thiocarbamate polymer, and to the manganese/thiocarbamate polymer, maneb. A 96% formulation of ziram used as an industrial biocide was labeled as both an irritant and a sensitizer, but no animal data were available for review. Both ziram-associated cases (83-298, 84-518) reported to the handler database were consistent with an irritant mechanism.
CH3
~N-C-S-Z~S-C-~
CHi
~
~
"
CH3 CH
3
Ziram
Matsushita tested maneb and zineb experimentally with the guinea pig maximization procedure and found both compounds to be potent sensitizers with a high degree of mutual crossreactivity. Concentrations of 5% or more were found irritating
CH2-NH-C-S Mancozeb
\\ S
11.2.3.4 Copper Fungicides Copper compounds are used as both fungicides and antimicrobial agents. Copper has been identified as a sensitizer in the public domain literature based on human case reports (Rademaker, 1998; Verhagen, 1974), and most of the copper fungicides are labeled as potential sensitizers even where there are negative animal sensitization studies (e.g., copper hydroxide). Elemental copper, copper naphthenate, and cuprous oxide are irritant, but not corrosive, in the Draize test. Technical formulations of other compounds (copper ammonium carbonate, copper hydroxide, copper oxide, copper sulfide, and cupric oxide) cause only minimal irritation in animal tests. In the handler database, elemental copper, copper hydroxide, copper naphthenate, and copper sulfate were associated with cases of contact dermatitis consistent with dermal irritation (Table 11.1).
11.2 Review of Use Categories
321
Copper naphthenate was also associated with one case (871724) described as a chemical bum.
11.2.3.5 Fungicides with Miscellaneous Structures Anilazine is a foliar and turf fungicide that has not been registered in California since 1990, but is currently being used elsewhere in the United States. Data on dermal irritation and sensitization from animal studies were not available for review, but the product has been reported as a human sensitizer in tomato harvesters (Schuman and Dobson, 1985; Schuman et aI., 1980) and in lawn care workers (Mathias, 1997). The sample case from the handler database described in Table 11.1 (92-732) was suspected to be caused by an allergic reaction, but a patch test was not carried out.
Carboxin is a systemic fungicide and seed protectant. It is a nonirritant in the Draize test. A mixture containing carboxin (15%), PCNB (15%), and metalaxyl (3.12%) was a nonsensitizer in the Buehler assay.
Chloroneb is a fungicide used for control of seedling diseases in beans, cotton, and soy beans. A formulation containing 30% chloroneb and 3.5% metalaxyl is a category III irritant in the Draize test. The same mixture was a nonsensitizer in the Buehler assay. There were no cases associated with either chloroneb or carboxin in the handler database. Cl
o
OCH 3
*
CH30
I
Chloroneb
Flusilazole is an organosilicon compound, formulated as dry granules and as an emulsifiable concentrate, used for control of ascomycetes and other fungi on cereals, fruits, and vegetables. A 60.7% formulation is an irritant in the Draize test, but a 20% formulation was a nonsensitizer in the Buehler assay. Fosetyl-aluminum is an aluminum salt of an organic acid, used as a systemic fungicide and bactericide. It is active against Oomycetes, Alternaria, and Penicillium on avocado, strawberries, and other crops. It is formulated as a water dispersable granule, as a liquid wettable powder, and as a liquid injectable.
The 80% formulation is a minimal irritant in the Draize test and a nonsensitizer in the Buehler assay. Imazalil (eni1conazole) is a systemic fungicide active against benzimidazole-resistant strains of fungi, formulated as an emulsifiable concentrate, as a water soluble powder, and as a soluble liquid. It is a nonirritant in the Draize test. Although the Buehler assay was negative on a 13.5% formulation, cases of sensitivity to the compound have been previously described in Europe (van Hecke and de Vos, 1983) and in Central America (Penagos, 1993). The technical material is labeled as a potential sensitizer.
There were no cases associated with ftusilazole, fosetylaluminum, or imazalil in the handler database. Iprodione is both a contact and a systemic fungicide used on a broad spectrum of crops. The 41.6% ftowable formulation was a nonsensitizer in the Buehler assay. The 50% formulation was a nonirritant in the Draize test; however, the iprodioneassociated case reported in the handler database (84-897) was compatible with an irritant mechanism following direct contact. CH(CH 3 h
NH\c
If ~CI
i f 'N----'\
Iprodione
~N o
0 Cl
Methylene bis(thiocyanate) is a reactive compound, used as a bactericide in water-cooling systems and pulp and paper mill operations. It is also used as a fungicide in wood preservation products. A 10% formulation was corrosive in the Draize test. The only sensitization study reviewed involved a mixture of methy lene bis( thiocyanate) (0.2%), chlorpyrifos (0.1 %), and TCMTB (0.2%), and did not show any evidence of an allergic
322
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides
response. Nevertheless, a 10% formulation used as an industrial preservative was labeled as a sensitizer.
Elemental Sulfur - transformation? products
S
--()g-
1/2S 2 0 3
",02,H 2 0 '~
NCS-CH 2 -SCN
Methylene bisthiocyanate
Myclobutanil is a systemic fungicide used on grapes and tree fruits. The dermal sensitization data reviewed were not sufficient to determine whether myclobutanil is a skin sensitizer. It caused no erythema or edema in the Draize test, but an accidental direct exposure involving myclobutanil and an adjuvant occluded against the skin without decontamination did prove capable of causing dermatitis (92-1379).
Pentachloronitrobenzene (PCNB) is a soil fungicide and seed treatment formulated as an emulsifiable concentrate, wettable powder, or flowable dust. Commonly treated crops include grain, cotton, cole crops, celery, strawberries, and ornamental flowers. No dermal irritation or sensitization data were available for review. Two reports in the public domain literature indicate that PCNB may be a human sensitizer. Cronin reported on a case seen at the St. John's Hospital in London (Cronin, 1980) and O'Malley described a positive patch test to PCNB as an incidental finding in a study of California nursery workers (O'Malley and Rodriguez, 1998a; O'Malley et aI., 1995). There were no reports involving PCNB in the handler database, as the cases identified in the cited study were not included in the California illness registry.
C I * C I Cl Cl
Cl
N02
S03
~
HS03 ,H 2 S04 _ _
1/2
H20
pH temperature ionic strength
TCMTB (thiocyanomethylthiobenzothiazole) is a compound used as both a fungicide and an antibacterial agent. It is contained in both wood preservation products and industrial preservatives. A 30% formulation was corrosive in the Draize test. A dilute wood preservative formulation (containing < 1% TCMTB) was a nonsensitizer in the Buehler assay, but a 30% formulation of TCMTB used as a bacteriocide, fungicide, and algicide was labeled as a sensitizer. There were no cases associated with TCMTB in the handler database.
ry--Nysr-
s~
~s TCMTB Busan
Thiabendazole is an agricultural fungicide with systemic activity against Fusarium, Penicillium, and other molds. It is formulated as a dust, a wettable powder, and a flowable dust. The 98.5% formulation is a minimal irritant in the Draize test, and no cases were associated with thiabendazole in the handler database. There were no sensitization studies done on thiabendazole per se, but the Buehler test was negative on a mixture of thiabendazole, captan, and pentachloronitrobenzene.
~---zJ N~N~ ~C') Thiabendazole S
PCNB
The large number of cases associated with elemental sulfur in California agriculture is striking and would seem to imply that sulfur is a potent skin irritant. Scattered cases have also been reported in applicators and field workers in the state of Washington. However, a 98% formulation caused only minimal irritation in the Draize test and was negative in the Buehler test. Other work has shown that a 25% concentration of wettable powder produced a 2+ irritant reaction on intradermal injection and was a sensitizer in the guinea pig maximization test (Matsushita et aI., 1977). Elemental sulfur has also been reported as a human allergen (Gregorczyk and Swieboda, 1968; O'Malley and Rodriguez, 1998b; O'Malley et aI., 1995; Schneider, 1978; Wilkinson, 1975). It is not clear which sulfur products (figure) may be responsible for its reported effects on the skin. The issue is explored more fully in the chapter on sulfur in this volume.
Triadimefon is a systemic fungicide for control of powdery mildew on cereals, deciduous fruit, and grapes. The 50% formulation is a minimal irritant in the Draize test. The 95% technical material is a sensitizer in the Buehler assay. The same test on a more dilute formulation was negative. Case 82-1412 in the handler database was a suspected allergic reaction to triadimefon.
Vinclozolin is a fungicide used for control of Botrytis, Sclerotinia, and Monilia on grapes, strawberries, and other crops.
11.2 Review of Use Categories Cl
It is a minimal irritant in the Draize test. It is labeled as a sen-
I
)0;(O~CH I
CH3
Cl
Vinclozolin
2
3
11.2.4 FUMIGANTS AND BIOCIDES Aluminum phosphide is a fumigant formulated as solid tablets that release phosphine gas on contact with air and water. It is used for both commodity fumigation and rodent control. Animal sensitization and irritation data were not available for review. Dermatitis cases have occurred in handlers following application (84-2184) and contact with partially spent dust. H20,02 AlP - - - PH3 PH3 ~
Cl
H-C-C=C I I ' H H H
sitizer, but no animal test data were available for review. The handler database did not contain any cases associated with the use of vinclozolin.
CI>rL)--o
/
323
1,3 Dichloropropene
Ethylene oxide is a commodity fumigant used in food processing and in hospital sterilization equipment. No dermal irritation or sensitization study was available for review; however, some ethylene oxide products are labeled as dermal sensitizers. Numerous case reports have described allergic contact dermatitis in hospital workers handling rubber products and other medical supplies sterilized with ethylene oxide (Alomar et aI., 1981; Alomar and Gimenez Camarasa, 1981; Fisher, 1988; Hanifin, 1971; Romaguera and GrimaIt, 1980; Romaguera et aI., 1977; Romaguera and Vilaplana, 1998; Taylor, 1977). The case described in the handler database (87-2720) involved a chemical bum following accidental direct contact with ethylene oxide gas. Fosthiozate is a compound that has limited use as a soil nematicide for root vegetable crops. It causes only minimal irritation in the Draize test, but the technical material is a sensitizer in the guinea pig maximization test. There were no cases associated with fosthiozate in the handler database.
+ AI(OHh
P04 + H20
Dazomet is a fumigant that releases irritant amines and mercaptans as it breaks down in soil. It is used as a preplanting treatment for ornamental beds and nurseries. The technical material causes minimal irritation in the Draize test, but most of the dazomet products reviewed (20-24% active ingredients) are all labeled as skin irritants. No dermal sensitization studies were available for review, but the compound has been reported as a sensitizer in the public literature (Black, 1973), and a series of cases of irritant contact dermatitis associated with dazomet has recently been reported from France (Gamier et aI., 1993). The latter cases were similar to a case reported to the handler database that occurred on contact of dazomet with skin covered with sweat (case 90-2448).
S==lSj CH3-N~N-CH3
Dazomet
Dichloropropene is a fumigant used principally for soil sterilization in the production of root vegetables and other crops. It is corrosive in the Draize tests. No animal sensitization studies were available for review, but the products registered in California are registered as sensitizers. A case reported by Nater and Gooskens describes an allergic reaction to a mixture of dichloropropane and dichloropropene, identifying dichloropropene as the most likely allergen (Nater and Gooskens, 1976). The 11 cases in the handler database (Table 11.1) appear consistent with a simple irritant mechanism (see case 88-2091).
Fosthiozate
Methyl bromide is a volatile fumigant used as a structural, soil, and commodity fumigant. It is corrosive in the Draize test, but no dermal sensitization data were available for review. The numerous cases in the handler database tended to involve prolonged occlusion of methyl bromide against the skin, as illustrated in Table 11.1 (82-32). Metam sodium, a soil fumigant and nematocide, is also effective against weeds and soil fungi. The reaction of metam with water produces methylisothiocyanate (MITC), carbon disulfide, hydrogen sulfide, and methyl amine. A 42% formulation was corrosive in the Draize test. In the Buehler assay, there was so much irritation present in the induction phase of the test that it was difficult to distinguish between irritant and allergic reaction. S
H2 0
CH3NH-~ Na+. 2 H20 _
"s
CH3N=C=S MITC
+ CH 3-NH 2 Methylamine
+ CS2
Carbon disulfide
+
CH 3N=C=O
MIC - up to 4% of the level of MITe
+ H2S Hydrogen disulflde
Cases of contact dermatitis associated with metam sodium have been reported in several jurisdictions around the world. In Germany, the cases stemmed from use of metam in the production of root vegetables (lung, 1975; Jung and Wolff, 1970a,
324
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides
1970b; Wolff and Jung, 1970). Cases of dennatitis were also reported from workers wading into the Sacramento river to clean up metam-sodium spilled into the river following a train derailment near Dunsmuir, California in July, 1991 (Koo et al., 1995). Between 1991 and 1995, one case of dennatitis associated with this compound was reported in a Washington applicator (O'Malley, 1997). There were 34 cases in the California handler database.
Acetochlor is not currently registered in California and consequently no animal sensitization or irritation data were available for review. However, the product infonnation in trade literature (Meister, 1995) indicates that acetochlor is a category 11 skin irri tan t. C H3
~<:~OH::~ CH 2CH3 Acetochlor
11.2.5 HERBICIDES
CH2CH3 O-N:CH20CH3
~
11.2.5.1 Bipyridyls Diquat is a contact herbicide and desiccant used for potato vines and seed crops, and for industrial and aquatic weed control. No animal or dermal sensitization study was reviewed for this compound. Fifteen cases of contact dennatitis associated with handling diquat were included in the handler database. The illustrative case described in Table 11.1 involved failure to decontaminate shoes after they became soaked with diquat and subsequent prolonged contact (see case 86-1498).
Diquat
Paraquat is a contact herbicide and dessicant used to control weeds on a variety of grain, vegetable, and fruit crops. The 37% technical material is a category III skin irritant in the Draize test. It is a nonsensitizer in the Buehler assay. There were 17 cases associated with paraquat in the California handler database. The case described in Table 11.1 involved a mild reaction following direct contact with a dilute paraquat spray (83-480). Skin injury associated with application of paraquat and with its misuse has been reported from many parts of the world (Angelo et al., 1986; Botella et al., 1985; Cooper et al., 1994; Gamier et al., 1994; George, 1989; Horiuchi and Ando, 1980; Howard, 1979; Li, 1986; Peachey, 1981; Sugaya, 1976; Swan, 1969; Vilaplana et al., 1993; Villa et al., 1995).
CH~~O)
·2CI
Paraquat dichloride
COCH2CI CH 2CH3
Alachlor
CH3 CH3
~N:::::2~IOCH3 CH2CH3 Metolachlor
Alachlor is labeled as a corrosive. A dennaI irritation study on a 45% emulsifiable formulation showed irritation persisting to the conclusion of the study (at 72 hours). The emulsifiable concentrate is also labeled as a sensitizer. Contact sensitization has been reported in the public domain literature by Won et al. (1993). The case reported in the handler database also involved a suspected allergic reaction (84-537). The Draize test demonstrates that an 85.1 % fonnulation of metolachlor is a moderate (category Ill) dennal irritant. A formulation containing 79% metolachlor also causes sensitization in the Buehler assay. 11.2.5.3 Nitroaniline Compounds Oryzalin is a selective pre-emergent herbicide used for control of annual grasses and broadleafs on fruit trees, vineyards, nuts, turf, and ornamentals. There was no dennal sensitization study available for review. No dermal irritation study was available for review, but the acute dermal toxicity study indicated that an 85% fonnulation did not cause significant skin irritation. However, the cases listed in the handler database associated with direct exposure to oryzalin were consistent with an irritant mechanism (84-51 and 84-272).
.';~~ S02NH" Oryzalin
11.2.5.2 Chloracetanilides Acetanilides are selective pre-emergence herbicides used for control of annual grasses and broadeafweeds in cabbage, citrus, orchard crops, and grapes. The skin effects of these compounds are quite similar, reflecting the high degree of similarity in their chemical structures.
Pendimethalin is a pre-emergent and early postemergent herbicide for control of annual grasses and broadleafs in corn, sorghum, soybeans, nursery crops, and is mixed with fertilizer in weed and feed lawncare products. A 2.45% fonnulation caused minimal irritation in the Draize test. A "weed and feed" fonnulation containing less than 1% pendimethalin was a nonsensitizer in the Buehler assay. Typical cases associated with
11.2 Review of Use Categories
pendimethalin in the handler database appear consistent with an irritant mechanism.
«2Ho
NO~~g:H5
~CKl CKl Pendimethalin
Benefin is a selective pre-emergent herbicide that controls annual grasses and broadleaf weeds in alfalfa and in turf. It does not affect established weeds. The technial material is a moderate skin irritant (category 11) in the Draize test. A combined formulation with oryzalin showed no sensitization in the Buehler assay. There were no benefin cases reported in the handler database. Ethalfluralin is used on beans, watermelons, sunflowers, cantaloupes, and cucumbers. No dermal irritation study was available for review, but dermal irritation was observed during an acute dermal toxicity study and equated to approximately category 11 in the Draize test. A 31.6% emulsifiable concentrate was a sensitizer in the Buehler assay. The ethalfluralin case reported in the handler database was consistent with an irritant mechanism (90-1832). Trifluralin is a selective pre-emergent herbicide. Its use profile is broader than that of the other trifluorotoluene herbicides, including numerous fruit and vegetable crops. The 80% dry flowable formulation causes minimal irritation in the Draize test. The same formulation is labeled as a dermal sensitizer. Nevertheless, Buehler sensitization studies were negative on other formulations (Trilin lOG, Trilin GRP). The trifluralin case listed in the handler database occurred following accidental direct contact and was compatible with an irritant mechanism (93-340).
325
defoliants, that are weak cholinesterase inhibitors, approximately comparable to malathion (Hayes, 1982). Bensulide is used for pre-emergent control of annual grasses and broadleaf weeds in turf and in a variety of vegetable crops. It causes minimal irritation in the Draize test, but a case reported to the handler database (83-1770) was consistent with an irritant mechanism. The animal sensitization study on file did not contain sufficient information to conclude whether or not bensulide is a dermal sensitizer.
Glyphosate is a nonselective herbicide with extensive agricultural and nonagricultural uses. The technical material is a nonirritant in the Draize test, but the formulated glyphosate products contain a surfactant that may produce some irritation on direct contact, as illustrated by the 70 cases reported to the handler database (Table 11.1). Glyphosate formulations are probably infrequent sources of dermatitis, given their volume of use, and have been determined experimentally to be no more irritating than detergents contained in topical shampoos (Maibach, 1986). Glyphosate is a nonsensitizer in the Buehler assay, but a photoallergic reaction to an isothiazolin preservative present in a glyphosate formulation was reported by Hindson and Diffey (l984a, 1984b).
o
0
11
11
HO-C-CH2--NI-t-C~-~-OH
Glyphosate
OH
Sulfosate is a nonselective herbicide with a range of uses similar to glyphosate. The 62% technical material is a nonirritant in the Draize test and is a weak sensitizer in the Buehler assay. There were no cases associated with sulfosate in the handler database.
o
CH3
OH
C~
I _ I HO-CO-CH2-NH-CH2-~-O 1-c~
Sulfosate
Tribufos (Fo1ex®) is a phosphotrithioate used as a defoliant to minimize boIl rot in cotton and to prepare the plants for mechanical harvesting. The 70.6% emulsifiable concentrate is labeled as a corrosive, but the technical material is a moderate irritant (category III) in the Draize test. The sensitization study reviewed did not contain sufficient information to determine whether or not tribufos is an allergen in the Buehler assay. No 11.2.5.4 Organophosphates
Although most important organophosphates are insecticides, the group includes several herbicidal compounds that do not inhibit cholinesterase to any significant degree, and two phosphorothioate compounds (butifos and merphos), used as cotton
C4 HS--S
:::p
C4 Hg-S C4 Hg-- S
Merphos (Folex)
150-50-6
C4 HS--S
C4 Hg-S :::p=o C4 Hg-- S DEF (tributos) 78-48-8
326
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides CH
tribufos cases were listed in the handler database. The oxidation product of merphos, DEF, is also used as a cotton defoliant.
I
CH2
11.2.5.5 Phenoxy Herbicides
3
/CO~CO-SCH
CH S
2,4-D and 2,4,5-T are the prototype phenoxy compounds, herbicides used for selective control of broadleaf weeds. The latter has been removed from the market since the later 1970s because of the concern about contamination with 2,3,7,8 TCDD resulting from the manufacture of the trichlorophenol component of 2,4,5-T. Chloracne, one of the principal clinical effects of 2,4,5-T, is discussed elsewhere in this volume.
CH3
3'-CH
~Q~
CF3
N
3
CF2H
Dithiopyr
90% formulation is a nonirritant. The 22.9% formulation is a nonsensitizer in the Buehler test. There were no cases associated with imazethapyr in the handler database.
Imazethapyr
Triclopyr is a systemic herbicide for control of woody plants, broadleaf weeds, forests, turf, and industrial sites. A 44% formulation of a triethylamine salt showed no sensitization in the Buehler assay, and a 61.6% formulation was a minimal irritant in the Draize test. The cases associated with triclopyr in the handler database were consistent with an irritant mechanism (88-2832).
MCPP
CH 3 ~
~N?-o----U-o-b -~-O-C2H5 ~ .. I
C I V CI
CI~O
CI~\!.J-OCH-COOH N Triclopyr
H Fenoxaprop Ethyl
2,4-D is manufactured by a different process and does not contain the same dioxin contaminants. The technical material is nonirritating in the Draize test, but the phenoxy compound fenoxapropethyl is a mild to moderate irritant. The results of animal sensitization tests on a mixture of 2,4-D dimethylamine salt, MCPP, and dichlorprop, an analog of fenoxaprop, did not clearly distinguish between irritant- and sensitization-related dermal reactions. There were no adequate dermal sensitization studies on file for fenoxaprop ethyl. There were no cases associated with any of the phenoxy herbicides in the pesticide handler database. Acute bullous dermatitis in German forestry workers handling a mixture of 2,4-D and 2,4,5-T was reported by J ung, including one case with contact allergy documented by patch testing with 0.4% dilution of the herbicide mixture (Jung and Wolf, 1975).
11.2.5.7 Thiocarbamates and Carbamates Molinate is a selective herbicide formulated as a liquid and granule to control watergrass in rice. The material was a nonsensitizer in the Buehler assay. A 15% formulation was found to cause minimal irritation in the Draize test, but causes mild irritation on wet skin. The molinate-associated case in the handler database was consistent with an irritant mechanism.
00 11
C2H5S-C-N Molinate
Thiobencarb is a pre-emergent and early postemergent herbicide for control of grasses and broadleaf weeds on rice fields. The 84% material is a minimal skin irritant.
11.2.5.6 Pyridine Derivatives Dithiopyr is an herbicide used for control of annual and smallseeded broadleafweeds in turf. The 91.5% technical material is a minimal irritant in the Draize test, but the 12.7% emulsifiable concentrate is labeled as a skin irritant and is a sensitizer in the Buehler assay. There were no cases associated with the use of dithiopyr in the handler database. Imazethapyr is an imidazolinone and pyridine compound, used for selective control of broadleaf weeds and grasses. The
11.2.5.8 Triazines Atrazine is a selective herbicide used in grain crops and turf. There are no data on animal sensitization and irritation, and no cases in the handler database. Cyanazine is a selective herbicide used principally on cotton in California, and on corn elsewhere in the United States. The technical material causes minimal irritation in the Draize test. Buehler studies conducted on mixtures containing metolachlor
11.2 Review of Use Categories
327
Chlorsulfuron is a selective broadleaf herbicide used on wheat and other grain crops, and is sold as a dry flowable formulation. It causes minimal irritation in the Draize test. There were no cases associated with its use in the handler database.
and cyanazine do not demonstrate sensitization. There were no cases associated with cyanazine in the handler database.
o
N~
OCH 3
o-SOrNHb-N~O N ~ _.
Cl
Chlorsulfuron
N===(
CH 3
Diuron is used for emerging broadleaf weeds and grasses in alfalfa, cotton, fruit, wheat, and vineyards. An 81 % formulation causes mild to moderate (category Ill) dermal irritation in the Draize test, but is a nonsensitizer in the Buehler assay. There were no cases associated with its use in the handler database. Prometryn is a selective herbicide with a spectrum of use similar to other triazines. The technical material is a minimal irritant in the Draize test, and a 45% formulation caused minimal reaction in the Buehler test. There were no prometrynassociated cases in the handler database. Thidiazuron is a plant growth regulator and defoliant used on cotton. A mixture of 12% thidiazuron and 6% diuron was corrosive in the Draize test. The same mixture also caused sensitization in the Buehler assay. There were no cases associated with its use in the handler database. N
Simazine is a selective herbicide used for control of broadleaf weeds and annual grasses in corn, established alfalfa, fruit, nuts, asparagus, ornamentals, and turf. A 90% formulation causes minimal irritation in the Draize test, and a 6.3% formulation is a nonsensitizer in the maximization test. The single case reported to the handler database involved contamination of the hands with simazine while mixing the material, and a secondary dermatitis of the genitalia (see case 83-2394). Cl
W,lN
0
C2 H,-NH---l... )-NH-C 2 H5 N
Simazine
0
~=:J-NH-8-N~ Thidiazuron
11.2.5.10 Herbicides of Miscellaneous Structure Dichlobenil is used for selective weed control in growing cranberries, ornamental flowers, orchard fruit, vineyards, and turf. It is available in wettable powder and granular formulations. It is a nonirritant in the Draize test, but it is a sensitizer in the maximization assay. The manufacture and formulation of dichlobenil was also associated with an outbreak of chloracne during the 1970s. This was perhaps due to a manufacturing contaminant that remained unidentified. Limited use of dichlobenil is reported in California, and there were no cases associated with its use included in the handler database.
11.2.5.9 Urea Herbicides Bromacil is a perennial grass, general weed, and brush control agent for noncrop areas. It is a nonirritant in the Draize test and a nonsensitizer in the Buehler assay. There were no cases associated with bromacil in the handler database.
Bromacil
CI~CI Dichlobenil
Flumetsulam (triazolopyrimidine sulfonanilide) is an herbicide used only for corn and soybeans. The only irritation and sensitization studies involved a mixture of metolachlor (79.9%) and flumetsulam (2.6%). The mixture was a category 11 irritant in the Draize test and a sensitizer in the Buehler assay. There
328
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides Stephan C-65 is a spray adjuvant contained in a commercial mixture with aromatic hydrocarbons (35%), phosphate ester of polyoxyalkylated fatty alcohol, and oleic acid. The mixture is labeled as a skin irritant, but no Draize study was available for review. The product was a nonsensitizer in the Buehler assay.
were no cases associated with the use of flumetsulam in the handler database. Isoxaben is a pre-emergent herbicide applied to the soil surface to control annual broadleafweeds in wheat and other grain crops. The technical material is a nonirritant in the Draize test and also a nonsensitizer in the Buehler assay. There were no isoxaben cases in the handler database.
Sethoxydim is a systemic postemergent herbicide for control of grasses in soybeans, peanuts, vegetables, and nursery ornamentals. The 13% formulation is a category 11 skin irritant, but demonstrated no sensitization in the maximization test. The sethoxydim-associated case in the handler database (88-1253) was consistent with an irritant mechanism.
11.3 PROTECTION OF PESTICIDE HANDLERS Protection of pesticide handlers from dermal exposure allows prevention of systemic as well as topical effects. In California, measures for minimizing exposure include requirements for mixing and loading corrosive agricultural materials (and other category I materials) in closed systems or using watersoluble packaging to reduce dust exposure. California Code of Regulations (Title 3. Food and Agriculture) Division 6. Pesticides and Pest Control Operations Chapter 3. Pest Control Operations Subchapter 3. Pesticide Worker Safety Section 6746. Closed Systems: Employers shall provide closed systems for employees who mix or load liquid pesticides in toxicity category one, or load diluted liquid mixes derived from dry pesticides in toxicity category one, for the production of an agricultural commodity. No employee shall be permitted to transfer, mix, or load these pesticides except through a closed system. The system's design and construction shall meet the director's closed-system criteria. In all V.S. jurisdictions, label recommendations for protective equipment have the force of legal requirements. For a 37% formulation of paraquat, for example, these include:
11.2.5.11 Adjuvants Adjuvants are an important component of many applications but are the focus of less attention than the active ingredients. The most commonly used adjuvants are spreading and sticking agents with chemical structures similar to detergents. Some are derivatives of simple long chain fatty acids (alkyl amino3-aminopropane hydroxy acetate alkyl derived from coconut oil fatty acids), and others are more complex synthetic molecules (e.g., alkyl aryl polyalkoxylated alcohols). Adjuvants not sold as stand-alone products are sometimes included as ingredients of formulated herbicides or insecticides (see discussion of glyphosate). Data on two adjuvants were available for this review. Neither product was included as an identifiable ingredient in the California illness surveillance data files. There were consequently no cases associated with this material in the handler database. Polymerized pinene is a spreading/sticking agent used for pesticide applications on turf. A Draize study conducted on this product showed mild irritation, but was not carried out for a sufficient length of time to adequately characterize the category. A human repeated insult test carried out on volunteer subjects showed no evidence of sensitization.
Applicators: Long-sleeved shirt and long pants Waterproof gloves Shoes plus socks Mixers and loaders: Long-sleeved shirt and long pants Shoes plus socks Waterproof gloves Face shield Chemical-resistant apron For a 30% formulation of propargite, all handlers are required to wear: Long-sleeved shirt and long pants; shoes plus socks; protective eyeware.
11.4 RESPONSE TO OUTBREAKS OF FIELD WORKER DERMATITIS Field workers are protected by means of "safe-entry waiting periods" or reentry intervals. Much of the regulatory efforts aimed at preventing the occurrence of illnesses in field workers has involved evaluation of the appropriateness of existing reentry intervals in the context of illness outbreak investigations.
11.4 Response to Outbreaks of Field Worker Dennatitis
Unlike pesticide handlers, field workers have indirect contact with pesticides, most commonly from contact with residues on foliage. The degree of exposure to pesticide residues depends upon the initial application rate, the half-life of residue dissipation, the nature of the crop, and the nature and timing of the work performed. The occurrence of skin reactions further depends upon the properties of the individual pesticides, to the extent these are established. For example, elemental sulfur is applied frequently to table grapes (every two weeks), and residues sometimes exceed 10 ).lg/cm2. This means that a large percentage of skin injuries occurring in grapes workers are associated with some exposure to sulfur. Single case reports have several potential explanations that require some effort to differentiate. Some cases might be allergic reactions unique to a single crew member; some might be sentinel cases signifying the occurrence of an otherwise unrecognized outbreak; and some cases might prove to be non-work-related skin conditions not expected, in most circumstances, to occur in co-workers. In investigating outbreaks, it is usually obvious whether the problem is work related. The central questions are whether the reported episode was related to pesticides, and which material, among those reported, was principally responsible. The variation in residue dissipation is illustrated by data on propargite. Residue studies (K Maddy et aI., 1977; 1979) performed in a coastal area of California showed 1- to 2-day dissipation half-lives. Residue studies in California's Central Valley typically showed half-lives of 5-7 days (M Reeve et aI., 1991), but some fields showed half-lives up to 11 days. Dissipation half-lives as long as 30 days have been measured in the context of outbreak investigations (O'Malley, 1998; O'Malley et aI., 1989; Smith, 1991b). In 1987, an outbreak of dermatitis occurred among workers turning cane, in a vineyard near Bakersfield, California; residue data revealed a history of applications of both sulfur and propargite. Residue data demonstrated foliar residues of sulfur ranging from 1.87 to 23.24 ).lg/cm2 and foliar residues of propargite ranging from 0.08 to 0.87 ).lg/cm2. Although a review of animal irritation studies demonstrated that the direct irritant capacity of propargite greatly exceeds that of sulfur, the levels of sulfur greatly exceeded the levels of propargite. It was not possible to conclude which compound may have been the principal cause of the outbreak. No change was therefore made in the 7-day reentry interval on propargite or the "spray dries" interval in place for sulfur. In 1988, an outbreak of dermatitis occurred among crews of nectarine harvesters in Tulare County (Fig. 11.1). Examination of a comparison group of workers who had no dermatitis allowed an analysis of work history and residue history. This showed a strong correlation between cumulative exposure to propargite and the occurrence of dermatitis (Fig. 11.2). A review of the work history for the group with no dermatitis showed their peak exposure to propargite residues of 0.2 ).lg/cm2. This value was used as an estimated no-ob servedeffect residue-level (NOERL) for purposes of determining a safe-reentry interval. The reentry interval for harvesting
329
5
:g4 Ul
rn
()3 '0
0>2
.0
E ~1
15161718192021222324252627282930 Date during June, 1988
D
Crew 1
•
Crew2
III
Crew3
Figure 11.1 Dennatitis in three crews of nectarine harvesters-June, 1998. Reprinted with pennission from Han1ey and Belfus, State of the Art Reviews in Occupational Medicine.
tree fruit was lengthened to 21 days following the episode (O'Malley et aI., 1990). In August, 1995, the Department of Pesticide Regulation (DPR), Worker Health and Safety (WHS) Branch, received a report regarding an outbreak of dermatitis (sunburnlike erythema on the chest, neck, arms, and face) among crews of workers performing hand labor activities on a table grape ranch in northern Fresno County near Kerman, California. Of 202 workers (8 crews) lifting grape foliage over vine guide wires (turning cane), 65 (32.2%) sought treatment between August 9 and August 20. No rashes were reported among 54 workers (2 crews) working in the same vineyards pulling leaves. The incidence of workers seeking treatment varied from 3.7% in the crew apparently least affected to 87.5% in the crew with the largest number of reported cases. Because the crews left the Fresno area, it was not possible to determine whether the variation in reported cases was attributable to differences in the occurrence of rash or to differences in the likelihood of seeking treatment following onset of rash. Rash incidence/10 employed workers
10
8
6 4 2
o
o
10 20 30
40 50 60
70
Propargite-Residue Hours
Figure 11.2 Correlation between cumulative exposure to propargite and the occurrence of dennatitis.
330
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides
Crew Number
•• • • •
•
11.5 RESIDUE PROBLEMS IN OTHER JURISDICTIONS
35 -
30 ...,
~ 25
!::o 20 -
.! 15 -4
8 9
Total
910 -
Z
54
0-- T 8/8195
8/9/95
8/10/95
8/11/95
Date of Reported Onset
Figure 11.3
Crews affected by dermatitis and dates of reported onset.
The crews affected by dermatitis worked between August 7 and August 11 in blocks of Thompson seedless and Red Globe grapes (Fig. 11.3). Application records showed that these blocks had received treatments with propargite, Britz Buffer, glyphosate, methomyl, dichloronitroaniline, myclobutanil, Latron B 1956, gibberelin, triflumizole, elemental sulfur, and iprodione. At site 304, three dislodgeable foliar residue (DFR) samples were taken 8/15/95 by DPR Pesticide Use Enforcement (PUE)lFresno County Agricultural Commissioner (CAC). Single samples were also taken at site 301 and site 302. Propargite levels on 8/15/95 ranged from 0.37 to 0.66 ug/cm 2. Detectable residue was also found for elemental sulfur (0.16-1.1 !l-g/cm2), iprodione (non-detect [nd]-0.16 !l-g/cm2), myclobutani1 (nd0.16 !l-g/cm2), and dich10ronitroaniline (nd-0.045 !l-g/cm2). At site 301, the sulfur residue (1.1 !l-g/cm2) exceeded the residue of propargite (0.66 !l-g/cm2), but at other sites the sulfur residue was <0.21 !l-g/cm2. Because animal studies had demonstrated that propargite was a more severe dermal irritant than the other compounds identified in the residue samples, it was considered to be the probable cause of the outbreak. The levels of dislodgeable propargite residue were well above the estimated 0.2 !l-g/cm2 (NOERL) for repeated dermal exposure to propargite in nectarine harvesters. Propargite dissipation half-lives observed in follow-up monitoring studies showed residue half-lives of 9.414.4 days. Although some blocks of Thompson seedless grapes on the ranch received two applications of propargite within a IQ-day interval, the residue levels on this site (0.55 !l-g/cm2) did not exceed the highest levels (0.66 !l-g/cm2) found on blocks receiving only a single treatment. Slow dissipation of propargite, comparable to that seen in prior episodes of dermatitis associated with this compound, appeared to be the principal reason for this outbreak. Possible measures to prevent recurrent episodes include prohibition of propargite use prior to high-contact work activities and routine monitoring of propargite dissipation prior to reentry. Additional studies might also evaluate whether dissipation of propargite is affected by spray adjuvants or other materials in the application tank mixes (O'Malley, 1998).
Although the problems observed with propargite in California have not been reported extensively from other jurisdictions, the dry environment in the Central Valley of California is not unique. One must assume that the lack of reports is due to the lack of illness surveillance. The few problems documented include an outbreak of dermatitis among tomato harvesters in Tennessee associated with exposure to residues of anilazine (Schuman and Dobson, 1985; Schuman et aI., 1980), which was reported in 1980. An episode of dermatitis associated with ziram in workers picking Bosc pears was reported to the Washington Department of Health (DOH) in September 1990. The cases occurred despite compliance with the existing reentry interval, which required waiting until the spray dried in the treated field. The rashes involved arms, hands, faces, and abdomens. The only workers affected were those picking Bosc pears. No rashes appeared in other workers harvesting D' Anjou pears and Golden Delicious apples. The findings were consistent with an interaction between skin abrasions that commonly occur while picking the Bosc variety and the exposure to the thiocarbamate fungicide. Half-life data for ziram on pears was estimated at approximately 9 days based on dissipation studies from California, New York, and Washington. The reentry interval was lengthened to 14 days, representing approximately 1.5 halflives, for most formulations of ziram on Bosc pears (O'Malley, 1997).
11.6 CONCLUSIONS Skin effects of pesticides depend upon the nature of the chemical and the specific circumstances of exposure. Corrosive materials include halogenated fumigants, bipiridyl herbicides, isothiazolin herbicides, and individual compounds such as propargite. The occurrence of allergic responses is less predictable, but some compounds, such as ani1azine, provoke a high incidence of sensitization. Animal data provide a useful cross-check on data from human use experience for both irritation and sensitization. Applicators and other handlers are at highest risk of contact with high concentrations of pesticides. Dermatitis in field workers typically occurs in clusters of cases, through indirect contact with foliar residues. Protection of field workers by means of reentry intervals or mandatory postapplication waiting periods requires frequent reevaluation as new episodes of dermatitis are reported.
REFERENCES Adams, R., and Fischer, T. (1990). Diagnostic patch testing. In "Occupational Skin Disease" (R. Adams, ed.), 2nd ed., pp. 223-253. WB Saunders, Philadelphia. Adams, R. M., and Manchester, R. D. (1982). Allergic contact dermatitis to Maneb in a housewife. Contact Dermatitis 8, 271.
References
Alomar, A, Camarasa, J. M., Noguera, J., and Aspinolea, E (1981). Ethylene oxide dermatitis. Contact Dermatitis 7, 205-207. Alomar, A, and Gimenez Camarasa, J. M. (1981). Dermatitis caused by ethylene oxide. Med. Cutan. Ibero Lat. Am. 9,293-296. Ancona, A, Suarez de la Torre, R, and Macotela, E. (1985). AIIergic contact dermatitis from povidone-iodine. Contact Dermatitis 13, 66-68. Angelo, c., Ruatti, P., and Comba, P. (1986). A case of acne in a subject exposed to paraquat. Med. Lav. 77, 247-249. Ashby, 1., Basketter, D. A., Paton, D., and Kimber, I. (1995). Structure activity relationships in skin sensitization using the murine local lymph node assay. Toxicology 103,177-194. Bashir, S. J., and Maibach, H. I. (2000). Methods for testing the irritation and sensitization potential of agricultural chemicals. In "Pesticide Dermatoses" (H. Penagos, M. O'Malley, and H. L Maibach, eds.). CRC Press, Boca Raton, Florida. Black, H. (1973). Dazomet and chloropicrin. Contact Dermatitis Newsletter 14, 410-411. Botella, R, Sastre, A, and Castells, A. (1985). Contact dermatitis to paraquat. Contact Dermatitis 13, 123-124. Bruynzee1, D. P. (1991). Contact sensitivity to lannate. Contact Dermatitis 25, 60-61. Bruze, M., and Gruvberger, B. (1988). Formaldehyde-induced depression of skin reactivity to 5-ch10ro-2-methyl-4-isothiazolin-3-one in the guinea pig. Contact Dermatitis 19, 231-232. Cagen, S. Z., Malley, L. A., Parker, C. M., Gardiner, T H., Van Gelder, G. A., and Jud, V. A (1984). Pyrethroid-mediated skin sensory stimulation characterized by a new behavioral paradigm. Toxico!. App!. Pharmaco!' 76, 270-279. California Department of Pesticide Regulation (1998). "Product Label Data Base."
Camarasa, G. (1975). Difolatan dermatitis. Contact Dermatitis 1, 127. Camarasa, J. G. (1986). Contact dermatitis due to Bronopol. Contact Dermatitis 14, 191-192. CoIlins, A., Nichol, A., and Elsbury, S. (1982). Porphyria cutanea tarda and agricultural pesticides. Aust. J. Derm. 23, 70-75. Cooper, S. P., Downs, T, Burau, K., Buffier, P. A., Tucker, S., Whitehead, L., Wood, S., Delc1os, G., Huang, B., Davidson, T, et a!. (1994). A survey of actinic keratoses among paraquat production workers and a nonexposed friend reference group. Am. J. Ind. Med. 25, 335-347. Cottel, W. (1972). Difolatan. Contact Dermatitis Newsletter 11,252. Cronin, E. (1980). Pesticides. In "Contact Dermatitis," pp. 393-414. Churchill Livingstone, Edinburgh. Edrnundson, W. E, and Davies, J. E. (1967). Occupational dermatitis from naled. A clinical report. Arch. Environ. Health 15, 89-91. Emmett, E. A., Ng, S. K., Levy, M. A., Moss, J. N., and Morici, L J. (1989). The irritancy and aIIergenicity of 2-n-octyl-4-isothiazolin-3-one (Skane M8), with recommendations for patch test concentration. Contact Dermatitis 20,21-26. Environmental Protection Agency (EPA) (1984). Pesticide Assessment Guidelines Subdivision F, Hazard Evaluation: Human and Domestic Animals. V.S. Environmental Protection Agency, Washington, DC. Erdrnann, S., Hertl, M., and Merk, H. E (1999). AIIergic contact dermatitis from povidone-iodine. Contact Dermatitis 40, 331-332. Everhart, L. P., and Holt, R E (1982). Potential benlate fungicide exposure during mixerlIoader operations, crop harvest, and home use. J. Agric. Food Chem. 30,222-227. Farkas, J. (1983). Irritative contact dermatitis to scabicides as a sort of postscabies dermatitis. Derm. Beruf Umwelt. 31, 189-190. Fisher, A. A. (1988). Burns of the hands due to ethylene oxide used to sterilize gloves. Cutis 42, 267-268. Flannigan, S. A, and Tucker, S. B. (1985). Variation in cutaneous perfusion due to synthetic pyrethroid exposure. Br. J. Ind. Med. 42,773-776. Flannigan, S. A., Tucker, S. B., Key, M. M., Ross, C. E., Fairchild, n, E. J., Grimes, B. A., and Harrist, R B. (l985a). Primary irritant contact dermatitis from synthetic pyrethroid insecticide exposure. Arch. Toxico!. 56, 288-294.
331
Flannigan, S. A, Tucker, S. B., Key, M. M., Ross, C. E., Fairchild, LE., Grimes, B. A., and Harrist, R B. (l985b). Synthetic pyrethroid insecticides: A dermatological evaluation. Br. J. Ind. Med. 42, 363-372. Foussereau, J., Brandle, L, and Boujnah-Khouadja, A. (1984). Allergic contact eczema caused by isothiazolin-3-one derivatives. Derm. Beruf Umwelt. 32, 208-211. Fregert, S. (1968). Allergic contact dermatitis from the pesticides captan and phaltan. Contact Dermatitis Newsletter 1, I!. Freundt, K. J., and Netz, H. (1977). Behavior of blood acetaldehyde in aIcoholtreated rats following administration ofthiurams. Arzneimittelforschung 27, 105-108. Frosch, P. J., White, L R, Rycroft, R. J., Lahti, A., Burrows, D., Camarasa, J. G., Ducombs, G., and Wilkinson, J. D. (1990). Contact allergy to Bronopol. Contact Dermatitis 22, 24-26. Gammeltoft, M. (1978). Tributyltinoxide is not allergenic. Contact Dermatitis 4,238-239. Gammon, D., and Casida, J. E. (1983). Pyrethroids of the most potent class antagonize GABA action at the crayfish neuromuscular junction. Neurosci. Lett. 40, 163-168. Gammon, D. W. (1985). Correlations between in vitro and in vivo mechanisms of pyrethroid insecticide action. Fundam. Appl. Toxico!. 5, 9-23. Gamier, R, Chataigner, D., Efthymiou, M. L., Moraillon, L, and Bramary, E (1994). Paraquat poisoning by skin absorption report of two cases. Veterinary Human Toxico!. 36, 313-315. Gamier, R, Prince, c., Reygagne, A, Azoyan, P., Dally, S., and Efthymiou, M. L. (1993). Contact dermatitis from dazomet: Seven cases. Archives des maladies professionnelles et de medecine du travail 54, 649-65 I. George, A. O. (1989). Contact leucoderma from paraquat dichloride? Contact Dermatitis 20, 225. Gregorczyk, L., and Swieboda, K. (1968). Vber den EinfluB von Schwefelverbindungen auf die Haut und auf die Schleimhaute [On the effect of sulfur compounds on the skin and mucuous membranes]. Polski Tygognik Lekarski 23, 463-466. Groundwater, J. R (1977). Difolatan dermatitis in a welder; non-agricultural exposure. Contact Dermatitis 3, 104. Hanifin,1. M. (1971). Ethylene oxide dermatitis. J. Am. Med. Assoc. 217,213. Hargreave, P. (1983). Benomyl residues on lichtis after post-harvest dipping. Aust. J. Exp. Agric. Anim. Husb. 23,95-98. Hatao, M., Hariya, T, Katsumura, Y., and Kato, S. (1995). A modification of the local lymph node assay for contact allergenicity screening: Measurement of interIeukin-2 as an alternative to radioisotope-dependent proliferation assay. Toxicology 98, 15-22. Hayes, W. (1982). "Pesticides Studied in Man," p. 408. WiIliams and Wilkins, Baltimore. Higo, A, Ohtake, N., Saruwatari, K., and Kanzaki, T (1996). Photoallergic contact dermatitis from mancozeb, an agricultural fungicide. Contact Dermatitis 35, 183. Hindson, C., and Diffey, B. (I 984a). Phototoxicity of glyphosate in a weedkiIIer. Contact Dermatitis 10, 51-52. Hindson, T. c., and Diffey, B. L. (1984b). Phototoxicity of a weedkiller: A correction. Contact Dermatitis 11, 260. Hirano, S., and Yoshikawa, K. (1982). Patch testing with European and American standard allergens in Japanese patients. Contact Dermatitis 8, 48-50. Horiuchi, N., and Ando, S. (1980). Contact dermatitis due to pesticides for agricultural use. Nippon Hifuka Gakkai Zasshi 90. Hostynek, J. J., Patrick, E., Younger, B., and Maibach, H. L (1989). Hypochlorite sensitivity in man. Contact Dermatitis 20, 32-37. Howard, J. K. (1979). A clinical survey of paraquat formulation workers. Br. 1. Ind. Med. 36, 220-223. Ikarashi, Y., Ohno, K., Momma, J., Tsuchiya, T, and Nakamura, A. (1994). Assessment of contact sensitivity of four thiourea rubber accelerators: Comparison of two mouse lymph node assays with the guinea pig maximization test. Food Chem. Toxico!. 32, 1067-1072. Ikarashi, Y., Tsuchiya, T, and Nakamura, A (1996). Application of sensitive mouse lymph node assay for detection of contact sensitization capacity of dyes. J. Appl. Toxico!. 16,349-354.
332
CHAPTER 11
Regulatory Evaluation of the Skin Effects of Pesticides
Jordan, W. P., Jr., and King, S. E. (1977). Delayed hypersensitivity in females. The development of allergic contact dermatitis in females during the comparison of two predictive patch tests. Contact Dermatitis 3, 19-26. Jung, H. D. (1975). Professional dermatoses in agriculture of the agricultural and industrial district of Neubrandenburg. Dtsch. Gesundheitwesch 30, 1540--1543. Jung, H. D., and Wolff, E (1970a). Nematin (Vapam) contact dermatitis in agricultural workers. Dtsch. Gesundheitwesch 25, 495-498. Jung, H. J., and Wolff, E J. (1970b). Occupational contact dermatitis from nematin (Vapam) in farming. Dtsch. Gesundheitwesch 25, 495-498. Jung, V. H., and Wolf, E (1975). Kontaktezekzeme durch das Herbizid Selest 100 in der Forstwirstschaft. Dtsch. Gesundheitwesch 30, 1540--1543. K Maddy, C. K., and Rivera, L. (1979). "A Study of the Decay Rate of Omite (Propargite) as a Foliar Spray on Strawberries in Ventura County, California, April 1977." California Department of Food and Agriculture, Sacramento. K Maddy, L. R., Kahn, C., and Rivera, L. (1977). "A Study of the Degradation of Propargite (Omite) on Strawberry Foliage and Fruit in Ventura County, California, April 1977." California Department of Food and Agriculture, Sacramento. Koch, P. (1996). Occupational allergic contact dermatitis and airborne contact dermatitis from 5 fungicides in a vineyard worker. Cross-reactions between fungicides of the dithiocarbamate group? Contact Dermatitis 34, 324-329. Koo, D., Goldman, L., and Baron, R. (1995). Irritant dermatitis among workers cleaning up a pesticide spill: California 1991. Am. 1. Ind. Med. 27,545-553. Kranke, B., Szolar-Platzer, c., and Aberer, W. (1996). Reactions to formaldehyde and formaldehyde releasers in a standard series. Contact Dermatitis 35,192-193. Lee, S., Chin, Y., Chang, W, and Kim, J. (1981). A study on hypersensitivity of Korean farmers to various agrochemicals. Determination of concentration for patch test of fruit-tree agrochemicals and hypersensitivity of orange orchard farmers in Che-ju Do, Korea. Seoul Medical lournal22, 137-142. Li, W M. (1986). The role of pesticides in skin disease. Int. J. Dermatol. 25, 295-297. Lisi, P. (1992). Sensitization risk of pyrethroid insecticides. Contact Dermatitis 26,349-350. Maddy, K. et al. (1977). Maddy, K. et al. (1979). Maibach, H. 1. (1986). Irritation, sensitization, photoirritation and photosensitization assays with a glyphosate herbicide. Contact Dermatitis 15, 152-156. Maibach, H. 1., and Johnson, H. L. (1975). Contact urticaria syndrome. Contact urticaria to diethyltoluamide (immediate-type hypersensitivity). Arch. Dermatol. 111, 726-730. Mancuso, G., Reggiani, M., and Berdondini, R. M. (1996). Occupational dermatitis in shoemakers. Contact Dermatitis 34, 17-22. Mathias, C. G. (1997). Allergic contact dermatitis from a lawn care fungicide containing dyrene. Am. 1. Contact Dermat. 8,47-48. Mathias, C. G., Andersen, K. E., and Hamann, K. (1983). Allergic contact dermatitis from 2-n-octyl-4-isothiazolin-3-one, a paint mildewcide. Contact Dermatitis 9, 507-509. Matsushita, T., and Aoyama, K. (1981). Cross reactions between some pesticides and the fungicide benomyl in contact allergy. Ind. Health 19, 77-83. Matsushita, T., Arimatsu, Y., and Nomura, S. (1976). Experimental study on contact dermatitis caused by dithiocarbamates maneb, mancozeb, zineb, and their related compounds. Int. Arch. Occup. Environ. Health 37, 169178. Matsushita, T., Nomura, S., and Wakatsuki, T. (1980). Epidemiology of contact dermatitis from pesticides in Japan. Contact Dermatitis 6, 255-259. Matsushita, T., Yoshioka, M., Aoyama, K., Arimatsu, Y., and Nomura, S. (1977). Experimental study on contact dermatitis caused by fungicides benomyl and thiophanate-methyl. Industrial Health 18, 141-147. Meister, R. (1995). "Farm Chemicals Handbook." Meister Publishing Company, Willoughby, OH. Menne, T. (1991). Relationship between use test and threshold patch test concentration in patients sensitive to 5-chloro-2-methyl-4-isothiazolin-3-one and 2- methyl-4-isothiazolin-3-one (MCIIMI). Contact Dermatitis 24, 375.
Menne, T., Frosch, P. J., Veien, N. K., Hannuksela, M., Bjorkner, B., Lachapelle, J. M., White, 1. R., Vejlsgaard, G., Schubert, H. J., Andersen, K. E., et al. (1991). Contact sensitization to 5-chloro-2-methyl-4isothiazolin-3-one and 2- methyl-4-isothiazolin-3-one (MCIIMI). A European multicentre study. Contact Dermatitis 24, 334-341. Mick, D. L., Gartin, T. D., and Long, K. R. (1970). A case report: Occupational exposure to the insecticide naled. J. Iowa Med. Soc. 60, 395-396. Milby, T., and Epstein, W (1964). Allergic contact sensitivity to malathion. Archives of Environmental Health 9, 434-437. M Reeve, L. O. C., Fong, B., and Edmiston, S. (1991). "Dissipation of Dislodgeable Foliar Residues of Propargite on Grape Foliage, 1989." California Department of Pesticide Regulation, Sacramento. Nater, J. P., and Gooskens, V. H. J. (1976). Occupational dermatosis due to a soil fumigant. Contact Dermatitis 2, 227-229. Nater, J. P., Terpstra, H., and Bleumink, E. (1979). Allergic contact sensitization to the fungicide maneb. Contact Dermatitis 5, 24-26. O'Malley, M. (2000). Pesticides included in the California handler data base. In "Pesticide Dermatoses" (H. Penagos, M. O'Malley, and H. 1. Maibach, eds.). CRC Press, Boca Raton, Florida. O'Malley, M. (1998). "Irritant Chemical Dermatitis Among Grape Workers in Fresno County, August 1995." California Department of Pesticide Regulation, Sacramento. O'Malley, M. A. (1997). Skin reactions to pesticides. Occup Med 12, 327-345. O'Malley, M., and Rodriguez, P. (1998a). "Contact Dermatitis in California Nursery Workers: Part I Surveillance of Skin Disease." California Department of Pesticide Regulation, Worker Health and Safety Branch, Sacramento. O'Malley, M., and Rodriguez, P. (1998b). "Contact Dermatitis in California Nursery Workers: Part 11 Pilot Field Study." California Department of Pesticide Regulation, Worker Health and Safety Branch, Sacramento. O'Malley, M., Rodriguez, P., and Maibach, H. I. (1995). Pesticide patch testing: California nursery workers and controls. Contact Dermatitis 32, 61-63. O'Malley, M., Smith, c., Krieger, R., and Margetich, S. (1989). "Dermatitis Among Stone Fruit Harvesters in Tulare County, 1988." California Department of Food and Agriculture, Sacramento. O'Malley, M., Smith, C., Krieger, R., and Margetich, S. (1990). Dermatitis among stone fruit harvesters in Tulare County, 1988. Am. J. Contact Dermatitis 1, 100--111. O'Malley, M. A., Mathias, C. G., Priddy, M., Molina, D., Grote, A. A., and Halperin, W. E. (1988). Occupational vitiligo due to unsuspected presence of phenolic antioxidant byproducts in commercial bulk rubber. 1. Occup. Med. 30,512-516. Peachey, R. D. G. (1981). Skin hazards in farming. Br. 1. Dermatol. 105,45-50. Peluso, A. M., Tardio, M., Adamo, E, and Venturo, N. (1991). Multiple sensitization due to bis-dithiocarbamate and thiophthalimide pesticides. Contact Dermatitis 25, 327. Penagos, H. (1993). Articulo de revision: la prueba del parche. Usos en dermatitis de contacto y en la dermatologia ocupacional. [Review article: the patch test-uses in contact dermatitis and occupational dermatology]. Dermatitis de Contacto-Boletin Informativo del GCIDC 2, 2-11. Pilger, c., Nethercott, J. R., and Weksberg, E (1986). Allergic contact dermatitis due to a biocide containing 5-chloro-2- methyl-4-isothiazolin-3-one. Contact Dermatitis 14,201-204. Rademaker, M. (1998). Occupational contact dermatitis among New Zealand farmers. Australas. 1. Dermatol. 39, 164-167. Romaguera, c., and Grimalt, E (1980). Irritant dermatitis from ethylene oxide. Contact Dermatitis 6,351. Romaguera, c., Grimalt, E, Torras, H., Castel, T., and Mascaro, J. M. (1977). Contact dermatitis caused by ethylene oxide. Med. Cutan. Ibero Lat. Am. 5, 309-312. Romaguera, C., and Vilaplana, J. (1998). Airborne occupational contact dermatitis from ethylene oxide. Contact Dermatitis 39, 85. Rudner, E. (1977). North American Group results. Contact Dermatitis 3, 208209. Rycroft, R. J. (1977). Contact dermatitis from organophosphorus pesticides. Br. 1. Dermatol. 97, 693-695.
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CHAPTER
12 Neurophysiological Effects of Insecticides Toshio N arahashi Northwestern University Medical School
The latter half of the 20th century has witnessed a considerable advance in our knowledge concerning the mechanisms of action of insecticides. This was due mostly to impressive developments of newer, mostly synthetic, insecticides, and rapid progress in various technologies in the field of biomedical sciences. Among various areas of the insecticide mechanism of action, studies of their metabolism were among the earliest developments, starting in the 1950s. However, it was not until the 1960s that studies of the cellular or physiological mechanism of action of insecticides became widespread. More recently, applications of molecular biology and genetics techniques have made it possible to identify the molecular species that are responsible for the toxic action of insecticides, particularly those related to the target resistance of insects to insecticides. Most insecticides are neuropoisons, but their target sites are rather limited. For example, voltage-gated sodium channels are the major target of pyrethroids and DDT; GABAA receptors are attacked by cyclodienes, hexachlorocyclohexane (HCH), and fipronil; neuronal nicotinic acetylcholine (nnACh) receptors are the target of nicotine, and nitromethylene and nitroimine heterocycles (e.g., imidacloprid). Organophosphate and carbamate insecticides inhibit acetylcholinesterase. This chapter covers the neurophysiological mechanisms of action of various insecticides. However, since a large number of review articles have already been published, emphasis will be placed on recent developments in the field. Readers are encouraged to refer to these review articles, each of which discusses similar issues from somewhat different points of view. These articles, though not limited to, are as follows: Narahashi (1971,1976, 1985, 1988, 1989, 1992, 1996), Narahashi et al. (1995, 1998), Ruigt (1984), Soderlund and Bloomquist (1989), Vijverberg and van den Bercken (1990), Salgado (1999), Clark (1997), Bloomquist (1996), and Casida and Quistad (1998). Handbook of Pesticide Toxicology
Volume 1, Principles
12.1 PYRETHROIDS AND DDT 12.1.1 SODIUM CHANNEL MODULATION Despite apparent differences in chemical structure, pyrethroids and DDT exert similar actions on the nervous system through modulation of the function of voltage-gated sodium channels. Pyrethroids may be divided into two groups: type I pyrethroids lack a cyano group in the a position, and their symptoms of poisoning are characterized by hyperexcitation, ataxia, convulsions, and paralysis; type 11 pyrethroids have an a cyano group, and cause hypersensitivity, choreoathetosis, tremors, and paralysis. At the level of nerve function, type I pyrethroids tend to produce repetitive action potentials as a result of the increase in depolarizing after-potential, whereas type 11 pyrethroids tend to cause membrane depolarization leading to discharges from sensory neurons. These apparent differences in nerve function alteration between the two types of pyrethroids can be ascribed to differences in modification of sodium channel kinetics. DDT has many features in common with type I pyrethroids with respect to the mechanism of action on the sodium channel. Changes in Sodium Channel Gating Kinetics Depolarizing after-potential is gradually increased after application of type I pyrethroids such as tetramethrin and allethrin, and reaches the threshold membrane potential for generation of action potentials (Lund and Narahashi, 1981a, b; Narahashi, 1962; Vijverberg et aI., 1982). The mechanism by which the depolarizing after-potential is increased can best be studied by the voltage clamp technique (Fig. 12.1). The tail current upon termination of a depolarizing pulse was greatly increased and prolonged in the presence of pyrethroid. Type 11 pyrethroids such as deltamethrin and fenvalerate caused much greater prolongation of sodium currents during and upon termination of a depolarizing pulse than type I pyrethroids (Brown and Narahashi, 1987, 1992; Ogata et aI., 1988; Salgado et aI., 1989; Song et aI., 1996; Tabarean and Narahashi, 1998). Cockroach neurons cultured from the brain of 21-day-old embryos did not express
335
Copyright © 200 1 by Academic Press. All rights of reproduction in any form reserved.
336
CHAPTER 12
Neurophysiological Effects of Insecticides (a)
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sodium channel activIty, yet deltamethrin unveiled "silent" sodium channels which were partly blocked by tetrodotoxin (TTX) (Amar and Pichon, 1992). Pyrethroid modulation of individual sodium channels was studied by single-channel patch clamp techniques using neuroblastoma cells. While normal sodium channels opened for a few milliseconds at the beginning of a depolarizing pulse, channels exposed to pyrethroid opened for a very long period of time often extending a few seconds and with a long delay from the beginning of the depolarizing pulse (Fig. 12.2) (Chinn and Narahashi, 1986; Holloway et aI., 1989; Yamamoto et aI., 1983). In the presence of pyrethroid, sodium channels often remained open after termination of the depolarizing pulse reflecting the whole-cell tail current. These observations have led to the conclusion that the kinetics of both activation and inactivation gates are slowed and the gates tend to be stuck at the open or closed position (Chinn and Narahashi, 1986; Vijverberg et aI., 1982). As expected from these results, the gating currents associated with both opening and closing of the sodium channel were inhibited by pyrethroid (Salgado and Narahashi, 1993). Extremely prolonged sodium channel openings (up to several seconds) were also observed in cockroach neurons in culture in the presence of deltamethrin (Amar and Pichon, 1992).
State Dependency of Sodium Channel Modification A drug could bind to a channel at its closed state, its open state, or both states. This is an important aspect of drug-channel interaction. Extensive studies along this line have led to the conclusion that pyrethroids modify the sodium channel function at its closed state but the open channel has a higher affinity for pyrethroids. Thus, pyrethroids act on the closed sodium channel, and opening further recruits modified channels (Brown and Narahashi, 1992; de Weille et aI., 1988; Ginsburg and Narahashi, 1999; Holloway et aI., 1989). It should be noted that the ratio of open sodium channel modification to closed sodium channel modification varies considerably in different preparations. For example, a large fraction of pyrethroid modification occurred in the closed channel
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8 ~W"A+i#..\Ip,~vi.-!f.'w./i.~..,.~~~.,,~W:r~\ ~.:, ·~N·l:rt'.1 .
~-100mV
-30mV
140 ms
.
3s
L
(b) Figure 12.2 Deltamethrin prolongation of single sodium channel currents recorded from a neuroblastoma cell (N1E-llS). (a) Currents from a cell before drug treatment in response to 140-msec depolarizing steps from a holding potential of -100 mV to - 30 mV with a 3-sec interpu1se interval. Records were taken at a rate of 100 j.lsec per point. (b) Currents after exposure to 10 j.lM deltamethrin. The membrane patch was depolarized for 3140 msec from a holding potential of -100 mV to - 30 mV. The interpulse interval was 3 sec. The time scale changed during the voltage step as indicated in the figure. During the first 140 msec, records were taken at a rate of 100 j.lsec per point, and after the vertical line, records were taken at a rate of 10 msec per point. From Chinn and Narahashi (1986).
12.1 Pyrethroids and DDT
state in squid giant axons (de Weille et al., 1988), in rat dorsal root ganglion neurons (Ginsburg and Narahashi, 1999), and in mouse neuroblastoma cells (Holloway et aI., 1989), whereas pyrethroid modification occurred largely in the open channel state in frog muscle fibers (Leibowitz et aI., 1986). Open Channel Properties While passing through an open sodium channel, the permeating cation must cross barriers, and temporarily binds to sites inside the channel. Thus, open sodium channels are not only permeable to but also blocked by various monovalent and divalent cations to a varying extent. The permeability ratios in squid axons for Na:Li:ammonium: guanidine:formamidine were 1: 1.19:0.21 :0.28:0.20 for the normal sodium channel, and 1:1.18:0.29:0.29:0.25 for the channel modified by tetramethrin (Yamamoto et aI., 1986). It is concluded that pyrethroid does not alter the permeability properties of open sodium channel. Site of Action of Pyrethroids in the Sodium Channel A variety of experimental approaches have been taken to determine the site of action of pyrethroids in the sodium channel. Pyrethroids have been shown to bind to a site different from any other known sites for various toxins and chemicals. n-Octylguanidine blocked the sodium channel by entering from inside the membrane when the gates are open (Kirsch et aI., 1980) in a manner similar to that of local anesthetics (Courtney, 1975; Hille, 1977; Strichartz, 1973; Yeh, 1978, 1980), pancuronium (Yeh and Narahashi, 1977), 9-aminoacridine (Yeh, 1979), and strychnine (Shapiro, 1977). The octylguanidine binding site was not the site for pyrethroids as they did not interact with each other (de Weille et aI., 1988). Batrachotoxin (BTX) and grayanotoxin (GTX) slow the kinetics of the activation and inactivation gates of the sodium channel and shift their voltage dependence in the hyperpolarizing direction, resulting in slow and prolonged sodium current and membrane depolarization (Albuquerque et aI., 1971; Khodorov, 1985; Khodorov et aI., 1976; Narahashi et aI., 1971; Narahashi and Seyama, 1974; Seyama and Narahashi, 1973, 1981; Tanguy and Yeh, 1991). BTX and GTX are also known to bind to site 2 of the sodium channel (Catterall, 1992). Tetramethrin action was not modified by either BTX (Tanguy and Narahashi, unpublished) or GTX (Takeda and Narahashi, 1988). Therefore, pyrethroids bind to a site other than site 2. TTX selectively blocks the sodium channel (Narahashi et aI., 1964) through binding to site 1 (Catterall, 1992). TTX blocked the tetramethrin-modified sodium channel in a noncompetitive manner, indicating that tetramethrin did not bind to the TTX site (site 1) (Lund and Narahashi, 1982). A recent study has shown that pyrethroids modify the a subunit of the sodium channel expressed in Chinese hamster ovary cells through binding to a site other than any other known binding sites for various toxins and chemicals (Trainer et aI., 1997). Binding of eH]batrachotoxinin A-20-a-benzoate (eHJBTX-B) to mouse brain sodium channels was modified by pyrethroids and DDT (Rubin et aI., 1993). Although deltamethrin and the 2S stereoisomers of fenvalerate enhanced
337
eHJBTX-B binding, nontoxic isomers inhibited the binding or caused no effect. DDT and its analogs and metabolites enhanced the binding. However, toxic type I pyrethroids enhanced, inhibited, or had no effect on the binding, and the effects were not correlated with toxicity. These data illustrate a limitation in the use of this assay as a screen for neurotoxicity (Rubin et aI., 1993). The role of a and 131 subunits of rat brain Ha sodium channel in pyrethroid action was studied using Xenopus oocyte expression and voltage clamp techniques (Smith and Soderlund, 1998). In both the a and the a plus 131 subunits expressed in oocytes, cypermethrin caused prolonged tail sodium currents. However, the cypermethrin affinity was 20 times higher in the a plus 131 combination than in the a subunit alone. Differential Pyrethroid Sensitivity to TTX-Sensitive and TTX-Resistant Sodium Channels Most sodium channels in the nervous system are highly sensitive to TTX block, with an ICso in the range of nanomolar concentrations. By contrast, cardiac sodium channels are less sensitive to TTX, with an ICso on the order of micromolar concentrations. During the past several years, TTX-resistant (TTX-R) sodium channels in the nerve have received much attention, partly because some of these channels in mammalian dorsal root ganglia (DRG) are related to pain sensation, opening the door for the possible development of drugs that selectively block TTX-R sodium channels as useful analgesics. Although the initial discovery ofTTX-R sodium channels in DRG was made almost 20 years ago by Kostyuk et al. (1981), it was not until the early 1990s that their significance received much attention after being revisited by Roy and Narahashi (1992). The ICso for TTX-R sodium channels was about 100 ).LM, a value 100,000 times higher than that for TTXsensitive (TTX-S) sodium channels. Analyses ofTTX-R as well as TTX-S sodium channels have been performed extensively not only for their physiology and biophysics (Elliott and Elliott, 1993; Ogata and Tatebayashi, 1993), but also for their molecular structures (Akopian et aI., 1996; Sangameswaran et aI., 1997). Significance of TTX-R sodium channels in insecticide toxicology has been demonstrated for pyrethroids. TTX-R sodium channels of rat DRG neurons were more sensitive to pyrethroid modulation than TTX-S sodium channels of DRG neurons (Ginsburg and Narahashi, 1993; Tatebayashi and Narahashi, 1994). An example of such a patch clamp experiment is shown in Fig. 12.3. Although TTX-S sodium channel current during a depolarizing pulse was only slightly affected by 1 ).LM tetramethrin (Fig. 12.3a), TTX-R sodium channel current underwent drastic changes including the appearance of a large tail current upon termination of the depolarizing step (Fig. 12.3b). Similar differential sensitivity to pyrethroids was also found between insect and mammalian sodium channels. Currents were recorded from Xenopus oocytes expressing para sodium channel a subunit from Drosophila and rat brain type HA sodium channels (Warmke et aI., 1997). Permethrin was over 100 times more potent in modulating sodium currents of para sodium channels than those of brain HA sodium channels. The
338
CHAPTER 12
Neurophysiological Effects ofInsecticides
o
-110:r-l~---------------CONTROL
-.-J
TETRAMETHRIN (1 ~M)
4nA
10 msec
(a)
-90~~-------------
TETRAMETHRIN
(1 ~M)
~4nA 10 msec
(b) Figure 12.3 Effects of tetramethrin on tetrodotoxin (TTX)-sensitive sodium current (a) and TTX-resistant sodium current (b) in rat dorsal root ganglion neurons. A step depolarization to 0 mV was applied from a holding potential of -110 mV (a) or -90 mV (b) in control and in the presence of I f.lM tetramethrin. From Tatebayashi and Narabashi (1994).
differential sodium channel sensitivity is one of the crucial factors that account for the selective toxicity of pyrethroids, as will be discussed later.
Amplification of Pyrethroid Toxicity from Sodium Channels to Animals An early study by Lund and Narahashi (1982) using squid giant axons suggested that only a very small fraction of the sodium channel population needed to be modified by pyrethroids to cause repetitive discharges. This was based on the calculation of the percentage of sodium channels needed to increase the depolarizing after-potential to the level of threshold membrane potential for generation of repetitive action potentials. However, a few assumptions had to be made for calculation, as not all data were available at that time. Later, Tatebayashi and Narahashi (1994) developed a method to calculate the percentage of sodium channel modification caused by pyrethroid based on patch clamp data using rat DRG neurons. Since the peak sodium current (fNa) during a depolarizing pulse was not affected by pyrethroid, it represented the activity of normal or unmodified sodium channels. The tail current (ftail) upon termination of a depolarizing pulse appeared only after application of pyrethroid, and therefore it represented the activity of modified sodium channels. The percentage of modification (M) can be calculated by the following equation: M
= [{Itail/(Eh - ENa)}/{INa/(Et
-
ENa)}] x 100
(1)
where [tail is the tail current amplitude obtained by extrapolation of the slowly decaying phase of the tail current to the moment of membrane repolarization assuming a single exponential decay, Eh is the potential to which the membrane was repolarized, E Na is the equilibrium potential for sodium ions obtained as the reversal potential for sodium current, and Et is the potential of step depolarization. The percentages of sodium channels modified by tetramethrin were very small: for example, for TTX-S sodium channels, 0.24%,3.53%, and 12.03% by 0.1,1, and 10 J.!M tetramethrin, respectively; forTTX-R sodium channels, 1.31%, 15.35%,57.82%, and 81.20% by 0.Q1, 0.1, 1, and 10 J.!M tetramethrin, respectively. Thus, TTX-R sodium channels are approximately 30 times more sensitive to tetramethrin than TTX-S sodium channels. A question arises as to the degree of pyrethroid modification needed to cause repetitive nerve activity. Using the same method of calculation and also comparing these calculated data with the threshold concentration for tetramethrin needed to induce repetitive discharges in rat cerebellar Purkinje neurons, an astonishingly small percentage was obtained, that is, 0.62%, as illustrated in Fig. 12.4 (Song and Narahashi, 1996). This provides one of the bases for high potency of pyrethroid action. It is also important to note that the significance of this "toxicity amplification" is not limited to pyrethroids. When a drug slightly suppresses the slow depolarization (e.g., caused by activation of T-type calcium channels or in epileptic seizure), repetitive discharges generated by the slow depolarization will stop, and for this action only a concentration of the drug (e.g., antiepileptic drug) much lower than the ICso for suppressing the depolarization (or calcium channels) will be needed, perhaps ICIO or even IC]. Thus, "pharmacological amplification" will become important for interpreting the drug action in vivo. The traditional concept of relating in vitro ICso to a patient's serum concentration of the drug may not necessarily be valid when the effect is exerted via the threshold phenomenon.
Temperature Dependence of Pyrethroid Action It is well known that the insecticidal activity of pyrethroids and DDT increases with decrease in temperature. This is important, as the negative temperature dependence is partially responsible for selective toxicity in insects and mammals. This phenomenon is also deemed to contain some keys to the molecular mechanism of action of these insecticides. The earliest study was performed by Yamasaki and Ishii [Narahashi] (1954a) for the action of DDT on repetitive discharges of cockroach nerve. It was clearly demonstrated that the major factor for the negative temperature dependence of insecticidal action was the nerve sensitivity to DDT, which showed the QIO value of 0.2. The effect of DDT in inducing repetitive discharges was reversible with respect to temperature change, and therefore, the metabolism of DDT did not come into play. Several studies have since been performed for the negative temperature dependence of DDT and pyrethroid actions on nerves (Ahn et aI., 1987; Gammon, 1978; Narahashi, 1962; Salgado et aI., 1989; Starkus and Narahashi, 1978). Binding of DDT to housefly brain increased with decreases in the temperature
12.1 Pyrethroids and DDT
-110
339
~----- TTX 0.5 f,LM CONTROL
35 CC
_12
IV--/-- TM 3 f,LM
nA
-110~---------------
10 msec TM 10 f,LM
~1
(a) 30
Q W
nA
10 msec
G: Ci
25
o
::::E
I
Cl)
~-I
20
(a)
c:::w
~ ~
w<
30 CC
15
::::EI
~
U
10
ti llL.
o
CONTROL
5
~
7
6
5
4
TETRAMETHRIN 3 f,LM
3
TETRAMETHRIN (-log molar concentration)
(b) 80
s:-
S
40 0
E w -40
50 msec
-80
(c) Figure 12.4 Concentration-dependent effect of tetramethrin on TTX-S sodium currents of rat cerebellar Purkinje neurons. (a) Currents were evoked by a 5-msec step depolarization to 0 mV from a holding potential of -110 mV under control conditions and in the presence of tetramethrin (0.3 J.!M, 3 J.!M, and 10 ~M). TTX (0.5 ~M) completely blocked both the peak current and the tetramethrin-induced tail current. (b) The concentration-response relationship for induction of tail current. Each point indicates the mean ± S.E.M. (n = 6). Data were fitted by the Hill equation. The percentages of channels modified by tetramethrin are 0.62 ± 0.15%,2.19 ± 0.36%,5.75 ± 0.87%, 13.58 ± 1.35%, 22.77 ± 2.26%, and 24.73 ± 2.11% at concentrations of 0.1, 0.3, 1,3,10, and 30 J.!M, respectively (n = 6). (c) Repetitive after-discharges caused by 100 nM tetramethrin, the threshold concentration. Action potentials were evoked by applying a current pulse (2 msec, 200 pA). Em refers to the membrane potential. From Song and Narahashi (1996).
(c)
20°C
(d)
(Chang and Plapp, 1983). The sodium tail current slowed by pyrethroids was further slowed by lowering the temperature (Vijverberg et aI., 1983). Despite these studies over many years, it was not until the mid-1990s that the physiological mechanism that underlies the negative temperature dependence of pyrethroid action on
Figure 12.5 Temperature-dependent effect of 3 J.!M tetramethrin on sodium currents recorded from a rat cerebellar Purkinje cell. The currents were evoked by a 5-msec step depolarization to 0 mV from a holding potential of -110 mV at various temperatures. The currents before and during application of tetramethrin are superimposed at each temperature. *, Current recording is truncated before the tail current returns to the base-line. From Song and Narahashi (1996).
340
CHAPTER 12
Neurophysiological Effects of Insecticides
the nerve was clearly elucidated. Song and Narahashi (1996) have perfonned current clamp and voltage clamp experiments using rat cerebellar Purkinje neurons. Repetitive discharges induced by tetramethrin at 15-20 o e subsided with an increase in the temperature to 30-35°C. The tail sodium channel current in the presence oftetramethrin was drastically affected by temperature changes (Fig. 12.5). Although the peak amplitude of the tail current was not changed by lowering the temperature from 30°C to 20°C, the decay phase of the tail current was greatly slowed, showing a QIO value of 0.07, and the charge movement during tail current was increased, with a QIO value of 0.2. Small QIO values (large negative temperature dependence) for pyrethroid-induced tail current decay were also observed with frog nodes of Ranvier (Vijverberg et aI., 1983). The percentage of sodium channels modified by tetramethrin was only slightly increased by lowering the temperature from 30°C to 20°C, with a QIO value of 0.77. Thus, the most critical factor for the negative temperature dependence of repetitive discharges is slowing of the tail current decay, which causes a sizable increase in tail charge transfer by lowering the temperature.
Selective Toxicity of Pyrethroids Pyrethroids are more toxic to insects than to mammals, with differences in LD50 ranging from 500- to 4500-fold (Elliott, 1977; Hirai, 1987; Miyamoto, 1993; Wiswesser, 1976). The selective toxicity of various insecticides has been generally ascribed to higher rates of enzymatic degradation of insecticides in mammals than in insects. It was assumed that this was also the case for pyrethroids, albeit without any solid data to justify the assumption. However, this is not the case as far as pyrethroids are concerned. Various factors pertaining to selective toxicity of pyrethroids are given in Table 12.1. Pyrethroids are more potent on nerve function at low temperature than at high temperature, and the QIO value is calculated to be 0.2, indicating that the potency increases fivefold with a decrease in temperature of 10°C, the body temperature difference between insects and mammals (Song and Narahashi, 1996). The intrinsic sensitivity of nerve (i.e., sodium channels) is at least 10 times, sometimes 100 times, higher in invertebrates than in mammals (Song and Narahashi, 1996; Warrnke et aI., 1997). Recovery after washout is approximately five times slower in invertebrates than in mammals (Song and Narahashi, 1996). Detoxication of pyrethroids
is known to involve various enzymes whose rates are approximately three times lower in insects due to lower body temperature. Smaller body size in insects makes detoxication less efficient before pyrethroids reach the target site. A difference of 2250-fold is obtained by multiplying differences in these factors, and this value is the same order of magnitude as the difference in LD50S described above. Therefore, the major factors responsible for the large difference in LD50 values between insects and mammals are all related to sodium channels.
Vitamin E Alleviation of Pyrethroid-Induced Paresthesia Any chemicals that block pyrethroid-modified sodium channels without effect on normal sodium channels could serve as antidotes for pyrethroid intoxication. Local anesthetics such as lidocaine were once considered (Oortgiesen et aI., 1990), but they also block nonnal sodium channels. One such possibility is vitamin E, which has been used for prophylactic and therapeutic purposes to alleviate paresthesia caused by pyrethroids. The paresthesia includes tingling, itching, and burning sensation of the skin, without the clinical symptoms of erythema, edema, or vesiculation (Knox et aI., 1984; LeQuesne et aI., 1980). Pyrethroids, particularly type 11 pyrethroids, cause such paresthesia in the facial skin, and vitamin E has been used for therapeutic purposes (Flannigan and Tucker, 1985; Tucker et aI., 1984). Vitamin E was found to be effective in blocking tetramethrin-modified sodium channels without effect on nonnal sodium channels in rat cerebellar Purkinje neurons and DRG neurons (Song and Narahashi, 1995). Vitamin E shortened the action potential duration prolonged by tetramethrin without affecting the peak amplitude. Reflecting this effect on action potential, the tail sodium channel current was blocked by vitamin E in a competitive manner while the peak sodium current remained unchanged (Fig. 12.6). However, the mechanism of this interesting antagonism is open to question, and it is not known whether the antioxidant action of vitamin E has anything to do with the antagonism. Recently, a-tocopherol was shown to antagonize the type I pyrethroid action in vivo in susceptible and kdr-resistant insects increases the LDso values by 4.3 to 6.6-fold (Scott, 1998). Thus, a-tocopherol opens the door for development of antidotes for pyrethroid intoxication (Song and Narahashi, 1995).
Table 12.1 Factors Contributing Selective Toxicity of Pyrethroids Mammals
Selectivity factor
Insects
Differences
Potency on nerve
(37°C)
(25°C)
5
Due to temperature dependence
Low
Due to intrinsic sensitivity
Fast
High
10
Recovery
Fast
Slow
5 3 3
High
Detoxication rate Due to enzymatic action
High
Low
Due to body size
High
Low
Overall difference =
2250. (From Song and Narahashi,
1996.)
12.1 Pyrethroids and DDT CEREBELLAR PURKINJE TTX-S 25 w20
z z <
CONTROL
o
TM 10 I'M
rs::sJ
en
....l
TM 10 I'M +a-TOCO 10 I'M ~ TM 10 I'M +a-TOCO 30 I'M
G15
J'
o o ::;:
u.. 5
nA
10 meec
o
~
(a) CONCENTRATION-RESPONSE RELATIONSHIP 30
en
....l
w 25
z z < :J: 20
0 - a- Toeo (n=6)
•+
a- TOCO (10 p.M. n=6)
U
0
w 15
ii:
CS
0 ::;: 10
u.. 0
~
5 0
-7 -6 -5 TETRAMETHRIN (LOG MOLAR CONCENTRATION)
same neuron remains totally unchanged (Ogata et aI., 1988). Thus, even though type 11 pyrethroids inhibit GABAA receptors to some extent under certain experimental conditions, the toxicological significance is rather questionable.
Pyrethroid Modulation of Calcium Channels Permethrin at a concentration as low as 50 pM increased the electrical activity of neurosecretory cells of the stick insect (Orchard and Osbome, 1979), and the effect was ascribed to the action on calcium channels (Gammon and Sander, 1985; Osbome, 1980). However, our patch clamp experiments using neuroblastoma cells showed a blocking, not a stimulating, action of pyrethroids on both T-type and L-type calcium channels (Yoshii et aI., 1985). It should be noted that the observed impulse discharges from the insect neurosecretory cells may originate in sodium channels of presynaptic neurons.
o w ~ 10
TM 10 I'M
341
-4
(b) Figure 12.6 (a) Suppression of 10 ).lM tetramethrin-induced tail currents by 10 and 30 ).lM (±)-a-tocopherol in TTX-S sodium channels of rat cerebellar Purkinje cells. Currents were evoked by depolarizing the membrane to 0 m V for 5 msec from a holding potential of -110 m V. Cells were first treated with 1O).lM tetramethrin, and then 10 or 30 ).lM (±)-a-tocopherol was added to the perfusion solution containing 10 ).lM tetramethrin. Records were taken 5 min after the addition of each chemical. The percentage of channel modification was calculated by Eq. (1). Mean ± S.E.M. with n = 6. (b) (±)-a-Tocopherol shifts the concentration-response relationship for tetramethrin modification in the direction of higher concentrations in TTX-S sodium channels of cerebellar Purkinje cells. Mean ± S.E.M. (n = 6). From Song and Narahashi (1995).
12.1.2 PYRETHROID ACTION ON OTHER RECEPTORS AND CHANNELS
Pyrethroid Modulation of GABA A Receptors Several papers have been published to report the block of GABAA receptors by type 11 pyrethroids (Abalis et aI., 1986; Bloomquist and Soderlund, 1985; Crofton et aI., 1987; Eshleman and Murray, 1990, 1991; Gammon and Sander, 1985; Lawrence and Casida, 1983; Lawrence et aI., 1985; Lummis et aI., 1987; Ramadan et aI., 1988). Despite these reports, the matter has been controversial, as the potency and efficacy of pyrethroids in blocking the GABAA receptors are low. Our patch clamp experiments using rat DRG neurons have unequivocally shown that while 10 J..I.M deltamethrin markedly prolongs the sodium current as expected, the GABA-induced current recorded from the very
Pyrethroid Modulation of Chloride Channels N1E-115 neuroblastoma cells are endowed with calcium-independent voltage-gated chloride channels. The type 11 pyrethroids deltamethrin and cypermethrin suppressed the channel activity by decreasing open probability, but type I pyrethroid cismethrin had much less effect (Forshaw et aI., 1993; Ray et aI., 1997). These chloride channels exhibited a high conductance of 340 pS. However, since the physiological function of these chloride channels is unknown, toxicological significance for pyrethroid action awaits further experimentation. Pyrethroid Modulation of Acetylcholine Receptors The binding of eH]perhydrohistrionicotoxin to the Torpedo electric organ was inhibited by type I and type 11 pyrethroids (Abbassy et aI., 1982, 1983a, b; Eldefrawi et aI., 1984; Sherby et al., 1986). On the contrary, the frog end-plate potential was not affected by allethrin (Wouters et aI., 1977). This paradox remains to be solved. Pyrethroids have also been shown to interact with muscarinic ACh receptors (Eriksson and Nordberg, 1990; Eriksson and Fredricksson, 1991). Deltamethrin suppressed ACh-induced currents in Helix neurons (Kiss and Osipenko, 1991). A question was raised whether the action of pyrethroids on ACh receptors represented a specific interaction, because both active and inactive isomers of pyrethroids exerted nonspecific, inhibitory effects on the nicotinic ACh receptors of N1E-115 neuroblastoma cells (Oortgiesen et aI., 1989). The significance of ACh receptors, especially that of neuronal nicotinic ACh receptors, has received much attention these days with respect to physiology and pharmacology; thus, more elaborate experimental analyses for pyrethroid interactions with these receptors are warranted. Pyrethroid Modulation of Glutamate Receptors The [3H] kainate binding to mouse brain homogenates was inhibited by pyrethroids: ICsos were 80 nM for deltamethrin and 8 J..I.M for cispermethrin (Staatz et aI., 1982). Cypermethrin at 1 J..I.M suppressed the glutamate sensitivity of the muscle of housefly larvae (Seabrook et aI., 1988). However, the toxicological sig-
342
CHAPTER 12
Neurophysiological Effects of Insecticides
nificance of glutamate receptors for pyrethroid action remains largely to be seen.
Role of Calcineurin and Other Enzymes in Pyrethroid Action Pyrethroids have been shown to inhibit Na-Ca ATP hydrolysis and Ca-Mg ATP hydrolysis (Clark and Matsumura, 1987). Deltamethrin stimulated protein phosphorylation and caused the release of calcium from the intracellular storage sites (Enan and Matsumura, 1991; Matsumura et aI., 1989). Pyrethroids, both type I and type II, stimulated phosphoinositide breakdown (Gusovsky et aI., 1986). A striking discovery was made regarding calcineurin, neural calcium-calmodulin-dependent protein phosphatase, which was inhibited by type II pyrethroids such as cypermethrin, deltamethrin, and fenvalerate with IC50 values of 0.0 1-1 nM (Enan and Matsumura, 1992). By contrast, insecticidally inactive chiral isomers of these pyrethroids, active type I pyrethroids, DDT and heptachlor expoxide were much weaker inhibitors. However, recent studies conducted by two independent groups cast doubt on the pyrethroid inhibition of calcineurin. None of the five pyrethroids tested, that is, bioallethrin, cyfluthrin, cypermethrin, deltamethrin, and fenvalerate, caused inhibition of the calcineurin-dependent dephosphory lation (Enz and Pombo-Villar, 1997). Both type I pyrethroids (cis-permethrin, trans-permethrin, and S-bioallethrin) and type II pyrethroids (cis-cypermethrin, trans-cypermethrin, deltamethrin, and fenvalerate) were unable to inhibit the phosphatase activity of purified calcineurin (Fakata et aI., 1998). Thus, the role of calcineurin in pyrethroid actions remains unclear. 12.1.3 SODIUM CHANNEL MUTATION IN
PYRETHROID RESISTANCE Earlier studies indicated that insecticide-resistant strains of insects acquired higher activity to detoxify insecticides (Wilkinson, 1983). However, a metabolic resistance mechanism could not completely explain the resistance to insecticides, because resistant strains of insects often contained unmetabolized insecticide in an amount much more than enough to kill susceptible strains. Insecticide resistance due to reduced nerve sensitivity was termed knockdown resistance (kdr) (Busvine, 1951; Milani, 1954). The mechanism of target site resistance was first studied for DDT, lindane and dieldrin. The sensitivity of the sensory nerves to DDT was lower in resistant houseflies than in susceptible houseflies (Smyth and Roys, 1955; Weiant, 1955). However, multiple discharges from the central nervous system (CNS) caused by insecticides are more closely related to the development of symptoms of poisoning. By electrophysiological measurements of such CNS multiple discharges as a measure of toxic action, resistant strains of houseflies were found to be less sensitive than susceptible strains for lindane and dieldrin (Yamasaki and Narahashi, 1958b) and for DDT (Yamasaki and Narahashi, 1962). While the identification of chromosome genes for low nerve sensitivity to DDT was made in the mid-1960s (Tsukamoto
et aI., 1965), studies for more precise sodium channel sites of mutations responsible for insecticide resistance were commenced only after thorough developments of molecular biology and genetic techniques in the 1990s. We now know mutations occur at several sites in the a subunit of sodium channels of pyrethroid-resistant kdr and super-kdr strains of various insects (Table 12.2).
12.2 CYCLODIENES AND HEXACHLOROCYCLOHEXANE The mechanisms of action of cyclodienes and hexachlorocyclohexane (HCH) have a long history of studies. In the 1950s, dieldrin and lindane (y-HCH) were shown to stimulate synaptic transmission in the cockroach nerve (Yamasaki and Ishii [Narahashi], 1954b; Yamasaki and Narahashi, 1958a). However, it was not until the 1980s that the GABA receptor was identified as their major target site by 36Cl- uptake and 5S]t-butylbicyclophosphorothionate (TBPS) binding experiments (Abalis et aI., 1986; Bermudez et aI., 1991; Bloomquist and Soderlund, 1985; Bloomquistet aI., 1986; Cole and Casida, 1986; Ghiasuddin and Matsumura, 1982; Llorens et aI., 1990; Lummis et aI., 1990; Matsumoto et aI., 1988; Matsumura and Ghiasuddin, 1983; Pomes et al., 1994; Olsen et aI., 1989; Thompson et aI., 1990). The first electrophysiological experiment to demonstrate that the GABAA receptor was the target site was performed by Ogata et al. (1988), who showed lindane suppression of GABA-induced chloride currents in rat DRG neurons. The effects of lindane and dieldrin on single-channel characteristics of cockroach GABA receptors were studied by noise analysis (Bermudez et aI., 1991). Both insecticides decreased the frequency of channel opening. Dieldrin was without effect on the single-channel conductance, but lindane decreased it. However, Zufall et al. (1989) found no effect of lindane on single channels of crayfish stomach muscle. Whereas lindane inhibited all three types of GABAA receptors of rat cerebral cortex expressed in Xenopus oocytes, a-, 8-, ,B-HCH had differential effects (Woodward et aI., 1992). Endrin, dieldrin, and lindane also suppressed electrophysiological responses of cockroach and locust GABA receptors (Bermudez et al., 1991; Wafford et aI., 1989). Similarity between lindane and picrotoxin in blocking GABA receptors is pointed out (Tokutomi et al., 1994; Zufall et aI., 1989).
e
Dual Action of Dieldrin on GABAA Receptors Dieldrin has been found to exert a dual action on GABAA receptors. During repetitive co-applications of GABA and dieldrin, the GABA-induced current was first increased but later suppressed irreversibly (Fig. 12.7) (Nagata and Narahashi, 1994). The dual action was not only time dependent but also dieldrin concentration dependent. There were two components of suppression with IC50 values of 5 and 92 nM; EC50 for potentiation was 754 nM. Analysis of picrotoxin-dieldrin interaction experiments led to the conclusion that dieldrin acts on the picrotoxin site which is closely associated with the chloride channel.
12.2 Cyclodienes and Hexachlorocyclohexane
343
Table 12.2 Mutation of Sodium Channel Amino Acids in Pyrethroid Resistance Locations of mutation in sodium channels Domain and transmembrane
Amino acid
Species
segment
sequence position
References
Drosophila melanogaster
IS4-S5
I253N
Pittendrigh et at. (1997)
Heliothis virescens
IS6
V410M
Park et at. (1997) Lee et at. (1999b)
Musca domestica
IIS4-S5
M9l8T
Williamson et at. (1996) Lee et at. (1999a)
Plutella xylostella
IIS5
T929I
Schuler et at. (1998)
Musca domestica
IIS6
L993F
Miyazaki et at. (1996)
Blattella germanica
IIS6
L993F
Dong (1997)
Heliothis virescens
IIS6
L993H
Park and Taylor (1997)
Musca domestica
IIS6
LlOl4F
Williamson et at. (1996)
Williamson et at. (1996) Miyazaki et al. (1996)
Smith et at. (1997) Lee et at. (1999a) Blattella germanica
IIS6
LlOl4F
Miyazaki et at. (1996)
Haematobia irritans
IIS6
LlOl4F
Guerrero et at. (1997)
Anopheles gambiae
IIS6
LlO14F
Martinez-Torres et at. (1998)
Plutella xylostella
IIS6
Ll0l4F
Schuler et at. (1998)
Heliothis virescens
IIS6
LlO14H
Park and Taylor (1997)
Heliothis virescens
IIS6
LlO29H
Lee et at. (I 999b )
Dong (1997)
F, phenylalanine; H, histidine; I, isoleucine; L, leucine; M, methionine; N, asparagine; T, threonine; V, valine.
Dieldrin suppression of GABAA receptors accounts for its excitatory action, but the role of dieldrin potentiation remains to be seen.
generalize the role of the y2 subunit in drug-induced potentiation of GABA responses.
12.2.1 GABAA RECEPTOR SUBUNIT SPECIFICITY OF DIELDRIN ACTION
12.2.2 GABAA RECEPTOR SUBUNIT SPECIFICITY OF THE ACTIONS OF HCHISOMERS
The GABAA receptor consists of five subunits which form a pentameric structure (Nayeem et aI., 1994). There are at least six as, four f3s, four ys including long and short splice variants of the y2 subunit, one 8, and one p. Pharmacological sensitivity and profile are known to differ depending on the combination of these subunits. Dieldrin suppressed GABA-induced currents regardless of the three combinations, a1f32, a1f32y2S, and a6f32y2S, but dieldrin potentiation required the y2S subunit (Nagata et aI., 1994). Dieldrin was more efficacious in potentiating the current in the a6f32y2S than in the a1f32y2S combination, indicating some role of the a subunits in potentiating the current. Whereas the y2 subunit was required for benzodiazepine potentiation (Kurata et aI., 1993; Pritchett et aI., 1989) and zinc inhibition (Draguhn et aI., 1990; Smart et aI., 1991) of GABA responses, n-octanol potentiation did not require the y2 subunit (Kurata et aI., 1993). Therefore, it is not possible to
HCH comprises geometric isomers which exhibit different insecticidal activity. The y-HCH (lindane) is most toxic to mammals and insects and is a strong stimulant. The a-isomer is a weak stimulant, the f3-isomer is a weak depressant, and the 8-isomer is a strong depressant. Pomes et al. (1994) reported differential effects of HCH isomers on GABA-induced 36Cluptake by cortical neurons. Patch clamp experiments using rat DRG neurons showed differential actions of the four HCH isomers (Nagata and Narahashi, 1995). y-HCH had a weak potentiating action and a strong inhibitory action on GABA-induced currents. 8-HCH had a strong potentiating action and an inhibitory action. a-HCH and f3-HCH had little or no effect on GABA-induced currents. The differential modulation of GABA response by HCH isomers accounts for variable symptoms of poisoning in insects and mammals. However, somewhat different results were obtained for the effects of HCH isomers on the a1f33y2S
344
CHAPTER 12 control
Neurophysiological Effects of Insecticides
1 min
2 min
3 min 4 min
5 min
vv
if
500
PAL 50 sec
(a)
g 250
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(b) Figure 12.7 Effects of dieldrin on GABA-induced chloride currents in a rat dorsal root ganglion neuron. (a) Current records in response to 20-sec application of 10 ~M GABA (solid bar) and to co-application of 10 ~M GABA and 1 ~M dieldrin (dotted bar) at the time indicated after taking control record. The peak amplitude of current was greatly enhanced but gradually decreased during repeated co-applications. Desensitization of current was accelerated. (b) Time course of the changes in peak current amplitude before and during (horizontal line) repeated co-applications. From Nagata and Narahashi (1994).
and a6t33y2S subunit combinations of GABAA receptors expressed in Xenopus oocytes (Aspinwall et aI., 1997). GABA responses were inhibited by y-HCH, potentiated by a- and o-HCH, and not affected by t3-HCH. Furthermore, the a subunit composition had no influence on these effects of HCH isomers. These differences in the responses to chemicals represent an example of the dissimilarity between native receptors and receptors expressed in Xenopus oocytes which is often encountered. 8-HCH altered calcium homeostasis and contractility of cardiac myocytes through interaction with ryanodine receptors (Buck and Pessah, 1999). 8-HCH also induced a profound increase in ionic permeability in lipid bilayers, and the calciumdependent current produced by o-HCH was selective for monovalent cations (K+ » Cs+ > Na+) (Buck and Pessah, 1999).
Cyclodiene Resistance The first direct demonstration of a target site resistance mechanism for dieldrin and lindane was reported by Yamasaki and Narahashi (1958b). Multiple discharges from the housefly CNS were induced by these insecticides, and resistant strains were less sensitive than susceptible strains. While low nerve sensitivity to dieldrin was also re-
ported in resistant strains of Drosophila (Bloomquist et aI., 1992; ffrench-Constant et aI., 1991), it was not until 1993 that a point mutation in a Drosophila GABA receptor was found to be responsible for dieldrin resistance (ffrench-Constant et aI., 1993). The cyclodiene resistance gene Rdl (resistance to dieldrin) was cloned from Drosophila resistant to cyclodienes and picrotoxinin. Single amino acid replacement from alanine to serine (A302S) occurs with the second membrane spanning domain, which is the region lining the chloride channel pore. Subsequently, similar mutations of amino acids were discovered in several other insect species resistant to dieldrin: in addition to A302S replacement in Drosophila melanogaster, A302G as well as A302S was found in Drosophila simulans, and a single mutation A302S also occurred in Aedes aegypti, Periplaneta americana, Musca domestica, and Tribolium castaneum (Anthony et al., 1998; Buckingham et aI., 1996; Co1e et aI., 1995; ffrench-Constant, 1994; Miyazaki et aI., 1995). In addition to Rdl, another GABA receptor subunit was also cloned from insects which represents a homolog of the vertebrate GABAA receptor t3 subunit. Contrary to the vertebrate GABAA receptor subunits, Rdl could form a functional homomultimeric receptor. The Rdl receptor was sensitive to the blocking action of picrotoxin but insensitive to that of bicuculline. GABA receptors formed by Rdl plus t3 subunits were insensitive to picrotoxin but sensitive to bicuculline (Zhang et aI., 1995).
12.3 FIPRONIL Fipronil is a phenylpyrazole compound and was developed as a useful insecticide in the mid-1990s. It is effective against some insects such as the Colorado potato beetle and certain cotton pests that have become resistant to the existing insecticides. Fipronil is much more toxic to insects than to mammals, another advantage it has as an insecticide. Fipronil has been found to block insect GABA receptor (Rdl). Wild-type Rdl of Drosophila was suppressed by TBPS, 4-n-propyl-4'-ethynylbicycloorthobenzoate (EBOB), picrotoxinin, and fipronil (Buckingham et aI., 1994a; Millar et aI., 1994). Insect GABA receptors are different from vertebrate GABAA receptors in that they are not blocked by bicuculline (Benson, 1988; Buckingham et aI., 1994a; ffrench-Constant et aI., 1993; Millar et aI., 1994; Sattelle et aI., 1988), and are not potentiated by benzodiazepines and barbiturates (Millar et aI., 1994). The insensitivity to bicuculline is reminiscent of the GABAc receptor of vertebrates (Qian and Dowling, 1993; Woodward et aI., 1993). Dieldrin-resistant Drosophila melanogaster and D. simulans were also resistant to fipronil but to a much lesser extent, and the [3H]EBOB binding to these resistant strains was less inhibited by fipronil compared to susceptible strains (Cole et aI., 1995). Mutant Drosophila Rdl (A302S) expressed in Xenopus oocytes was also less sensitive to fipronil than wild-type receptors (Hosie et aI., 1995). Fipronil and desulfinyl derivative were more potent in houseflies than in mice as toxicants and in competing with eH]EBOB binding (Hainzl and Casida, 1996).
345
12.4 Imidacloprid
LDso values of fipronil were 0.13 mg/kg and 41 mg/kg for housefly and mouse, respectively, and receptor ICso values were 6.3 nM and 1010 nM for housefly and mouse, respectively. Fipronil block of GABAA receptors of rat DRG neurons has recently been analyzed in detail (lkeda et aI., 1999). Fipronil suppressed the GABA-induced whole-cell currents reversibly with an ICso of 1.66 ± 0.18 J.tM. Preapplication of fipronil through the bath suppressed GABA-induced currents without channel activation. These results indicate that fipronil acts on the GABA receptors in the closed state. From co-application of fipronil and GABA, the ICso value for the activated GABA receptor was estimated to be 1.12 ± 0.21 J.tM. The association rate and dissociation rate constants and the equilibrium dissociation constant of fipronil effect were estimated to be 673 ± 220 M-I sec-I, 0.018 ± 0.0035 sec-I, and 27 J.tM for the resting GABA receptor, respectively, and 6600 ± 380 M-I sec-I, 0.11 ± 0.0054 sec-I, and 17 J.tM for the activated GABA receptor, respectively. Thus, both the association and dissociation rate constants of fipronil for the activated GABA receptor are approximately ten times higher than those for the resting receptor, with a resultant lower Kd value for the activated receptor. Experiments with co-application of fipronil and picrotoxinin indicated that they did not compete for the same binding site. It is concluded that although fipronil binds to the GABAA receptor without activation, channel opening facilitates fipronil binding to and unbinding from the receptor. Single-channel recording experiments using the GABAA receptor of rat DRG neurons have revealed that fipronil prolonged the closed time without much effect on open time and burst duration (Ikeda et aI., 1999). Thus, fipronil reduces the frequency of channel opening, thereby suppressing the receptor activity.
12.4 IMIDACLOPRID A number of factors must be taken into consideration for developing new insecticides and for using existing insecticides, mammalian toxicity and insecticide resistance being among the most important. In order to cope with the situation, a new group of chemicals has been developed into commercial insecticides during the past 10 or so years, that is, nitromethylene or chloronicotinyl insecticides. Soloway et al. (1978, 1979) found that nithiazin was among the most insecticidally active nitromethylenes tested. Imidacloprid was later shown to have excellent insecticidal activities against leafhoppers, planthoppers, white flies, aphids, and various coleopteran insects (Elbert et aI., 1990,1991). Newer derivatives ofimidacloprid were also developed, including nitenpyram (Minamida et aI., 1993a, b) and acetamiprid (Takahashi et aI., 1992). Imidacloprid exhibits a unique mechanism of action on nicotinic acetylcholine (nACh) receptors. It bound to insect nACh receptors with a high affinity (Bai et aI., 1991; Buckingham et aI., 1995; Chao and Casida, 1997; Chao et aI., 1997; Lind et aI., 1998; Liu and Casida, 1993; Liu et aI., 1994). Imidacloprid depolarized nerve membrane and caused spontaneous discharges in cockroaches (Buckingham et aI., 1995;
SOms
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(a) 10 J1.M Imidacloprid
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10 J1.M ACh + 10 J1.M Imidacloprid
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(c) Figure 12.8 Single-channel currents activated by 10 J.!M ACh, 10 j.lM imidacloprid, and co-application of 10 j.lM ACh and 10 j.lM imidacloprid to cell-attached membrane patches clamped at a membrane potential 40 mV more positive than the resting potential in PCI2 cells. (a) Currents induced by 10 j.lM ACh occurred during brief isolated openings or longer openings interrupted by a few short closures or gaps. Main conductance state currents were observed more frequently than subconductance state currents. (b) Currents induced by 10 J.!M imidacloprid. Subconductance state currents were more frequently observed than main conductance state currents. (c) Co-application of JO J.!M ACh and 10 j.lM imidacloprid. Main conductance and subconductance state currents were induced, and channel openings were shortened. From Nagata et al. (1998).
Nishimura et aI., 1994, 1998; Sone et aI., 1994). Imidacloprid, acetamiprid, and nitenpyram also acted on Torpedo nACh receptors, but only as weak agonists (Tomizawa et aI., 1995). Mammalian end-plate nACh receptors were less sensitive than those of locust neurons (Zwart et aI., 1994). The effects of imidacloprid on single-channel activity of nACh receptors were analyzed in detail using PC12 cells (Nagata et aI., 1996, 1997, 1998). First, whole-cell currents were analyzed in the absence and presence of imidacloprid. Imidacloprid itself generated whole-cell currents with a low potency and efficacy. The minimum effective concentration was 1 J.tM, and the current amplitude reached a maximum at 30 J.tM. The imidacloprid-induced current was approximately 10% of the carbachol-induced current. Imidacloprid also suppressed carbachol-induced currents with a low potency: even at the maximum concentration tested (100 J.tM), imidacloprid suppressed the currents only by 30%. Single-channel analyses have
346
CHAPTER 12
Neurophysiological Effects of Insecticides
disclosed an interesting feature of imidacloprid action. Application of ACh induced primarily main conductance (25.4 pS) currents and some low conductance (9.8 pS) currents, while imidacloprid generated primarily the low conductance currents (Fig. 12.8a and b). Co-application of ACh and imidacloprid generated both types of currents (Fig. 12.8c). The mean open time and burst duration of the main conductance current were decreased by the co-application of ACh and imidacloprid. These changes in single-channel behavior by imidacloprid can account for the changes in whole-cell ACh receptor currents. Imidacloprid has both agonist and antagonist effects on the mammalian neuronal nicotinic ACh receptors. Nitenpyram behaved similarly to imidacloprid in modulating single AChinduced currents ofPCl2 cells (Nagata et aI., 1999). ACh receptor subunit specificity for imidacloprid action has recently been studied (Matsuda et aI., 1998). Imidacloprid was a partial agonist in generating currents in the recombinant chicken a4f32 subunit combination and in the hybrid receptor of Drosophila a subunit (SAD) with the chicken 132 subunit, both expressed in Xenopus oocytes. However, imidacloprid was more potent on the SADf32 subunit combination than on the a4f32 combination. Furthermore, imidacloprid was a weak potentiator of ACh-induced currents in the a4f32 receptors, whereas it was a weak antagonist of ACh-induced currents in the SADf32 receptors. Binding experiments indicated that imidacloprid, acetamiprid, and nitenpyram had low to moderate potency at the a3 and a4f32 ACh receptors and were essentially inactive at the al and a7 ACh receptors (Tomizawa and Casida, 1999). Insect ACh receptor subunits were also studied for imidacloprid action (Huang et aI., 1999). In the peach-potato aphid Myzus persicae, five a subunit cDNAs have been cloned: Mpa 1, Mpa2, Mpa3, Mpa4, and Mpa5. Although the insect a subunits evolved in parallel with the vertebrate neuronal nACh receptors, the insect non-a subunits are different from vertebrate neuronal 13 and muscle non-a subunits. The aphid nACh receptor a subunit cDNAs were co-expressed with the rat 132 subunit in Drosophila S2 cells. The affinity of recombinant nACh receptors for [3H] imidacloprid was a sUbtype dependent, being high in Mpa2 and Mpa3 subunits, but low in Mpal subunit.
ACKNOWLEDGMENTS Author's studies quoted in this chapter were supported by NIH Grant NS14143. Thanks are also due to Julia Irizarry for secretarial assistance and to Nayla Hasan for technical assistance.
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Stewart, G. R., Hennan, R. C., Chan, H., Eglen, R. M., and Hunter, J. C. (1997). A novel tetrodotoxin-sensitive, voltage-gated sodium channel expressed in rat and human dorsal root ganglia. J. BioI. Chem. 272, 1480514809. Sattelle, D. B., Pinnock, R. D., Wafford, K A, and David, J. A. (1988). GABA receptors on the cell body membrane of an identified insect motor neurone. Proc. R. Soc. London Ser. B 232, 443-456. Schuler, T. H., Martinez-Torres, D., Thompson, A J., Denholm, 1., Devonshire, A L., Duce, 1. R., and Williamson, M. S. (1998). Toxicological, electrophysiological, and molecular characterisation of knockdown resistance to pyrethroid insecticides in the diamondback moth, Plutella xylostella (L.). Pesticide Biochem. Physiol. 59, 169-182. Scott, J. G. (1998). a-Tocopherol antagonizes the toxicity of the pyrethroid insecticide pennethrin in susceptible and kdr-resistant insects. J. Pesticide Sci. 23, 399-401. Seabrook, G. R., Duce, 1. R., and Irving, S. N. (1988). Effects of the pyrethroid cypennethrin on I-glutamate-induced changes in the input conductance of the ventrolateral muscles of the larval house fly, Musca domestica. Pesticide Biochem. Physiol. 32,232-239. Seyama, 1., and Narahashi, T. (1973). Increase in sodium penneability of squid axon membranes by a-dihydrograyanotoxin 11. J. Pharmacol. Exp. Ther. 184,299-307. Seyama, 1., and Narahashi, T. (1981). Modulation of sodium channels of squid nerve membranes by grayanotoxin 1. J. Pharmacol. Exp. Ther. 219, 614624. Shapiro, B. 1. (1977). Effects of strychnine on the sodium conductance of the frog node of Ranvier. J. Gen. Physiol. 69, 915-926. Sherby, S. M., Eldefrawi, A. T., Deshpande, S. S., Albuquerque, E. X., and Eldefrawi, M. E. (1986). Effects of pyrethroids on nicotinic acetylcholine receptor binding and function. Pesticide Biochem. Physiol. 26, 107-115. Smart, T. G., Moss, S. J., Xie., X., and Huganier, R. L. (1991). GABAA receptors are differentially sensitive to zinc; dependence on subunit composition. Br. J. Pharmacol. 103, 1837-1839. Smith, T. J., Lee, S. H., Ingles, P. J., Knipple, D. c., and Soderlund, D. M. (1997). The LlOl4F point mutation in the house fly VsscJ sodium channel confers knockdown resistance to pyrethroids. Insect Biochem. Mol. BioI. 27,807-812. Smith, T. J., and Soderlund, D. M. (1998). Action of the pyrethroid insecticide cypennethrin on rat brain lIa sodium channels expressed in Xenopus oocytes. Neurotoxicology 19(6), 823-832. Smyth, T., Jr., and Roys, C. C. (1955). Chemoreception in insects and the action of DDT. BioI. Bull. 108, 66-76. Soderlund, D. M., and Bioomquist, J. R. (1989). Neurotoxic actions of pyrethroid insecticides. Annu. Rev. Entomol. 34, 77-96. Soloway, S. B., Henry, A. c., Kollmeyer, W. D., Padgett, W. M., Powell, J. E., Roman, S. A, Tieman, C. H., Corey, R. A., and Home, C. A. (1978). Nitromethylene heterocycles as insecticides. In "Pesticide and Venom Neurotoxicity" (D. L. Shankland, R. M. Hollingworth, and T. Smyth, eds.), pp. 153-158. Plenum Publishers, New York. Soloway, S. B., Henry, A. C., Kollmeyer, W. D., Padgett, W. M., Powell, J. E., Roman, S. A, Tieman, C. H., Corey, R. A., and Home, C. A. (1979). Nitromethylene Insecticides. In "Advances in Pesticide Science Part 2" (H. Geissbiihler, G. T. Brooks, and P. C. Kearney, eds.), pp. 106-127. Pergamon Press, Oxford. Sone, S., Nagata, K, Tsuboi, S.-I., and Shono, T. (1994). Toxic symptoms and neural effect of a new class of insecticide. J. Pesticide Sci. 19, 69-72. Song, J.-H., and Narahashi, T. (1995). Selective block oftetramethrin-modified sodium channels by (±)-a)-tocopherol (vitamin E). J. Pharmacol. Exp. Ther. 275, 1402-1411. Song, J.-H., and Narahashi, T. (1996). Modulation of sodium channels of rat cerebellar Purkinje neurons by the pyrethroid tetramethrin. J. Pharmacol. Exp. Ther.277,445-453. Song, J.-H., Nagata, K, Tatebayashi, H., and Narahashi, T. (1996). Interactions of tetramethrin, fenvalerate and DDT at the sodium channel site in rat dorsal root ganglion neurons. Brain Res. 708, 29-37.
Staatz, C. G., Bloom, A. S., and Lech, J. J. (1982). Effect of pyrethroids on [3Hlkainic acid binding to mouse forebrain membranes. Toxicol. Appl. Pharmacol. 64,566-569. Starkus, J. G., and Narahashi, T. (1978). Temperature dependence of allethrininduced repetitive discharges in nerves. Pesticide Biochem. Physiol. 9, 225230. Strichartz, G. (1973). The inhibition of sodium currents in myelinated nerve by quaternary derivatives oflidocaine. J. Gen. Physiol. 62, 37-57. Tabarean, 1. v., and Narahashi, T. (1998). Potent modulation of tetrodotoxinsensitive and tetrodotoxin-resistant sodium channels by the type 11 pyrethroid deltamethrin. J. Pharmacol. Exp. Ther. 24, 958-965. Takahashi, H., Mitsui, 1., Takakusa, N., Matsuda, M., Yoneda, H., Suzuki, J., Ishimitsu, K, and Kishimoto, T. (1992). NI-25, a new type of systemic and broad spectrum insecticide. Brighton Crop Protection Conferences-Pest and Diseases 1, 89-96. Takeda, K., and Narahashi, T. (1988). Chemical modification of sodium channel inactivation: separate sites for the action of grayanotoxin and tetramethrin. Brain Res. 448, 308-312. Tanguy, J., and Yeh, J. Z. (1991). BTX modification of Na channels in squid axons. 1. State dependence of BTX action. J. Gen. Physiol. 97,499-519. Tatebayashi, H., and Narahashi, T. (1994). Differential mechanism of action of the pyrethroid tetramethrin on tetrodotoxin-sensitive and tetrodotoxinresistant sodium channels. J. Pharmacol. Exp. Ther. 270, 595-603. Thompson, R. G., Menking, D. E., and Valdes, J. J. (1990). Comparison of lindane, bicyclophosphate and picrotoxin binding to the putative chloride channel sites in rat brain and Torpedo electric organ. Neurotoxicol. Teratol. 12,57-63. Tokutomi, N., Ozoe, Y., Katayama, N., and Akaike, N. (1994). Effects of lindane (y-BHC) and related convulsants on GABAA receptor-operated chloride channels in frog dorsal root ganglion neurons. Brain Res. 643, 66-73. Tomizawa, M., and Casida, J. E. (1999). Mino structural changes in nicotinoid insecticides confer differential subtype selectivity for mammalian nicotinic acetylcholine receptors. Brit. J. Pharmacol. 127, 115-122. Tomizawa, M., Otsuka, H., Miyamoto, T., and Yamamoto, 1. (1995). Pharmacological effects of imidacloprid and its related compounds on the nicotinic acetylcholine receptor with its ion channel from the Torpedo electric organ. J. Pesticide Sci. 20, 49-56. Trainer, V. L., McPhee, J. C., Boutelet-Bochan, H., Baker, C., Scheuer, T., Babin, D., Demoute J.-P., Guedin, D., and Catterall, W. A. (1997). High affinity binding of pyrethroids to the a subunit of brain sodium channels. Mol. Pharmacol. 51, 651-657. Tsukamoto, M., Narahashi, T., and Yamasaki, T. (1965). Genetic control ofIow nerve sensitivity to DDT in insecticide-resistant houseflies. Botyu-Kagaku (Scientific Pest Control) 30,128-132. Tucker, S. B., Flannigan, S. A., and Ross, C. E. (1984). Inhibition of cutaneous paresthesia resulting from synthetic pyrethroid exposure. Internat. J. Dermatol. 23,686-689. Vijverberg, H. P. M., and van den Bercken, J. (1990). Neurotoxicological effects and the mode of action of pyrethroid insecticides. Critical Reviews in Toxicology 21(2), 105-126. Vijverberg, H. P. M., van der Zalm, J. M., and van den Bercken, J. (1982). Similar mode of action of pyrethroids and DDT on sodium channel gating in myelinated nerves. Nature 295, 601-603. Vijverberg, H. P. M., van der Zalm, J. M., van Kleef, R. G. D. M., and van den Bercken, J. (1983). Temperature- and structure-dependent interaction of pyrethroids with the sodium channels in frog node of Ranvier. Biochim. Biophys. Acta 728, 73-82. Wafford, K A., Sattelle, D. B., Gant, D. B., Eldefrawi, A T., and Eldefrawi, M. E. (1989). Noncompetitive inhibition of GABA receptors to insect and vertebrate CNS by endrin and lindane. Pesticide Biochem. Physiol. 33, 213219. Warmke, J. W., Reenan, R. A., Wang, P., Qian, S., Arena, J. P., Wang, J., Wunderler D., Liu, K, Kaczorowski, G. J., Van der Ploeg, L. H., Ganetzky, B., and Cohen, C. J. (1997). Functional expression of Drosophila para sodium channels. Modulation by the membrane protein TipE and toxin pharmacology. J. Gen. Physiol. 110, 119-133.
References
Weiant, E. A. (1955). Electrophysiological and behavioral studies on DDTsensitive and DDT-resistant house flies. Ann. Entomol. Soc. Amer. 48, 489492. Wilkinson, C. F. (1983). Role of mixed function oxidases in insecticide resistance. In "Pest Resistance to Pesticides" (G. P. Georghiou and T. Saito, eds.), pp. 175-205. Plenum, New York. Williamson, M. S., Martinez-Torres, D., Hick, C. A., and Devonshire, A. L. (1996). Identification of mutations in the housefly para-type sodium channel gene associated with knockdown resistance (kdr) to pyrethroid insecticides. Mol. Gen. Genetics 252, 51-60. Wiswesser, W. J. (1976) "Pesticide Index," 5th ed., p. 328. Entomological Society of America, College Park, MD. Woodward, R. M., Polenzani, L., and Miledi, R. (1992). Effects of hexachlorocyclohexanes on y -aminobutyric acid receptors expressed in Xenopus oocytes by RNA from mammalian brain and retina. Mol. Pharmacol. 41,1107-1115. Woodward, R. M., Polenzani, L., and Miledi, R. (1993). Characterization of bicucullinelbaclofen-insensitive (p-like) y-aminobutyric acid receptors expressed in Xenopus oocytes. n. Pharmacology of y-aminobutyric acidA and y-aminobutyric acidA receptor agonists and antagonists. Mol. Pharmacol. 43,609-625. Wouters, W., van den Bercken, J., and van Ginneken, A. (1977). Presynaptic action of the pyrethroid insecticide allethrin in the frog motor end-plate. Europ. J. Pharmacol. 43,163-171. Yamamoto, D., Quandt, F. N., and Narahashi, T. (1983). Modification of single sodium channels by the insecticide tetramethrin. Brain Res. 274, 344-349. Yamamoto, D., Yeh, J. Z., and Narahashi, T. (1986). Ion permeation and selectivity of squid axon sodium channels modified by tetramethrin. Brain Res. 372,193-197. Yamasaki, T., and Ishii [Narahashi], T. (1954a). Studies on the mechanism of action of insecticides (VIII). Effects of temperature on the nerve susceptibility to DDT in the cockroach. Botyu-Kagaku (Scientific Insect Control) 19, 39-46. (In Japanese); English translation (1957) "Japanese Contributions to the Study of the Insecticide-Resistance Problem," pp. 155-162. Published by the Kyoto University for the World Health Organization. Yamasaki, T., and Ishii [Narahashi], T. (1954b). Studies on the mechanism of action of insecticides (X). Nervous activity as a factor of development of y-BHC symptoms in the cockroach. Botyu-Kagaku (Scientific Insect
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CHAPTER
13 Ecotoxicological Risk AssessIllent of Pesticides in the EnvironIllent Keith R. Solomon University of Guelph
13.1 INTRODUCTION Risk assessment is a process of assigning magnitudes, ranks, and probabilities to the adverse effects than may result from a particular activity or set of activities. Ecotoxicological risk assessment, as we use it today, had its scientific origins in the area of human health protection and has undergone many years of evolution since it was first implemented. In the context of physical risks, such as those associated with walking, running, climbing, or hunting large and dangerous mammals, or avoiding being hunted by large and dangerous animals, a form of risk assessment is probably conducted by animals (the consciousness of this decision-making process in animals is subject to debate) and its successful application was undoubtedly a major factor in the early evolution of humans on this planet. For human health risk assessment, it is important to distinguish between risk assessment and risk management decisions made on an individual basis from those made on a group basis. Risk assessment and risk management decisions made by individuals are part of an individual survival strategy and are common in the everyday life of all humans. Group or societal risk assessment and risk management decisions can only be made if a population, group, or society exists in the first place and are probably unique to humans. The process of risk assessment for individuals is very different from that of populations, groups, or societies. Risk assessment by single humans may be modified by individual differences in perception, lack of information, or individual stupidity. Some people are more likely to take individual risks than others. Individually, humans make risk assessment and risk management decisions every day (sometimes consciously, sometimes not). For example, simply turning on the light in the morning is a potential risk-in the United States about 200 people are electrocuted each year (Laudan, 1994). Walking downstairs to get your breakfast is even more risky, as is the choice of bicycle transportation over the Handbook of Pesticide Toxicology Volume J. Principles
use of an automobile (Laudan, 1994). These risks, based on actual or actuarial observations (Wilson and Crouch, 1987), are mostly examples of risks that are individually controllable and are therefore subject to perception, interpretation, and rationalization. Risks to the group or society are assessed differently from those that accrue to individuals. Societal risk management decisions (such as are taken by governments) are usually aimed at protecting the individual in the society. This individual is usually the weakest individual or the individual judged to be at highest risk, such as, for example, children. Ecological risks are usually perceived by humans in a similar fashion to group or societal risks. This is particularly true where there is a lack of identification with the organisms to which the risk accrues (Suter et ai., 1993). This is illustrated in the difference between the public acceptability of a risk that accrues to nearly invisible freshwater invertebrates and that which accrues to anthropomorphically identifiable creatures such as pandas or to endangered or threatened species. In the latter case, risks to these organisms may be treated in the same way as risks to individual humans. 13.1.1 RISK ASSESSMENT OF PESTICIDES Pesticides are deliberately used to control organisms for the protection of crops, human health, or structures. Use in these circumstances always occurs after some form of risk assessment has occurred in relation to the particular situation. For example, the cost of the pesticide may be a considered in relation to the benefit resulting from the control of the pest. Thus, the cost of a treatment to control termites in a structure may be assessed against the benefit of the structure lasting for a longer time and not having to be replaced. Similarly, an agriculturalist may consider the cost of the pesticide used to control an infestation of fungi in a fruit crop against the benefit resulting from increased value of the crop to the ultimate consumer, be this
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Ecotoxicological Risk Assessment
in improved quality or better storage properties. These types of risk assessments are familiar to most of the public and have been used for thousands of years in agricultural decision making. They are relatively easy to conduct because the risks (loss of the crop, etc.) are measured in financial units and the cost of the control measures is measured in the same units, and simple arithmetic can be used to determine if the risk of use (cost) is worth the benefit (profit). Integrated pest management (IPM), as commonly practiced in agriculture and other pests management situations, involves similar risk management decisions, although the risks are not as easily measured and the benefits are more indirect. For example, the use of an insecticide may be recommended for the control of an outbreak of an insect pest (A), but it may affect a beneficial mite that would control another pest (B). If the pesticide were used, pest A would be controlled, but pest B would be released from natural controls and may increase in numbers to the point that it may damage the crop to a greater extent than pest A. This risk assessment considers the effect of the pesticide on a nontarget organism in the agricultural unit, but the measures used to compare risk to benefits can ultimately be expressed in financial units, reducing the decision to more familiar terms. Risk benefit decisions made during pest control or pest management operations are internalized to the act of agriculture or crop production (Fig. 13.1). These decisions are made with very specific uses in mind. The history of the crop or animal production system is known and the particular pest situation may be known in great detail, especially in cases where pest numbers and life table information are available, such as in intensive IPM practices where scouting of pest populations and descriptive information on climate and other environmental factors are collected.
Although a large number of risk assessment and risk management decisions may be taken as part of the use of pesticides in pest management, these decisions are internalized to the production unit, the agricultural field, the structure, or the health district. Seldom do these decisions consider effects of the pesticide outside the area of specific use, such as may result from movement away from the area of application to areas that support nontarget organisms or to nontarget organisms that utilize the agroecosystem as habitat (birds, mammals, or other terrestrial organisms). These external risks are those that are focused on in environmental regulations and those that will be dealt with in more detail in this chapter. This focus does not imply that risk assessment in relation to management of pests in a particular situation of crop production or protection of health or property is unscientific or oversimplified, it is merely a recognition that risk assessment inside the system is different from that outside the system. This chapter is therefore devoted to the risk assessment of pesticides as it applies to pesticides that have moved off the agroecosystem and their effects are not part of the risk assessments conducted as part of production or to organisms that are exposed to pesticides because they make use of the agroecosystem as habitat or as a source of food. 13.1.2 ASSESSING RISKS FROM PESTICIDES IN RELATION TO OTHER SUBSTANCES Substances that have adverse effects in the environment have certain combinations of characteristics that lead to the possibility of adverse effects. These characteristics include the inherent toxicity of the substance and the potential for exposure to the substance. The toxicological properties of a substance are determined by its molecular structure and the biochemistry and
PESTICIDE RISK ASSESSMENT IN PEST MANAGEMENT
Risk assessments andriskmclnagej~e,D~~~
decisions made internally to the process of pest management and production
PESTICIDE RISK ASSESSMENT IN THE ENVIRONMENT Movement of pesticides off the agroecosystem to non-traget organisms in other parts of the ecosystem
Use of the ....... a~7rOecOsyste'm by organisms from "natural areas n
Agricultural field Figure 13.1
Agroecosystem including surrounding buffer zone Illustration of risk assessment and risk management decisions in pesticide use in agriculture.
13.1 Introduction
physiology of the organisms exposed to it. Exposure to the substance is also dependent on interactions between the chemical and physical properties of the substance and its environment. This is as true for pesticides as it is for all other substances. Contrary to public perceptions, pesticides do not possess special properties that make them chemically unique as a class or give them special toxicological or environmental properties that are not found in other substances. Depending on the structure of the molecule, the physical and chemical properties of pesticides (and other substances) span a large range of values. As for other substances, some pesticides are highly persistent, highly mobile, and highly toxic. It must be recognized that relatively few substances possess the necessary properties to place them in this category, however, pesticides are usually toxic to at least one class of organisms (otherwise they would not be used as such). However, the physical, chemical, and environmental properties of pesticides span the entire range observed in other substances. The process of pesticide risk assessment allows the proper identification and categorization of these substances according to their risks to the environment. 13.1.2.1 The Need for Pesticide Risk Assessment Pesticide risk assessment plays a crucial role strategic planning and priority setting, and in helping society to determine environmental or other priorities. Risk assessment is used in a number of forms by pesticide regulatory agencies in many
countries as well as internationally through the Food and Agricultural Organization (FAO) of the World Health Organization (WHO), which itself conducts risk assessment because of the use of these substances in the protection of public health. In the United States, the Federal Insecticide, Fungicide and Rodenticide Control Act (FIFRA) specifically requires risk assessment as well as risk benefit analysis. Canada has a Pest Control Products Act and many other jurisdictions such as members of the European Union, Japan, Australia, etc. , also have judicial instruments that require the use of pesticide risk assessment procedures in some form or another. 13.1.2.2 The Basic Concepts of Risk Assessment in Relation to Pesticides The basic concepts of the process of risk assessment have been discussed and developed in a number of documents (Environment Canada, 1997; NRC, 1993; SETAC, 1994; U.S. EPA, 1992, 1998) and books (Reinert et at., 1998; Suter et aI., 1993). The EPA Framework for Risk Assessment (U.S. EPA, 1992, 1998) has been taken as a general model, however, many of the other risk assessment schemes are essentially similar. The first step in the risk assessment framework (U.S. EPA, 1992, 1998) is "problem formulation" or the definition of the problem (Fig. 13.2). This is an important step because it lays down the foundation upon which the rest of the assessment depends. This step defines the objectives and scope of the entire
Discussion ECOLOGICAL RISK ASSESSMENT between the risk assessor and risk 14-- '" I manager (planning) ANALYSIS ,------------,
,------------,
Characterization of ecological l*--~ Characterization of exposure effects
Discussion between the risk assessor and risk manager (results)
Risk Management Figure 13.2
355
The framework used by the U.S. EPA for ecological risk assessment.
Data acquisition, verification and monitoring
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CHAPTER 13
Ecotoxicological Risk Assessment
process. This is important because ecological risk assessments are often complex, involving several biological components ranging from organism, population, community to ecosystem, and possibly several stressors and/or responses as well. A detailed plan is required at the onset of the assessment to identify and prioritize all of the issues that need to be addressed, and the data requirements of the assessment. Communication between the risk assessor and risk manager should occur during this phase as well, ensuring that all information required by the manager is provided by the assessment, and that all ecologically relevant issues are also addressed during the assessment (Fig. 13.2). This process may be informal in its initial stages, but will become more formal as one or more iterations are made through the risk manager and the risk assessor. The EPA framework for risk assessment (U.S. EPA, 1992, 1998) further divides problem formulation into five subsections: (1) stressor characteristics, (2) ecosystems potentially at risk, (3) ecological effects, (4) endpoint selection, and (5) the conceptual model. 13.1.2.3 Stressor Characterization
During stressor characterization, all potential stressors are identified and characterized by type (i.e., chemical or physical stressor), intensity, duration, scale, frequency, and timing-whether the stress occurs during an important biological cycle. For pesticides, the type of stressor is normally an organic or inorganic substance. Physical stressors may be important in other ecological risk assessments, but are less relevant to pesticides. Intensity refers to the exposure concentration or intensity of the stressor as experienced by organisms in the relevant environmental matrix (water, soil, etc.). The concentration of the stressor likely to result from its release into various matrices may also be useful for assessing organisms most likely to be affected. This is important in assessing risks from pesticides because environmental fate is an important determinants of exposure concentration. Duration and frequency of the release of the stressor are also important. Risk characterizations based on long-term, continuous releases will be very different from those based on sporadic, short-term releases. Likewise, labile substances must be assessed differently from more persistent substances. Timing in relation to biological cycles may also be important. This is particularly relevant in noncontinuous releases or the use of stressors such as pesticides where application is usually linked to the biological cycle of the pest organism and, by default, to those of other organisms in the environment. In the process of risk assessment as applied to IPM, concurrent cycles of beneficial organisms are taken into account in the choice of pesticide and the timing of application. In the case of long-term exposures, organisms may be more sensitive at some times of the year than others (spawning in fish) and risks may need to be assessed in relation to these events. Scale is important, both in terms of the range over which these effects occur and the degree of heterogeneity in the distribution of the stressor. Lack of mixing of an effluent stream in a river or lake may result in differential exposures and the treat-
ment of relatively small blocks of land with pesticides in agriculture and forestry may leave large untreated areas as refugia from which repopulation may rapidly occur. In fact, for many pesticides, annual or more frequent applications are required because pest populations will recover, either from protected individuals or as a result of immigration from untreated refugia (metapopulations ). For pesticides, hazard identification is easier than for many other environmental risk situations where the identity of the substance(s) is unknown. In the case of pesticides, the identity of the pesticide(s) plus any contaminants or formulants is known (see other chapters in this book). In addition, environmental breakdown products will be known as will pathways of metabolism in a number of species. In addition, because pesticides carry directions for use, the maximum amounts used per unit area are known and so is the frequency of use and region/crop type where the pesticide is to be used. This makes risk assessment for pesticides easier in one sense, but may misdirect attention to pesticides because there is generally more information available for them. 13.1.2.4 Receptor Characterization
The receptor organisms in the system may be well identified or unknown. In the former case, the motivation for the risk assessment may have been the observation of changes in populations or in community structure such as bird or fish kills or changes in distribution of plants. Where the pesticide has not yet been used in the environment, such field observations are not possible and laboratory tests on a range of organisms or systems (micro- and mesocosms) may be necessary to identify populations or communities at risk. This is often the case when a new pesticide is brought forward for initial registration. However, for pesticides, a great deal of additional relevant information is often available. For example, the mechanism of action may be known and, from this, likely sensitive organisms can be deduced. Even where the mechanism of action is not known, the wide-ranging screening tests used in the development of a pesticide may be used to identify most sensitive classes of organisms. The organisms or communities most at risk may also be more easily identified when the use pattern of the pesticide is use to characterize the system. 13.1.2.5 System Characterization
A similar process to receptor characterization occurs with characterization of the ecosystem potentially at risk. This involves the identification of the ecosystem and characterization of some of its properties such as the abiotic environment and the ecosystem structure and function. For example, with pesticides, a knowledge of the crop types upon which the pesticide will be used and when it is likely to be used will help to identify regions of use and likely nontarget organisms.
13.1 Introduction
13.1.3 ASSESSMENT ENDPOINTS AND EFFECT MEASURES Selection of appropriate responses upon which to base the risk assessment (assessment endpoints) is particularly important in any risk characterization. Poorly selected assessment endpoints have resulted in more risk assessment failures than any other possible error in risk assessment (Suter et al., 1993). An endpoint is a characteristic of a receptor that may be affected by the stressor (for example, mortality in birds, sustainability of a fish population). It is also important to realize that there are two types of measurements that are used in risk assessmentassessment endpoints and effect measures-and that these are not necessarily the same. Assessment endpoints are explicit expressions of the actual values that are to be protected. These are the ultimate focus in risk assessment and act as a link to the risk management process (such as the policy goals). Assessment endpoints usually have the following characteristics: They must be ecologically relevant, they should be susceptible to the stressor and, more controversial, they should have societal value. This last criterion, social value, must be treated with caution because it is subject to perceptual interpretation. For example, societal values may change over time. An example of this is the public regard for wetlands (Suter et al., 1993). Before the 1970s, the public saw little value in wetlands and many were, in fact, drained to reduce nuisance from mosquitoes or to increase cropland area. Since risk assessors and activists have brought to the attention of the public the linkage between wetlands and amphibians, birds, waterfowl, and flood protection, the social value of wetlands has increased and their functions are now accepted as a worthy of protection. This suggests that we should seek endpoints that have great ecological as well as societal relevance. This may require that the risk assessor educate the public on the value of the ecologically relevant endpoints. Assessment measures must also be unambiguous and must not be confused with policy goals. Thus trite statements such as "protection of indigenous populations" or "ecosystem health" have no place in risk assessment, even though they may be very laudable political or cultural objectives. These goals are usually not measurable and, therefore, progress and achievement cannot be gauged. Effect measures are responses that usually are more easily measured than assessment measures, but are, however, related quantitatively or qualitatively to the assessment measure. Therefore, effect measures should be directly related to a biologically significant responses. For example, a physiological or biochemical change in an organism (often referred to as an indicator, biomarker, or bioindicator) must be related to the survival or fecundity of the population before it can become a useful effect measure. The bioindicator is useless unless it is linked to relevant effects at the population level in organisms that are ecologically important in the ecosystem. A number of considerations must be taken into account in selecting effect measures (U.S. EPA, 1992, 1998). These include relevance to an assessment measure, consideration of indirect
3~'1
effects, sensitivity, and response time, diagnostic ability, and cost effectiveness and ease of measurement, but also of ecological significance. The availability of a large data base of effect measures is desirable to facilitate comparisons and to develop models. When an assessment measure can be measured directly, the assessment and effect measures are the same. However, this is seldom the case. Examples of assessment and effect measures are shown in Table 13.1.
13.1.3.1 Ecosystem Function and Ecological Risk Assessment of Pesticides Assessment and effect measures can be defined at all levels of organization in ecosystems, from that of the individual to that of the community. However, these are not necessarily of equal importance (Suter et al., 1993). In contrast to human health protection, individual organisms in the ecosystem are generally regarded as transitory and, because they are usually part of a food chain, are individually expendable (Suter et al., 1993). A self-maintaining or reproducing population is persistent on a human time scale and an be easily conceptualized by humans as being in need of protection. Thus, most assessment measures in ecological risk assessment are defined at the population rather than at the organismal level. Only in the case of the protection of rare, endangered, or long-lived species are organisms in the environment afforded similar protection to that enjoyed by humans. Generally, ecological risk assessment is aimed at protection of the functions of populations, communities, and ecosystems. This acknowledges the fact that populations are less sensitive than their most sensitive member and, likewise, that communities and ecosystems are less sensitive than their most sensitive components. Effects on a population are not necessarily of concern (to the ecosystem) as long as the functions of the popUlation can be taken over by other organisms. In this context, function is the interaction of the population with other populations or the abiotic environment. Functions in ecosystems are normally related to energy and nutrient flow: production of biomass (primary production), consumption of biomass (grazing or predation), controlling the abundance of other (prey) species, providing food to predators, or processing organic detritus, such as shredding plant tissue, macerating animal remains, and mineralizing organic compounds (Suter et aI., 1993). Functional redundancy is essential to the continuance of ecosystems in the face of natural stressors. Redundancy is the result of selection imposed by fluctuating and unpredictable environmental conditions. Most ecosystems exhibit functional redundancy, where multiple species are able to perform each critical function (Baskin, 1994; Walker, 1992, 1995). Functional redundancy is particularly relevant to ecotoxicological risk assessment. It is the basis for being able to tolerate effects in some sensitive populations because these effects are unlikely to impair the functions of the ecosystem as a whole. This is the basis for being able to tolerate some species being affected, such as in setting water quality guidelines (Stephan
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Table 13.1 Examples of Ecosystem Protection Goals, Assessment Endpoints, Effects, and Effect Measures Ecosystem goal
Assessment endpoint
Effect assessed
Effect measure
No unacceptable
Probability of a > I 0%
Toxicity of the
Fathead minnow
loss of fish resulting
reduction in fish
compound to fish in
from use of an
production
the laboratory
insecticide in
Toxicity of the
sugarcane
compound to food chain organism
LCSOs LCsos and ECs for
Daphia magna, Selanastrum capricornutum
Toxicity to fish under
Mortality of caged
field conditions
bass or other appropriate fish
Populations of fish in affected areas
Catch per unit effort or size/age ratios of age classes
No unacceptable
Probability of a >20%
reduction in
reduction in songbird
songbird
populations for a
populations in
period of more than 2
organophosphorus
years
Toxicity to songbirds in the laboratory
and white-throated sparrow
Toxicity to birds in
Number of bird carcases per hectare
the field
insecticide treated forests
LDSOs in Zebra finch
by species Poisoning of birds in the field
Brain acetylcholine esterase activities in birds exposed in the field
Adapted from Suter et al. (1993).
et aI., 1985). As illustrated in Fig. 13.3, there is a general relationship between exposure concentration and impact of any substance. However, there are deviations from this general rule. For example, functions may be maintained where few species are affected, but as the number of species affected increases, indirect effects amplify the effects of the substance to greater than predicted levels. Redundancy of function has been observed in a number of experimentally manipulated systems ranging from terrestrial (Tilman, 1996; Tilman et aI., 1996) to aquatic (Giddings et aI., 1996, 2000, 2001; Giesy et aI., 1999; Solomon et aI., 1996; Stephenson et aI., 1986). These observations support the concept that in ecotoxico1ogical risk assessment, some effects at the level of the organism and population can be allowed, provided that these effects are restricted on the spatial and temporal scale. In other words, they do not affect all communities all of the time and that keystone organisms are not adversely affected. In the context of selecting assessment measures, it has become increasingly recognized that these should be at the functional level of populations and the community, and that some effects on populations and species diversity may thus be tolerated. It is possible that a chain of events can occur whereby effects on one population may cause ripple effects throughout the ecosystem. These effects would occur in keystone species. Keystone species often supply physical habitat or modify the habitat in a way that cannot be replicated. Thus, removal of
Deviation resulting from interactions between species
Expected ecological effect
Deviation resulting from community resiliency and -"iIro-- - - - species redundancy
Concentration ----. Figure 13.3 Illustration of ecosystem resiliency in response to stressors and effects caused by interactions between organisms. (Adapted from Solomon and Takacs, 2001).
habitat (other populations) may be the root cause of the risk to a population designated for protection. An example of this is seen in the case of the spotted owl in the Pacific Northwest and removal of nesting habitat (James, 1994). Effects of pulp mill effluents on larval and juvenile fish in the Baltic is another example of the importance of habitat (Lehtinen et aI., 1991) as is
13.2 The Risk Assessment Analysis
the relationship between sea urchins, kelp, and sea otters (Estes et aI., 1998). Examples of these types of keystone responses are infrequently reported for pesticides in the nonagricultural environment, but are a very real problem in IPM situations within the agroecosystem. 13.1.3.2 Extrapolating from Single-Species Tests to Other Levels of Organization The costs of obtaining measurements of endpoints at the population level are often very great and, as a result, the bulk of toxicological testing has been focused at the organismallevel. It is thus necessary to extrapolate from data at the organismal level to that at the population or community. This concept is not foreign to risk assessment-most human health risk assessments are based on interspecies extrapolation from rodents or other test animals to humans. Extrapolation of data to another situation (data extrapolation) is more an act of faith than a statistical process, although some consideration may be given to scaling the exposure or the response. An example of this is the extrapolation of fathead minnow reproductive responses as measured in response to pesticide exposure in the laboratory to the reproductive success of other fish in a lake receiving pesticide runoff. When a pesticide is first registered, there is usually little toxicity data from which to extrapolate and conservative assumptions are normally used (Urban and Cook, 1986). When a pesticide has been in use for several decades, a larger data base may exist that allows for more realistic risk assessment (Giddings et aI., 2001; Giesy et aI., 1999; Solomon et aI., 1996, 2001a). 13.1.3.3 Uncertainty in Risk Assessment for Pesticides Uncertainty is a very important component of risk assessment because it influences the probability (risk) that an adverse effect will occur. As has been pointed out (Suter et aI., 1993), it is necessary to understand the uncertainties associated with the data that are used to make risk management decisions. Traditionally, risk assessors have used conservative assumptions rather than actually estimating uncertainty. This approach has several drawbacks: worst cases scenarios ignore probability of occurrence; may not be multiplicative or additive; are inconsistent because it is always possible to conceive of a still worst case; and are based on the premise that there are no societal or environmental costs resulting from the regulation of false positives (Suter et aI., 1993). Use of worst case assumptions is only applicable in early tiers of risk assessment where data sets are small and greater uncertainty exists.
13.2 THE RISK ASSESSMENT ANALYSIS After the risk assessment problem has been defined, the next two steps in preparing for the ecological risk assessment are the characterization of the effects of the pesticide and estimating exposures to the pesticide.
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13.2.1 CHARACTERIZING EFFECTS For organisms exposed through the matrix they inhabit (soil, air, or water), measurement of effects involves the consideration of three variables: exposure concentration, exposure duration, and response. Exposure concentration is a continuous variable and the proportion responding is a quantal variable. Duration of exposure is usually a continuous variable, but under field conditions, may be complicated by variation in the concentration during the exposure period. In the simplest case, the proportion of organisms responding is a function of concentration. This relationship is nonlinear, but may be analyzed by a linear model where the concentration is expressed as a logarithm and the percentage of organisms responding is expressed as a probability (probit scale). The most commonly reported number with respect to toxicity is the median response concentration. This the concentration at which half the organisms respond with the defined response. Commonly, the defined response is lethality (LCso), but it may also be a specific effect such as immobility of small organisms (ECso). The general use of the 50% response is because of the higher precision of the estimate and the narrower confidence intervals about this point. In practice, risk assessors are usually more interested in low or high proportions of responses. Other responses that a commonly used are lowest observed adverse effect concentration (LOAEC), the lowest concentration at which adverse effects are observed (this depends on the range of doses tested) or no observed adverse effect concentration (NOAEC), the highest concentration at which no effects are observed. Another measure of effect that is often used is the maximum allowable toxic concentration (MATC). This is the geometric mean of the NOAEC and the LOAEC, and is used to better approximate an exposure concentration equivalent to an incipient level of response. However, like the data that it is derived from, the MATC is dependent on the choice of concentrations used in the study. Presently, there is a move toward using a low level of response derived from a regression of the data points, in a similar way to the derivation of the benchmark response suggested for use in the interpretation of mammalian studies (U.S. EPA, 1995b). It is important to recognize that time is a critical dimension in the relationship between exposure and effect. This is particularly the case when exposure is through the matrix and uptake kinetics affect the concentration of the substance in the body of the organism. It is also important when pulsed exposures are considered. In almost all cases in the environment, exposures are not to constant concentrations, but to some form of pulse, the peak and duration of which is determined by the interaction between the pesticide and its environment. After initial application of the pesticides to the matrix, breakdown, adsorption, or hydraulic dilution occur. Exposures may thus be of unpredictable duration unless the specific characteristics of the receiving system are known. In practice, they range from very short to occasionally very long. Few compounds show a linear relationship between exposure concentration, exposure time, and effect (Giesy and
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Graney, 1989). For instance, pulsed exposures to greater concentrations may be less toxic than longer-term exposures to lesser concentrations. Although these exposures result in the same integrated concentration multiplied by time of exposure (J.!glL . h), they may not account for pharmacokinetic or repair processes, which control the effects that are observed. In answer to this concern, it has been suggested that customized toxicity tests of a specific duration be carried out (SETAC, 1994). A similar result can be achieved by assessing mortality at earlier time periods (e.g., 4 h, 8 h, etc.). A 48- or 96-hour study would allow data to be created that compare short-term exposure to short-term toxicity (ECOFRAM, 1999). This has the additional advantage of improving the statistical analysis to allow more accurate estimates of the effects (LCsos or ECsos) to be made (because more observations are available) and to allow the estimation of responses for any time period between those at which measurements have been made (ECOFRAM, 1999). However, it should be noted that observation of responses at a particular time interval during a test does not necessarily mean that toxicity will not change after that time if the organisms are moved to an uncontaminated medium (latency of response). Very few substances are instantly toxic and some time may have to elapse before the organism shows maximal response even though exposure has ceased. 13.2.1.1 Measuring Effects of Pesticides in Individual Species
Standardized test methods for pesticides are routinely used and required by a number of regulatory agencies. As in Europe (European Union, 1994), an evaluation procedure and a number of test methods have been developed by the U.S. EPA (U.S. EPA, 1982). The Organization for Economic Cooperation and Development (OECD) has published a number of methods (OECD, 1981, 1984) and American Society for Testing and Materials (ASTM) has a long history of methods development, testing, and validation (ASTM, 1991). Assessment procedures and test methods for pesticides have been reviewed (Lynch, 1995) and are constantly being updated, changed and, added to. To attempt to address all these methods in detail in this chapter would make the chapter too long and it would be out of date before publication. Instead, the general principles of the tests will be summarized. The basic principle behind the use of standardized laboratory toxicity tests is not that the particular organisms in the test are those that require protection in the environment, but rather that these organisms act as surrogates for all those other organisms in the ecosystem that could be exposed, but that for one reason or another cannot be tested in the laboratory. Because of this, test organisms are usually selected for ease of use (easy culturing techniques or easy availability in the environment) and because historical testing has shown that the species is particularly sensitive and, therefore, provides a worst case measure of effect. To make the effect measure even more conservative, the tests are normally conducted under conditions where the exposures are maintained at a constant concentration, usually
by continuous addition to a continuous flow treatment system. Most regulatory organizations now require that all toxicity testing be conducted under Good Laboratory Practice Guidelines (GLP) to ensure that the tests have been carried out appropriately and that proper quality controls have been implemented (OECD, 1992a-g, 1993a, b; U.S. EPA, 1986). For terrestrial systems, plants, microorganisms, invertebrates, and mammals may be tested. Testing of terrestrial plants is not yet widely required for pesticide risk assessment, but it should be recognized that a wide range of plants, including both crop and weed plants, are routinely tested in the discovery and development of a pesticide. These data give a general indication of sensitivity within classes of plants and can be used to assess likely sensitive nontarget species. Endpoints used in plant tests include germination, wet or dry biomass, or root elongation. For soil microorganisms (bacteria and fungi), nitrification (aerobic), denitrification (anaerobic), heterotrophic nitrification, and mineralization of nutrients may be used as effect measures. Tests on terrestrial invertebrates are directed toward assessing effects in beneficial invertebrates such as the honeybee (Apis mellifera) or other beneficial arthropods, various types of earthworms, such as Eiseniafetida, Lumbricus terrestris, or Enchytraeus albidus, and Collembola (springtails) such as Folsomia candida. Tests of terrestrial vertebrates are normally focused on mammals and birds. For mammals, the extensive testing on rodents and other laboratory test species used in the process of human health risk assessment is normally used as a surrogate for mammalian wildlife. Birds used in tests include the bobwhite quail, Japanese quail, and the hen. In most cases, oral toxicity is measured, either as a dose or as a concentration in the diet. Long-term feeding studies for either chronic or lifetime exposures may also be undertaken. For aquatic systems, plants, invertebrates and vertebrates are tested. Tests on aquatic plants are required for most herbicide registrations with the U.S. EPA (Urban and Cook, 1986) and in Europe (European Union, 1994). Organisms used include algae such as the freshwater algae Selanastrum capricornutum, Anabaena fios-aquae, Microcystis aeriginosa, and Navicula peliculosa and the saltwater algae Chlorella spp., Chlorococcum spp., Dunaliella tertiolecta, Isochrysis galbana, Nitzchia closterium, Skeletonema costatum, and Porphyridium cruentum. The freshwater macrophyte Lemna spp. (duckweed) is also used. Measures of effect include growth (as a percent of control) and cell numbers are usually reported as ECsos or ICsos. Measures of effect for algae are based on growth and the tests are normally run for periods of time that include many generations. Thus the ECsos reported are similar to those that would be expected from chronic studies in other organisms. For invertebrates in freshwater, Daphnia magna and Ceriodaphnia spp. are used for acute assay and chronic studies that include one life cycle. The midge, Chironomus spp. is normally used for acute assays only. For saltwater invertebrates, Mysidopsis bahia and Penaeus duorarum are used for acute and chronic assays for reproduction, mortality, and growth. The oyster embryolarval test Crassostrea spp. is also used in an acute assay. The need for testing of saltwater species is usually dictated
13.2 The Risk Assessment Analysis
by the use of the pesticide and the potential for contamination of saltwater systems such as estuaries. For aquatic vertebrates, standardized tests include acute, life-cycle tests and early lifestage tests. Some assays have been reported with amphibians (mostly larval stages) and are likely to become more widely required with increased interest in amphibian declines. Freshwater species used in acute tests include the rainbow trout (Onchorynchus mykiss), brook trout (Salvelinus Jontinalis), channel catfish (lctalurus punctatus), fathead minnow (Pimephales promelas), and bluegill (Lepomis macrochirus), whereas saltwater fish include the sheep she ad minnow (Cyprinodon variegatus), mummichog (Fundulus heteroclitus), longnose killifish (Fundulus similis), silverside (Menidia spp.), and the threespine stickleback (Gasterosteus aculeatus). Early life-stage tests may also be carried out. These tests focus on development and growth during the first 30-90 days of development from the egg. Species used include the fathead minnow (P. promelas), bluegill (L. macrochirus), brook trout (S. Jontinalis), flagfish (1ordanellafioridae), and sheepshead minnow (c. variegatus). Depending on several factors related to fate of the pesticide and results observed in other tests, full life-cycle tests may be required. Test organisms include fathead minnow (P. promelas), bluegill (L. macrochirus), brook trout (S. Jontinalis), flagfish (1. fioridae), and sheepshead minnow (c. variegatus). Measures of effect for acute studies normally include mortality, whereas early life-stage and life-cycle tests include measures related to growth, development, and reproduction. Although the preceding organisms are favored for regulatory testing, data from other species are also considered in pesticide risk assessment and may be submitted to the registration authorities, obtained from the open scientific literature, or databases such as ECOTOX (U.S. EPA, 2001). In these situations, the studies may not have been conducted according to specific guidelines or under good laboratory practice protocols and the data must be used with appropriate critical review. For aquatic organisms, the Quality Standards for the Great Lakes System (U.S. EPA, 1995a) criteria (GU) for acceptance of a study are useful. Preferred studies are those conducted under flow-through exposure conditions and with the concentration of the pesticide measured.
13.2.1.2 Measuring Effects at the Ecosystem Level Population-level assessment procedures (those carried out in populations of single species such as in chronic laboratory tests in organisms with a short life cycle) cannot take into account effects that involve interactions between populations of different species in communities or those that affect ecosystem function, such as recovery and changes in productivity or nutrient flow. Experimental testing of large populations of humans with a pesticide would be unlikely to receive public approval, but such studies are possible with ecosystems or subsets of ecosystems. A number of procedures have been proposed for ecosystem and community-level tests and there are numerous examples of their utility (Hill et aI., 1994). Most of this work has been carried out in aquatic systems, but some terrestrial systems have also been used. The aquatic systems range from
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simple laboratory systems to complex flowing stream systems, usually referred to as mesocosms or microcosms. Microcosm and mesocosm studies with pesticides provide effect measures that are closer to the assessment measures for several reasons (Solomon et aI., 1996). Measurements of productivity in microcosms incorporate the aggregate responses of many species in each trophic level. Because organisms will likely vary widely in their sensitivity to the stressor, the overall response of the community may be quite different from the responses of individual species as measured in laboratory toxicity tests. Microcosm studies allow observation of popUlation and community recovery from the effects of the pesticide and indirect effects of pesticides on other trophic levels. Indirect effects may result from changes in food supply, habitat, or water quality. Microcosm studies can be designed to approximate realistic stressor exposure regimes more closely than standard laboratory singlespecies toxicity tests. Most studies, especially those conducted in outdoor systems, incorporate partitioning, degradation, and dissipation, important factors in determining exposure. These factors are rarely accounted for in laboratory toxicity studies, but may greatly influence the magnitude of ecological response. As has been pointed out (Suter et aI., 1993), there are a number of difficulties associated with ecosystem-level assessments and assays. As originally proposed for use by the U.S. EPA, mesocosm studies were generally quite costly, were not standardized, and often had a large number of uncontrollable variables. Because of low replication, an analysis of variance (ANOVA) design with only three treatment concentrations, and a large number of measured parameters, statistical analysis of the data frequently produced false positive results and the studies lacked sufficient power to clearly demonstrate no effects. For this and other reasons, mesocosm and microcosm testing was discontinued as a regulatory requirement in the United States, but microcosms continue to be used in Europe. If the microcosm experiment is set up to test a specific hypothesis (SETAC, 1991), the results can be quite valuable for calibration of responses in laboratory tests (Giddings et aI., 2000,2001; Giesy et aI., 1999; Hall et aI., 1999; Solomon et aI., 1996) or to ask what-if questions relating to secondary and indirect effects. They have also been used to determine ecosystem- or community-level NOAECs (Giddings et aI., 1996,2000,2001; Giesy et aI., 1999; Liber et aI., 1992; Okkerman et aI., 1993; Solomon et aI., 1996; Stephenson et aI., 1986; Van den Brink et aI., 1996).
13.2.2 CHARACTERIZING EXPOSURE Millions of kilograms of pesticides move through the global ecosystem each year. During this process, many of these substances may be transformed into other products. Organic compounds are eventually transformed into simple compounds, such as carbon dioxide, ammonia, and water (mineralized), or incorporated into the biological cycle via small carbon units such as acetate. However, some may be rendered more toxic or more bioavailable in this process. Pesticides may be trans-
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Table 13.2 Environmental Processes, Driving Parameters, and Matrices Process
Driver and matrix
Photolysis
Light intensity (water and atmosphere)
Oxidation
Oxidant concentration (water, air, and soil)
Reduction
Reductant concentrations (water, air, and soil)
Rain-out
Precipitation rate, sticking coefficient (atmosphere)
Transport
Wind velocity (atmosphere); current velocity (water),
Vo1ati1ization
Henry's constant, surface texture (water and soil)
percolation (ground water); particle transport (soil runoff) Sorption
Organic matter, lipid, clay content (soil, sediment)
Bioconcentration
Lipid content of organisms (biota)
Hydrolysis
Temperature, pH (water)
Biotransformation
Organism populations, nutrient concentrations, temperature, pH and other factors that affect the energetics of organisms (biota)
formed, transported, or transferred to another matrix by a number of processes that are controlled by certain environmental parameters. These are shown in Table 13.2. For each pesticide, in each matrix, one or two processes usually dominate, however, the relative importance of these may change over time or space (i.e., photolysis may dominate during the day and hydrolysis during the night). Most of these processes result in smaller concentrations in the matrix, but for some pesticides such as DDT (now largely discontinued), bioconcentration results in higher concentration in the organism or through the food chain. Most toxicity tests for environmental risk assessment make use of exposures via the matrix in which the organism is present. Therefore, in the characterization of exposures in the environment, the concentration in the matrix is most important. Body dose could be measured in toxicity tests, but the equivalent measures from environmental samples may be difficult to obtain. 13.2.2.1 Measuring Exposure Measuring exposure in environmental matrices is one of the critical components of risk assessment, but is subject to errors through improper sampling techniques and sometimes by incorrect analyses. Obtaining an unbiased and representative sample from an environment may be very difficult and costly, and yet is probably the most important part of any exposure characterization. Sampling needs to consider both temporal and spatial heterogeneity of the pesticide residues. For example, the concentration of a pesticide may vary with water depth or distance from the shore immediately after a spray-drift contamination of water. Similarly, the concentration of a pesticide in flowing water may decrease over distance from the source of contamination due to breakdown in the water, adsorption to sediments, or dilution from uncontaminated water entering downstream of the source of the pesticide. Concentrations in soil may vary with the crop and soil type, and with the chemical and physical properties of the pesticide as well with climatic factors such as rainfall and percolation or leaching through the soil.
The objective of sampling the environmental matrix is to obtain a characterization of exposure that will be useful in the risk assessment process. Even with a good sampling design to address spatial heterogeneity, temporal variations in concentrations may be very important in assessing risks in relation to duration of exposure and choosing the appropriate exposure time for the toxicity data. Because of hydraulic flows in a headwater stream system, peak exposure concentrations in these systems may be very narrow and may be easily missed with a single daily grab sample. They would be incorporated into a continuous sampling system where daily integrated sampling was carried out, but very narrow peaks would be obscured. Sampling intervals should be designed to take into account the known hydraulics and breakdown kinetics of the pesticide in question. Thus, in small head water streams, more frequent sampling with a frequency of less than one day may be more appropriate. For slow-flowing rivers or for a rapidly degrading pesticide in a pond or reservoir, daily sampling may be adequate. For slowly degrading pesticides in stagnant pools, ponds, or reservoirs, even less frequent sampling may be needed. The concentration-time series of data that results from this type of sampling can then be analyzed by means of postprocessor tool such as the Risk Assessment tool to evaluate Duration And Recovery (RADAR) developed as part of the efforts of ECOFRAM (ECOFRAM, 1999). This tool provides information on pulse height, pulse width, and interpulse interval that is particularly useful for assessing likely effects on classes of organisms with known recovery times and time-exposure responses. Modem analytical methods may be exquisitely sensitive and be able to detect pesticide residues at amounts as small as 10- 12 g. In reporting residue data, it is normal also to report the limits of detection and the limits of quantification. The limits of detection (LOD) are method dependent and are often determined by the specific analytical equipment. In essence, the LOD is that amount that is statistically different from the blank. The LOD is usually set at a concentration 3 times the standard
13.2 The Risk Assessment Analysis
deviation (SD) greater than the mean of the blank (American Chemical Society Subcommittee on Environmental Analytical Chemistry, 1980). The limit of quantification (LOQ) is the smallest concentration to which a definite numerical value can be assigned with confidence. This concentration is greater than the LOD and is dependent on the repeatability or precision of the analytical method. By convention, the level of quantification is set at 10 SDs above the mean value for the blank. In practice, it is recommended that a response below the LOD be reported as not detected (ND). A measurement between the LOD and the LOQ can be reported as numbers along with the LOD in brackets or as a trace. Measurements of environmental concentrations of pesticides may contain a significant number of values below the LOD or LOQ. Conventionally, if a mean concentration is to be calculated, those numbers below the LODILOQ are assumed to be half the LOD. It is unlikely that all of the NDs would be exactly equal to any single fraction of the LOD. Therefore, a more realistic estimator of these concentrations is to assume a continuation of the distribution of the detected concentrations. Uncensoring methods can be used to estimate these numbers, but these would be useful only if a mean were to be calculated. A mean concentration that ignores all of the information inherent in a temporal or spatial array of data is not very useful in risk assessment. 13.2.2.2 Estimating Exposures In many risk assessments, the actual pesticide concentrations in the environment cannot be measured and risk assessors must make use of models to predict what these concentrations will likely be. All models, regardless of complexity, require input data. In general, as the problem definition becomes more complex and highly resolved, the model must become more complex and sophisticated. More complex models generally require more input data and a greater understanding of the system and the input data usually must be available at greater resolution. If the model is too simple, it may not be capable of answering the complicated question. On the other hand, if it is too complex, the available data and knowledge may be insufficient to accurately parameterize the model and its usefulness will be decreased. Also, models that are too complex can become difficult to interpret. Models may be used in Monte Carlo simulations, where measured or estimated distributions of input values are used to generate distributions of output values (ECOFRAM, 1999; Klaine et aI., 1996). Output from Monte Carlo simulations is useful for distributional and probabilistic analyses, but if the model is in error, the error is propagated through the entire data set. Use of Monte Carlo analysis also requires additional information on the distributions of input values, data that may not be available, thus forcing the use of default or assumed values. A number of models to predict surface water concentrations of pesticides exist. They simulate water flow, sediment deposition, and pollutant fate and transport in surface waters that receive dissolved runoff, soil erosion, and aerial deposition loadings. Outputs include dissolved concentrations of
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pesticide in the water column and concentrations adsorbed to suspended and bottom sediment as a function of time and location. One of the most widely used of these surface water models is EXAMS 11 that was developed and is supported by the U.S. EPA Athens Laboratory (Burns, 1997). EXAMS 11 can accept multiple loadings to multiple stream segments as input. The model can give contaminant concentrations in the water column and benthic sediment as a function of time for a given stream segment or as a function of stream segment at a given time. In addition, EXAMS can simulate some specific dissipation processes, including hydrolysis, direct photolysis, oxidation/reduction, biodegradation, and volatilization. Soil runoff models simulate water and soil movement along with contaminant fate and transport to the edge of the field, and from the soil surface through the vadose zone to at least the bottom of the root zone. Outputs include event, monthly, and annual runoff volumes and sediment yields along with their associated dissolved and adsorbed contaminant loadings to surface waters. They are commonly used to model pesticide movement in the soils. PRZM-2 (Mull ins et aI., 1993) is a field-specific, daily time scale model that can be used to estimate runoff, leaching, and associated contaminant loadings. This model allows for different inputs of soil erosion, cropping, and management factors over time; it simulates foliar as well as soil degradation, foliar interception, foliar washoff, plant uptake, crop rotations, daughter products, and some types of pesticide application, tillage, and irrigation methods. 13.2.2.3 Models for Estimating Exposure to Pesticides As a first simple step in assessing likely exposure concentrations of pesticides, several single-compartment formulae and tables are available to estimate initial or maximum exposure concentration for a pesticide. These are useful as initial screens to estimate the highest likely concentrations of pesticides. Drift of pesticide sprays is influenced by a number of processes, but exposure concentrations through drift are normally reduced with distance from the site of application. There are several approaches to assessing spray drift. One method is to assume that drift is equal to 5% of the application rate (as was the case with an earlier version of the GENEEC model, which is subsequently discussed). Another is to use models such as AGDRIFT, a spray drift model currently under development for use in the United States (Teske and Scott, 2000). Alternatively, actual measurements of drift resulting from the use of pesticides in the field can be used, as is currently the case in Germany and other parts of Europe (Ganzelmeier et aI., 1995). The German (BBA) drift tables were developed from drift deposition values measured in a number of actual applications of pesticides with similar equipment and crop types, but in different locations and times. The data were analyzed statistically and the values presented in the final tables represent the 95th centile drift deposition values for the replicate studies. Thus, although the data are derived from a distribution, the single values derived from the table represent reasonable worst-case deposition
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rates. The BBA drift table was developed from data generated in Europe; deposition rates may differ in other crops and environments as well as with the type of application equipment. It can, however, be used as a guideline in areas with similar crops and where similar environmental factors will affect drift. Several assumptions can be made with regard to pesticide concentrations in water. The U.S. EPA assumes a water depth of 2 m (based on design specifications for farm ponds) and a direct overspray with the pesticide (SETAC, 1994). In Europe it is assumed that the water depth of a pond is 30 cm (Riley, 1993). In Canada, forest pools are assumed to be 15 cm deep and a similar assumption is made by the U.S. EPA when considering wetlands. Assuming complete and rapid mixing, simple volumetric calculations give worst-case concentrations of 50, 333, and 670 I-lg/L, respectively, resulting from a direct surface application of 1 kg/ha. Concentrations from other application rates can be determined by simple proportion. Assumptions for the calculation of soil concentrations of pesticides are based on a direct application to soil and mixing in the upper 2.5 or 5 cm of soil. If the pesticide is incorporated (ploughed into) into the soil, mixing is assumed to extend to a depth of 20 cm. If a cover crop is present, it may be assumed to intercept 50% of the spray. Based on a soil density of 1.5 kg/ha, initial concentrations would be 2.67, l.34 and 0.34 mg/kg from a 1-kg/ha application rate. Interception by the crop will reduce the concentration in the soil and is normally assumed to be 50%. For beneficial arthropods (bees) it is assumed that they receive a direct application of the pesticide. Based on empirical observations, a general guideline has been suggested for assessing hazard to honeybees (Felton et aI., 1986). If the ratio of g of pesticide active ingredient (AI) applied per hectare divided by the LDso for bees (I-lg Allbee) is less than 50, bees are judged to be at low risk. If the ratio is 50-2500, a moderate risk is judged to exist and if it is greater than 2500, a high risk exists. Assumptions used to assess the exposure of birds and wildlife are based on estimates of contamination of foliage, seeds, and insects used by the FAO (1989). These are similar to the Kenaga Nomograms used by the U.S. EPA (Urban and Cook, 1986) as recently reassessed (Fletcher et al., 1994). Small birds (up to 100 g) are assumed to eat a maximum of 30% of their body weight in seeds or insects in a day, whereas larger birds (500 g) are assumed to consume 10% of their body weight in a day. Based on an application rate of 1 kg/ha, foliage, seeds, large insects, and small insects would be expected to contain 200, 10, 10, and 100 mg/kg wet weight of pesticide. For a small bird (lOO g), the expected dose would be expected to be 60, 3, 3, or 30 mg/kg body weight if they ate foliage, seeds, large insects, or small insects, respectively. For a large bird (500 g), the dose would be 20, 1, 1, or 10 mg/kg, respectively. All of these assumptions of exposure concentration are derived from averages of measures. They are deterministic values that are best used for lower tiers of risk assessment and do not consider heterogeneity in spray deposition, either within the agroecosystem or as a result of spray drift. Incorporation of measures of heterogeneity such as exist in the raw data from which the BBA drift tables were developed (Ganzelmeier et at.,
1995) would allow use of distributional approaches to characterize exposure concentrations. Several multi compartment models for estimating pesticide concentrations in environmental media are available. The most simplified of these is GENEEC. GENEEC version 1.3 (Parker, 1999) mimics a PRZMIEXAMS simulation of a generic 1O-ha row crop field draining into a I-ha farm pond of depth 2 m. It incorporates spray drift to estimate concentration in water at various times after a contamination event and has a choice of several crop types. GENEEC is designed as a Tier 1 model. It is conservative and only gives one output for each use scenario. A more complex combination of EXAMS and PRZM has been used with a preprocessor called the multiple scenario risk assessment tool (MUSCRAT; ECOFRAM, 1999). MUSCRAT, is an application program that links chemical, crop, soil, and climate data bases and facilitates the creation of PRZM-3 and EXAMSII input files, batch processes multiple model simulations, and performs statistical analyses on predicted exposure concentrations for pesticides (ECOFRAM, 1999). MUSCRAT gives multiple values as output and these output values can be analyzed as distributions rather than as single deterministic values. MUSCRAT is designed to provide mode led data for use in higher tiers of the probabilistic risk assessment process.
13.3 RISK ASSESSMENT OF PESTICIDES Risk assessment for pesticides is generally done by comparing the concentration of pesticide estimated or found in the environmental matrix to response concentrations reported for that pesticide in the laboratory. As has been recommended numerous times, risk assessments should be conducted in a series of steps or tiers (ECOFRAM, 1999; SETAC, 1994). The use of tiered approaches in risk assessment has several advantages for the risk assessor and those being assessed. The initial use of conservative criteria allows substances that truly do not present a risk to be eliminated from the risk assessment process, thus allowing the focus of expertise to be shifted to more problematic substances. Progression through the tiers results in estimates of exposure and effects becoming more realistic as uncertainty is reduced through the acquisition of more data. Tiers are normally designed such that the lower tiers in the risk assessment are more conservative (less likely to pass a hazardous chemical), whereas the higher tiers are more realistic, with assumptions more closely approaching reality. Because lower tiers are designed to be protective, failing to meet the criteria for these tiers is merely an indication that a more data-rich and more realistic risk assessment is needed. Risk assessment of pesticides can be conducted for many reasons. These range from simple ranking systems to more complex probabilistic risk assessments (Fig. 13.4). 13.3.1 SCORING SYSTEMS
The least complex form of risk assessment is the use of scoring systems to rank substances on the basis of toxicological or other
13.3 Risk Assessment of Pesticides
Ranking of concems in the absence of specific exposure information
Assessment of hazard based on a ratio of single deterministic exposure and toxicity values
365
Assessment of risk based on likelihood of exposure and/or toxicity
Figure 13.4 Illustration of various types of risk assessment from ranking/scoring systems to probabilistic approaches. (Redrawn from Solomon and Takacs. 2001).
properties. A number of scoring systems are used by national and international organizations. The International Joint Commission has used a scoring system (HC, 1993) to identify candidate substances for virtual elimination and the Ontario Ministry of Environment and Energy (OMEE) developed a scoring system to assess environmental contaminants (OMEE, 1990). The WHO uses a ranking scheme based on pesticide toxicity to classify pesticides into categories that can be used to regulate access to the products as well as transportation and storage. Many jurisdictions have followed the WHO ranking scheme. The basic principle of a scoring system is to assign a rank or priority to a list of substances. This is usually accomplished by assigning a score to several of the properties of the substances being assessed, manipulating these scores in some way or another, and then using the scores to rank (and select) some of these substances for further action. Some scoring systems use single criteria for a property (greater than/less than), whereas others may use multiple criteria (discrete values or ranges) that are assigned numerical scores. Very few scoring systems use decision criteria for multiple values, that is, where different data sources report different values. Most systems use the most conservative (worst-case) value, regardless of source, provenance, or general applicability. Scores may be combined in several ways. Some scoring systems use simple arithmetic procedures to combine scores assigned to different properties of the pesticide, whereas others make use of more complex algorithms. Yet others use the scores in a taxonomic key, where, after the scores have been evaluated in a series of questions, the compound is classified into a particular category. This has the advantage that scores may be combined in specific ways for different combinations of properties and that the system may be integrated into a computerized expert system. Correctly used, scoring systems have been employed to rank substances in order of priority for further assessment. This is usually part of the first tier of the risk assessment process, where
comparisons are needed for assignment of priorities. Before being used in a final risk management decision, a more detailed risk assessment would be required because the scoring systems rarely consider exposures, commonly make use of worst-case data, and do not handle missing values, weighting, or scaling in clear or appropriate ways. The rank numbers produced from combinations of scores have no meaning in the real world, their only use is to allow prioritization of substances for more detailed assessment. Incorrectly used, scoring systems have been and are employed in place of a full risk characterization. This may be politically expedient because it allows rapid action, but it may result in very controversial or costly decisions being made on the basis of poor science.
13.3.2 THE HAZARD QUOTIENT The first real tier in the risk characterization process is the use of hazard quotients (Fig. 13.4). These are simple ratios of single exposure and effects values, and may be used to express hazard or relative safety. For example, the calculation of hazard quotients has normally been conducted by utilizing the susceptibility of the most sensitive organism or group of organisms and comparing this to the greatest exposure concentration measured or estimated in the environmental matrix. This may be made more conservative by the use of an uncertainly (application) factor (CWQG, 1999), for example, division of the effect concentration by a number such as 20. This is done to allow for unquantified uncertainty in the effect and exposure estimations or measurements. In this case, if the hazard ratio is greater than 1, a hazard exists. Under the guidelines for pesticide risk assessment currently used in the United States (Urban and Cook, 1986), the hazard ratio is referred to as the level of concern (LOC) and different LOCs are used for different classes of organisms, depending on the nature of the effect measure or whether endangered species are likely to be affected
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(Urban and Cook, 1986). These criteria range from 0.05 to 1. In the European Union, the margin of safety concept is used instead of the hazard ratio [it is called the toxicity exposure ratio (TER)]. The acceptability criteria for the TER may depend on amount of data available, but a ratio of 10 to 100 is normally deemed acceptable (European Union, 1994). All hazard ratio assessments incorporate some form of uncertainly factor, either implicitly as part of the calculation itself or in the criteria for acceptance of the hazard ratio. In the absence of an adequate range of toxicity tests or exposure measurements, decisions based on hazard quotients may be underprotective and the use of uncertainty factors is justified. However, where an acceptable range of toxicity data is available, the inherent variation in receptor response is better defined and use of safety factors may be overprotective. The quotient approach is thus acceptable for early tiers or preliminary risk assessments, but it fails to consider the range of variation that may exist in terms of exposures and susceptibility when more data are available. 13.3.3 PROBABILISTIC RISK ASSESSMENT
The probability of occurrence of a particular event is and has been widely used in the characterization of risk from many physical and medical events in human society (the insurance industry) and for protection against failure in mechanical and civil engineering projects (time between failures, one-in-onehundred-year floods, etc.). This concept has been applied in ecotoxicological risk assessment for the characterization of distributions of both exposures and effects. Concentrations of substances in the environment can be affected by a large number of processes that relate to the amount released, the spatial and temporal distributions of the releases, and the results of the action of a large number of transportation and transformation processes (fate processes) on the substance. The likelihood and extent to which these myriad of fate processes will affect a particular quantity of substance in the environment is essentially random, and frequency distributions of exposure concentrations in the environment can be used to describe and characterize the data set and to use it to make predictions similar to those made in other areas of risk management (Carrington, 1996; McBean and Rovers, 1992). In many cases, these distributions fit the log-normal model reasonably well (Klaine et aI., 1996; SETAC, 1994; Solomon et aI., 1996; Solomon and Chappel, 1998). The assumption of a reasonable fit to a model makes calculations of exceedence probabilities relatively easy, but is not necessary for the concept of probabilistic risk assessment to be used. Centiles of distributions may be estimated from large data sets by simple ranking and interpolation or by using a suitable model, such as a polynomial, to describe the relationship. This is best used with large data sets where extrapolation beyond the observations is not needed. The same observations of log-normality generally apply to distributions of toxicological data. Many of the reactions through which toxicity mechanisms are mediated are first-order or pseudo first-order and, with a large enough data set and appropriate groupings of organisms to avoid mixing susceptible
and non susceptible species, good fits to the normal distribution are obtained (Giddings et aI., 2000; Giesy et aI., 1999; Solomon et aI., 1996, 2001a; Solomon and Chappel, 1998). Some care should be taken when using exposure or toxicity data in the distributional analyses. Exposure data should be screened to make sure that the data are consistent. Ideally, exposure data should be expressed over constant intervals such as daily samples. Distributional analysis of a data set with unequal time intervals will distort the distribution to overrepresent periods where more samples were taken. In this case, samples taken more frequently can be combined as time-weighted averages. In situations where samples are taken less frequently, interpolation can be used to "generate" data with the proviso that this will introduce a bias into the data set. In temperate regions, sampling of environmental matrices may not be carried out in winter. Such winter is a period of low biological activity, it may be more ecologically appropriate to focus risk assessment on the more biologically productive months when more analytical data are available. Effects data used in distributional analyses should be reviewed for appropriate quality (see preceding text). However, particularly with older pesticides, multiple studies on the same species that satisfy the appropriate quality criteria may be used. If this is the case, several procedures may be adopted. If one of the data points represents a more sensitive life-stage and life-table analyses indicate that survival of this stage is key to population sustainability, then this datum should be used. If no life-stage data are available and/or multiple tests still satisfy the review criteria, it is recommended that test data be combined as a geometric mean to determine a measure of central tendency (ECOFRAM, 1999). Pesticide toxicity data reported at concentrations in excess of the maximum water solubility of the pesticide may not be reliable descriptors of responses; however, they can be used for risk assessment. These data are almost always from the least susceptible organisms and, although they are less relevant in the risk assessment, they can be used in the characterization of the toxicity distribution. Because the data are less reliable, they should not be used in fitting a model to the distribution, but should be included in the calculation of the total sample size (n) and the ranks. The toxicity and exposure data are thus analyzed as distributions on the assumption that the data represent the universe of observations. Obviously, it is not possible to test all the species in an ecosystem and, for this reason, an approximation is usually made and the test data are used as a surrogate. The same is true of exposure data, because it is not practical or feasible to sample all possible locations or times. Whereas it is unusual for sufficient toxicity and exposure data to be available to allow a cumulative frequency distribution of data to be plotted directly, an approximation is normally used (Parkhurst et al., 1996). The data are ranked and then mathematically expressed or displayed graphically as log concentration versus percent probability. The y-plotting positions are calculated as percentages using the formula [100 x i/(n + 1)] (from Parkhurst et aI., 1996), where i is the rank number of the datum point and n is the total number of data points in the set. This gives an empirical cumulative
13.3 Risk Assessment of Pesticides
probability based on the Weibull equation. Similar empirical probabilities can also be calculated using other formulae such as the Blom equation [P = (i - 0.375)/(n - 0.25) x 100] or the Hazen equation [P = (i - 0.5)/n x 100] (Cunnane, 1978). These two equations may be useful for small data sets (Cunnane, 1978). These formulae all compensate for the size of the data set. Small (more uncertain) data sets are more likely to give more conservative estimates of high or low centiles than larger (more certain) data sets. The principle of probabilistic approach has been described (Cardwell et aI., 1993; Giesy et aI., 1999; Klaine et aI., 1996; Parkhurst et aI., 1996; SETAC, 1994; Solomon, 1996; Solomon and Takacs, 2001) and is illustrated diagrammatically in Fig. 13.5. Distributional analysis can be applied to concentrations of substances in the environment with due consideration for the fact that these data are usually censored by the limits of analytical detection (Fig. 13.5A). In practice, all exposure concentration data below the LOD or LOQ are assigned a dummy value of zero. These data are used in the calculation of n, but are not plotted or used to estimate centiles. As in this illustration, when plotted as a cumulative frequency distribution using a probability scale on the y axis as a function of loglO concentration (Fig. 13.5B), these distributions approximate a straight line, which can be used to estimate the likelihood that a particular concentration of the substance will be exceeded. A similar approach can be taken with susceptibility of organisms to the substance (Fig. 13.5C and D). The combination of these distributions in the probabilistic characterization of risk is illustrated in Fig. 13.5E. In this procedure, it is assumed that the distributions of sensitivity represent the range of responses that are likely to be encountered in the ecosystems where the exposures occur (SETAC, 1994). If the exposure data were collected over time at a particular site, the degree of overlap of the exposure distribution with the effects distribution can be used to estimate the joint probability of exposure and toxicity, leading to estimates of exceedence probabilities for responses at a fixed effect assessment criterion, such as, for example, the concentration equivalent to the 10th centile of the species distribution (Fig. 13.5E). This can be applied to a number of data sets and the resulting probabilities can be used for priority setting or to further assess ecological relevance. Expressing the results of a refined risk assessment as a distribution of values rather than a single point estimate is an approach that has been used by the Dutch government (Health Council of the Netherlands, 1993) and recommended for use in ecological risk assessments of pesticides (ECOFRAM, 1999; SETAC, 1994). This approach has been used in a number of risk assessments of pesticides (Cardwell et aI., 1999; Giddings et al., 2000,2001; Giesy et aI., 1999; Hall et aI., 1999; Klaine et aI., 1996; Solomon et aI., 1996, 200la, b; Solomon and Chappel, 1998; Versteeg et aI., 1999). The major advantage of this approach is that it uses all relevant single species toxicity data and, when combined with exposure distributions, allows quantitative estimations of risks. In addition, the data may be revisited again, the decision criteria become more robust with
367
additional data, and the method is transparent (will give the same results with the same data sets). The method does have some disadvantages. More data are usually needed, although these are mostly low cost studies. For new products, models have to be used to estimate exposures and models have not been widely validated for these uses. It is not easily applied to highly bioaccumulative substances where exposure is via the food chain as well as the matrix; however, if appropriate data are available, this can be overcome. Probably most critical of all is that it requires education of risk assessors and risk managers to increase their ability to evaluate decisions and to increase their comfort levels with the process (Bier, 1999; Solomon, 1996). Probabilistic risk assessment can be applied to assessments based on acute or chronic responses. All that is necessary is to ensure that the toxicity data and the exposure data are expressed in the same units. Although more widely used to assess risk to aquatic systems, the techniques are applicable to terrestrial systems as well (ECOFRAM, 1999; Solomon et aI., 2001a). 13.3.3.1 Probabilistic Risk Assessment for Mixtures
Pesticides are rarely used in consistent mixtures or applied in such a way that they will enter the environment in predictable combinations of concentrations and, because pesticides are regulated as single substances, risk assessments of pesticides have traditionally been conducted on one active ingredient only. These same constraints apply to probabilistic risk assessments. However, environmental monitoring has shown that pesticides do co-occur in the same location and time (Giddings et aI., 2000) and this has raised questions regarding risk assessment of mixtures. Where substances are known to act additively, it is possible to use the toxic equivalent (TE) or toxic unit (TU) approach to add concentrations and assess risks from the mixtures. This has been applied to several classes of compounds such as the dioxins (Ahlborg et al., 1994; Parrott et aI., 1995), chlorinated phenols (Kovacs et aI., 1993), and polyaromatic hydrocarbons (Schwarz et aI., 1995). This approach has been used to assess the combined risk from atrazine and its metabolites (Solomon, 1999) using probabilistic approaches, and could be used to assess risks from other pesticides with a common mode of action, such as the organophosphorus pesticides. The approach is similar to that used for dioxins in that the concentrations of the components of the mixture are converted to concentration equivalents of one reference component of the mixture, usually the most toxic. These concentrations are then added and compared to toxicity values for the reference compound. Traditionally, these equivalents have been based on responses measured in the same organism, for example, the laboratory rat. This is appropriate if the risks are to be assessed in the same organism or extrapolated to another (humans) with appropriate uncertainty factors. However, ratios of potencies measured in one animal may not be the same in another and wide interspecific extrapolations, such as from rats to fish, may
368
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not be possible (Parrott et aI., 1995). This situation becomes more complex when dealing with ecological risk assessments. If the toxic equivalents are based on responses measured in a single species, the relationship between the potency of the components in the mixture may be different from those in other organisms. Thus, as was the case with dioxins in fish and rats, extrapolation of the risk, whether assessed via hazard quotients or probabilistic risk assessment, could be incorrect. Alterna-
tively, TEs could be based on point estimates of potency derived from distributions of toxicity values. Again, use of these TEs would require that the distributions were similar (same slope when expressed as cumulative frequency curves) and that the order of the species in the distributions was the same. If this is not the case, risk assessments may be incorrect. Because of the number of uncertainties in the use of probabilistic risk assessment for mixtures, there is a clear need for further research.
13.3 Risk Assessment of Pesticides 13.3.3.2 Refinements to the Probabilistic Risk Assessment Process
A number of refinements to the probabilistic approach have been suggested by the ECOFRAM working group (ECOFRAM, 1998, 1999). Many pesticides have some degree of specificity in their mechanism of action. For example, herbicides may be selectively toxic to some groups of plants (weeds versus corn) as well as be less toxic to animals and other organisms that do not possess the receptor system (say photosynthesis). Similarly, an insecticide that acts on the nervous system of insects is unlikely to be highly toxic to plants. Specificity of action may not always be the case. For example, some biocides, such as the chlorophenols, are similarly toxic to a wide range of organisms (Liber et aI., 1994) and the grouping of all organisms together for distributional analysis may be appropriate. Thus, from a basic understanding of the mechanism of action of a pesticide and from the toxicity data, it may be possible to identify and group sensitive organisms that are the most likely to be adversely affected. This is helpful from the point of view of risk assessment because it allows the assessor to focus on the groups at higher direct risk and to devote less time and resources to groups exposed to very low or negligible direct risks. In addition, with a knowledge of the ecology of the potentially impacted system, it is possible to assess the likelihood that indirect effects will occur as a result of an effect on keystone groups of predator or prey/food organisms, should these be in the sensitive groups. Although the mechanism of action of the pesticide is an important criterion for grouping of organisms, habitat may also be important. For example, there may also be good mechanistic reasons to separate effects data for freshwater and saltwater organisms, where it is known that one group has an inherently different sensitivity because of interactions between salinity and the pesticide of concern (Hall and Anderson, 1995; Solomon et aI., 2001a; Solomon and Takacs, 2001). It is also possible to group organisms together on the basis of their reproductive strategy and life cycle. Thus, organisms that are able to recover rapidly from an adverse effect at the population level (reduction in population caused by mortality) may be considered differently from another group of organisms that may require a longer period of recovery. For example, aquatic algae have short reproductive cycles and would be expected to recover from a decrease in population more rapidly than a population of fish subjected to a similar reduction. Thus, the frequency of occurrence and the intensity of the effect that could be tolerated would be different. This is also important when deciding how the exposure data should be analyzed. Instead of estimating the likelihood that a specific toxicity criterion (say the 10th centile of the species sensitivity distribution) will be exceeded, exceedence probabilities can be presented as a continuum of likelihoods. This allows the risk assessor to judge the possible adverse outcomes over a range of possible combinations in an exceedence profile or, as it was named in the ECOFRAM reports, a joint probability curve (Fig. l3.6). These approaches are useful for communication of risks (ECOFRAM, 1999; Giesy et aI., 1999; Solomon and Takacs, 2001).
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13.3.3.3 Using Multiple Lines of Evidence in Ecotoxicological Risk Assessment of Pesticides
The results of ecological risk assessments need to be interpreted in the context of a number of lines of evidence, which include ecological, abiotic, and biotic, components of the ecosystem. Whereas the fact that the probabilistic approach is a purely numerical methodology is an advantage from the point of view of the transparency of the procedure, it cannot, nor is it designed to, assess the ecological relevance of the exceedences that may be identified. For example, an assessment criterion of the 10th centile may include keystone organisms of value to ecosystem function. Effects on keystone species would be expected to extend to other species that are dependent on them, for example, as a source of food or as a predator. For this reason, it is necessary to assess the role of the potentially affected species in terms of their function in the ecosystem and whether this can be taken over by other organisms (Hall and Giddings, 2000). The probabilistic approach can be used to refine the assessment process by allowing a rational ranking of scenarios by risk (likelihood of exceeding assessment criteria) and by identifying species in the distributions for which functional redundancy may exist (less sensitive organisms that can also perform the same function as the more sensitive and more affected organisms). Ecological relevance can most usefully be assessed from a basic knowledge of ecology and from tests, such as microcosms, where community resiliency, productivity, and function can be evaluated directly. For this reason, refinement of the effects characterization in a probabilistic risk assessment gives a greater reduction of uncertainty. The temporal and spatial scale of pesticide exposure is important in ecological risk assessment. The return frequency of an event (how often the event happens) is an important consideration in the choice of methods for probabilistic risk assessment and is related to the ecological cost of recovery from the event (Solomon, 1996). In assessing exposure, the return frequency protected should be consistent with the resiliency of vulnerable populations. Resiliency is determined by life-cycle characteristics and reproductive capacity of the potentially affected organisms and the ability of their populations (or their function in the ecosystem) to recover from the episode. The Report of the Aquatic Risk Assessment and Mitigation Dialogue Group (SETAC, 1994) recommended conservative approaches to ecological risk assessment, such as the use of low return frequencies, for example, one or fewer occurrences in 30 years. This safeguards all organisms in situations where limited information is available on mode of action or sensitivity of species. Where better information is available, more appropriate return frequencies may be used. For example, more frequent adverse events may be tolerated where a stressor affects organisms with short life cycles and high rates of reproduction. In temperate regions, many ecosystems undergo a period of dormancy and the system is, in a sense, reset seasonally by the winter. Thus, for some organisms, mechanisms for propagation beyond the winter reset already exist
370
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99
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..
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Figure 13.6 Presentation of exceedence probabilities (A) as a continuum of likehoods in an exceedence profile (B) and the use of these curves in decision making (C and D). Adapted from Solomon and Takacs, 2001.
and resting and other dormant stages are produced from which populations in the next season will develop. Similar mechanisms exist in environments with a dry season where ephemeral water bodies are subjected to drying out. Therefore, as many organisms in these regions undergo seasonal resets, a stressor return that occurs less frequently than once per season is likely to be tolerable from the viewpoint of the long-term productivity of the population and the sustainability of function in the ecosystem, especially if the effects are spatially restricted. Protection of longer-lived species without seasonal resets, such as some fish, birds, or mammals, may, however, require the consideration of return frequencies of several years or more. If a stressor is present nonuniformly in the environment, unexposed areas \\' ill act as refugia (metapopulations) for repopulation of potentially impacted areas. The relative size of the exposed and unexposed areas and their closeness is important, but this issue is particularly significant for assessing risks from pesticide use, where untreated fields, set-aside land, conservation headlands, crop rotations, and mixed farming practices guarantee that refugia will be present. Similarly, refugia exist in streams and rivers, and many organisms have resistant stages or propagules from which population recovery can occur. Thus, probabilistic risk assessments (and hazard quotients) are additionally conservative because they do not consider repopulation from unexposed refugia. The example of the more rapid than expected recovery of the biota in the Rhine River from an endosulfan spill illustrates this point (Friege, 1986).
13.4 UNCERTAINTY IN RISK ASSESSMENTS Uncertainty analysis is important in ecotoxicological risk assessment because it both identifies and, to the extent possible, quantifies the uncertainty in the entire process of problem formulation, analysis, and risk assessment. In addition, an assessment of uncertainty may allow identification of ways in which uncertainty can be reduced. Uncertainties in risk assessment have three sources: 1. Ignorance or imperfect knowledge of things that should be known is the first source of error. An example of this is the lack of prior knowledge that DDT would biomagnify in the food chain. 2. Systematic errors in the risk assessment process are those that may occur through computational mistakes or through incorrect instrumental calibration. These errors can be addressed through better quality control and quality assurance. 3. Nonsystematic errors are random or stochastic errors that result from the random nature of the system being assessed. These types of errors can be described and quantified, but cannot be avoided or corrected for. To the extent that probabilistic risk assessment uses distributions of data, errors of this type are incorporated in the assessment process. There is a reluctance in many risk assessors to admit that uncertainty exists in any decision; however, all scientific data have some uncertainty, as do all risk decisions made by humans. A description of sources of uncertainty is, in fact, helpful
References
to the risk manager because this allows identification of mechanisms by which additional certainty can be added to the risk assessment with the appropriate increase in comfort level in the result of the process on the part of the public and other users of the information.
13.5 RISK COMMUNICATION Once risk has been assessed, it will almost always be necessary to develop a risk communication strategy. This strategy may be needed for communication between the assessor and the manager or between the manager and politicians and the public. Communicating risk is not an easy task, especially if the result of the assessment is contrary to conventional wisdom or to the interests of certain stakeholder groups. Written risk communications must be designed to be understood by the lay public and this is a skill that few have. Verbal communication is even more difficult, especially if it is carried out in front of an audience where nonverbal communication can give mixed signals to the audience. Part of the difficulty with communicating risk assessment is that, by human definition, risk is assumed to be adverse. In the final analysis, in the ecosystem there is neither "good" nor "bad," and certainly nothing "adverse." An ecological change is labeled "adverse" by individuals or society and is basically a value judgment. Thus, to conduct a risk assessment means that someone has made a value judgment of which conditions will be defined as adverse. The public is often suspicious of the motives of those who communicate risks and may perceive conflicts of interest (Lackey, 1995). Risk communication is a form of persuasive communication that is designed to change behavior or what may be deeply held beliefs. Two of the keys to successful risk communication are (1) expressing the technical evaluation in a way that is meaningful to the audience (this can be achieved by using appropriate analogies to describe the risk assessment process and not using technical terminology that can be misunderstood) and (2) anticipating potential misunderstandings and dealing with them in a sympathetic way. In doing this, the communicator must recognize that groups who oppose a particular risk management strategy have a right to question a decision that affects them.
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Hall, L. W. J., and Giddings, J. M. (2000). The need for multiple lines of evidence for predicting site-specific ecological effects. Human Eeo!. Risk Assess. 6, 679-710. Hall, L. W. J., Giddings, J. M., Solomon, K R., and Balcomb, R. (1999). An ecological risk assessment for the use of Irgarol 1051 as an algaecide for antifoulant paints. Crit. Rev. Toxieol. 29,367-437. Health Council of the Netherlands (1993). Ecotoxicological risk assessment and policy-making in the Netherlands-Dealing with uncertainties. Network 6/7,8-11. Hill, I. R, Heimbach, E, Leeuwangh, P., and Matthiessen, P., eds. (1994). "Freshwater Field Tests for Hazard Assessment of Chemicals." CRC Press, Boca Raton, FL. HC (1993). "A Strategy for the Virtual Elimination of Persistent Toxic Substances, Vo!. I. Report of the Virtual Elimination Task Force to the HC." International Joint Commission, Windsor, ON. James, E C. (1994). Society Actions. Joint EAS/AIBS review of President Clinton's plan for the management of forests in the Pacific Northwest. Bul!. Entomo!' Soc. Am. June, 69-75. Klaine, S. J., Cobb, G. P., Dickerson, R. L., Dixon, K R., Kendall, R J., Smith, E. E., and Solomon, K R (1996). An ecological risk assessment for the use of the biocide, dibromonitrilopropionamide (DBNPA) in industrial cooling systems. Environ. Toxieo!. Chem. 15,21-30. Kovacs, T. G., Martel, P. H., Voss, R H., Wrist, P. E., and Willes, R E (1993). Aquatic toxicity equivalency factors for chlorinated phenolic compounds present in pulp mill effluent. Environ. Toxieol. Chem. 12, 684-691. Lackey, R T. (1995). The future of ecological risk assessment. Human Eeo!. Risk Assess. 1, 339-343. Laudan, L. (1994). "The Book of Risks. Fascinating Facts About the Chances We Take Every Day." Wiley, New York. Lehtinen, K.-J., Axelsson, B., Kringstad, K, and Stromberg, L. (1991). Characterization of pulp mill effluents by the model ecosystem technique: SSVL investigations in the period 1982-1990. Nordie Pulp Paper Res. 1. 2,81-88. Liber, K, Kaushik, N. K, Solomon, K R., and Carey, J. H. (1992). Experimental designs for aquatic mesocosm studies: A comparison of the "ANOVA" and "regression" design for assessing the impact of tetrachlorophenol on zooplankton populations in limnocorals. Environ. Toxiea!. Chem. 11, 6177. Liber, K, Solomon, K R., Kaushik, N. K., and Carey, J. H. (1994). Impact of 2,3,4,6-tetrachlorophenol (DIATOX) on plankton communities in limnocorals. In "Aquatic Mesocosm Studies in Ecological Risk Assessment" (R. L. Graney, J. L. Kennedy, and J. H. Rogers, eds.), pp. 257-294. Lewis Publishers, Boca Raton, FL. Lynch, M. R, ed. (1995). "Procedures for Assessing the Environmental Fate and Ecotoxicology of Pesticides," pp. 1-54. SETAC Europe, Brussels. McBean, E. A., and Rovers, E A. (1992). Estimation of the probability of exceedences of a contaminant concentration. Ground Water MonU. Rev. 12, 115-119. Mullins, J. A, Carsel, R. E, Scarbrough, J. E., and Ivery, A M. (1993). "PRZM-2 A Model for Predicting Pesticide Fate in the Crop Root Zone and Unsaturated Soil Zones: Program and User's Manual for Release 2.0," Rep. 600/R-93/046. U.S. Environmental Protection Agency, Athens, GA NRC (1993). "Issues in Risk Assessment." National Academy Press, Washington,DC. OECD (1981). "OECD Guidelines for Testing of Chemicals." OECD, Paris. OECD (1984). "OECD Guidelines for Testing of Chemicals, Update of 1984." OECD, Paris. OECD (l992a). "GLP Consensus Document. Compliance of Laboratory Suppliers with GLP Principles. OECD Series on Principles of Good Laboratory Practice and Compliance Monitoring 5," Rep. 49. OECD, Paris. OECD (l992b). "GLP Consensus Document. Quality Assurance and GLP. OECD Series on Principles of Good Laboratory Practice and Compliance Monitoring 4," Rep. 48. OECD, Paris. OECD (I 992c). "GLP Consensus Document. The Application of GLP Principles to Field Studies. OECD Series on Principles of Good Laboratory Practice and Compliance Monitoring 6," Rep. 50. OECD, Paris.
OECD (l992d). "Guidance for GLP Monitoring Authorities Guides for Compliance Monitoring Procedures for Good Laboratory Practice 2," Rep. 46. OECD, Paris. OECD (l992e). "Guidance for GLP Monitoring Authorities. Guidance for the Conduct of Laboratory Inspections and Study Audits 3," Rep. 47. OECD, Paris. OECD (1992f). "Guidance for GLP Monitoring Authorities. Guidance for the Conduct of Laboratory Inspections and Study Audits," Rep. 47. OECD, Paris. OECD (1992g). "The OECD Principles of Good Laboratory Practice. OECD Series on Principles of Good Laboratory Practice and Compliance Monitoring 1," Rep. 45. OECD, Paris. OECD (l993a). "GLP Consensus Document. The Application of the GLP Principles to Short-Term Studies. OECD Series on Principles of Good Laboratory Practice and Compliance Monitoring 7," Rep. 73. OECD, Paris. OECD (l993b). "GLP Consensus Document. The Role and Responsibilities of the Study Director in GLP Studies. OECD Series on Principles of Good Laboratory Practice and Compliance Monitoring 8," Rep. 74. OECD, Paris. Okkerman, P. C., van der Plassche, E. J., Emans, H. J. B., and Canton, J. H. (1993). Validation of some extrapolation methods with toxicity data derived from multiple species experiments. Eeataxiea!' Environ. Safety 25, 341359. OMEE (1990). "The Ontario Ministry of the Environment Scoring System: A Scoring System for Assessing Environmental Contaminants," Hazardous Contaminants Branch, Ontario Ministry of the Environment, Toronto. Parker, R. (1999). "GENEEC." Environmental Fate and Effects Division, Office of Pesticide Programs, U.S. EPA, Washington, DC. Parkhurst, B. R, Warren-Hicks, W., Cardwell, RD., Volison, J., Etchison, T., Butcher, J. B., and Covington, S. M. (1996). "Aquatic Ecological Risk Assessment: A Multi-Tiered Approach to Risk Assessment," Rep. 91-AER-1. Water Environment Research Foundation, Alexandria, VA. Parrott, J. L., Hodson, P. v., Servos, M. R., Huestis, S. L., and Dixon, D. G. (1995). Relative potencies of polychlorinated dibenzo-p-dioxins and dibenzofurans for inducing mixed function oxygenase activity in rainbow trout. Environ. Taxieol. Chem. 14,1041-1050. Reinert, K H., Bartell, S. M., and Biddinger, G. R, eds. (1998). "Ecological Risk Assessment Decision-Support System: A Conceptual Design," pp. 1-98. SETAC Press, Pensacola, FL. Riley, D., ed. (1993). "Principles of Risk Assessment," Rep. 73. The Winand Staring Centre for Integrated Land, Soil and Water Research, Wageningen, Netherlands. Schwarz, R c., Schults, D. W., Ozretich, R. W., Lamberson, J. 0., Cole, E A, DeWitt, T. H., Redmond, M. S., and Ferraro, S. P. (1995). Sigma PAH: A model to predict the toxicity of polynuclear aromatic hydrocarbon mixtures in field-collected sediments. Environ. Taxieol. Chem. 14, 1977-1978. SETAC (1991). "Report of the Wintergreen Workshop on Microcosms." SETAC Foundation for Education, Pensacola, FL. SETAC (1994). "Pesticide Risk and Mitigation. Final Report of the Aquatic Risk Assessment and Mitigation Dialog Group." SETAC Foundation for Environmental Education, Pensacola, FL. Solomon, K R. (1996). Overview of recent developments in ecotoxicological risk assessment. Risk Ana!. 16,627-633. Solomon, K R (1999). Integrating environmental fate and effects information: the keys to ecotoxicological risk assessment for pesticides. In "Pesticide Chemistry and Bioscience: The Food-Environment Challenge" (G. T. Brooks and T. R. Roberts, eds.), pp. 313-326. Royal Society of Chemistry, London. Solomon, K. R., Baker, D. B., Richards, P., Dixon, K R., Klaine, S. J., La Point, T. w., Kendall, R. J., Giddings, J. M., Giesy, J. P., Hall, L. W. J., Weisskopf, c., and Williams, M. (1996). Ecological risk assessment of atrazine in North American surface waters. Environ. Taxieol. Chem. 15, 3176. Solomon, K R, and Chappel, M. J. (1998). Triazine herbicides: Ecological risk assessment in surface waters. In "Triazine Risk Assessment" (L. Ballantine, J. McFarland and D. Hackett, eds.), Vo!. 683, pp. 357-368. American Chemical Society, Washington, DC.
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CHAPTER
14 Developmental and Reproductive Toxicology of Pesticides Poomi Iyer California Environmental Protection Agency
14.1 INTRODUCTION Pesticides have been in use since the early days of modern agriculture and, as a class of chemicals, are well studied and subject to much regulation. There is growing concern about the safety of pesticides and how exposure (occupational/via food residues/contamination of air and water) may affect human health in general and reproductive outcome in particular. Attention to pesticide residues in food (fresh and processed), spurred by a demand for organically [as defined by the Organic Foods Act of California (1990)] grown produce, and reports on the levels of pesticides in the diets of infants and children (NAS, 1993) have led to the passage of federal regulations in the United States (EPA, 1996a). Exposure to pesticides is inherent in most agriculture-related occupations, and studies on pesticide use and pregnancy outcome generally focus on birth defects and the effects on the reproductive system. Endpoints investigated cover a broad range, including early and late fetal loss, alteration in gestational age at delivery, formation of terata (birth defects), infant/child morbidity and mortality, male/female sexual dysfunction, sperm abnormalities, amenorrhea, dysmenorrhea, and illness during pregnancy and parturition. Pesticides acting on the developing organism or on the reproductive system may produce adverse effects by one of several mechanisms. They may be direct-acting by being chemically reactive and (1) cause germ cell destruction (e.g., alkylating agents) or (2) exert their effects due to their structural similarity to endogenous molecules (e.g., hormone agonists/antagonists such as phytoestrogens). They could also act indirectly and interrupt reproduction (1) by metabolism to a direct-acting compound or reactive intermediate, (2) via endocrine alterations such as increased/decreased steroid clearance, and (3) by stimulating or inhibiting neuroendocrine responses at the level of the hypothalamus or pituitary. Developmental toxicants, through a direct- or indirect-acting mechanism, may result in embryolethality, frank malformations, or other undesirable sequelae such as growth retardation or functional alteration. Similarly, pesticides affecting reproHandbook of Pesticide Toxicology
Volume I. Principles
duction may act on selected stages targeting the prenatal stage, the prepubertal stage, or the adult, resulting in damage to the reproductive organs and/or impaired fertility. The potential of pesticides to adversely affect development is determined from studies conducted on animals to meet the regulations of the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA; EPA, 1982). Although this chapter is not intended to be encyclopedic in nature, the reproductive and developmental toxicity of a number of compounds regarded as pesticides will be discussed.
14.1.1 DEVELOPMENTAL TOXICITY Adverse effects on the conceptus, covering the period from conception to the completion of morphological structure and functional capability of the individual, are included in this category. After the malformations caused by the drug thalidomide in the early 1960s, the role of chemical exposure in the causation of birth defects has received much attention. Accordingly, following the pattern for drugs developed by the Food and Drug Administration (FDA) , animal studies using pesticides are conducted and findings from such studies are extrapolated to humans. In addition to malformations in fetuses, endpoints such as prenatal death, growth alterations, developmental variations, and maternal effects as well as postnatal development are also investigated. Postnatal neurodevelopment in animals and the field of behavioral changes are now gaining attention especially as alterations in these areas have been reported for several pesticides (Boyes et aI., 1997; Chernoff et aI., 1979a; Gray et aI., 1986; Ostby et aI., 1985; Sette et aI., 1989; Tilson et aI., 1988). Increased perinatal mortality has been reported in laboratory animals from excessive exposure to pesticides; in one case, the deaths may have been related to functional cardiac disorders (Grabowski and Daston, 1983). Also, postnatal exposure via lactation has resulted in the induction of cataracts in rat pups (Chernoff et aI., 1979b; Gaines and Kimbrough, 1970). Epidemiological studies have documented an association between spontaneous abortions and fetal deaths and maternal exposure (Barlow and Sullivan, 1982; Goulet and The-
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CHAPTER 14
Developmental and Reproductive Toxicology of Pesticides
riault, 1991; Weinberg, 1993; Wi1son, 1979). The potential of pesticides to cause adverse effects on the developing individual has been demonstrated in laboratory animal studies with a range of effects at various stages of development. Of the numerous pesticides tested so far, about 43% have been documented to induce birth defects in experimental animals (J. L. Schardein, personal communication). Although this could be alarming, it must be remembered that all chemicals can interfere with some aspect of development if administered at a sufficiently high dose level at the appropriate time of development to certain species of animals. Development is a complex process, and the effects of a chemical depend on the time of exposure, the exposure level, and the extent of maternal effects. The nature of an insult during embryonic development is often less important than the developmental stage at which it occurs. This is because the steps in the sequence of tissue interactions during development are susceptible to disruption for a specific period of time. Typically, early exposure (i.e., during preimplantation and early postimplantation) results in fetal death, whereas exposure during the organogenesis period (3 weeks after conception through 2 months in humans) results in structural birth defects. However, there is evidence that exposure during the preimplantion period can also result in teratogenic effects (Rutledge et aI., 1992; Spielmann and Vogel, 1989). Pre- or postimplantation exposure of the developing conceptus to toxicants may also result in a "derailment" in the genetic control of development and the coordinated cascade of events that occur during normal development. Thus, developmental abnormalities may be induced by disrupting the coordinated expression of developmental genes involved in genomic imprinting, cell lineage specification, cell mixing and recognition, cell-cell interaction, cell migration and differentiation, and segmentation, depending on the time of exposure (Kimmel et aI., 1993). Exposure after the critical stage of organogenesis often results in growth retardation or other functional deficits. For regulatory purposes, hazard identification is based on the dose level at which an effect is noted, the observation of a dose-response relationship, and whether the adverse effect on the conceptus occurs at an exposure level below that which causes severe maternal toxicity. This is done partly to determine if the maternal effects are the underlying mechanism for the developmental effects noted. Testing for developmental toxicity therefore requires high doses even though humans may be exposed to low doses in practice. The details on testing protocols will be elaborated later on, suffice it to say at this point that the purpose for testing at high doses is to get an understanding of the mechanism of action of the chemical. The limitations of such testing, as in other toxicology studies, include the range of sensitivity within humans, extrapolation of effects observed at high doses to predict those likely to occur at low doses, as well as extrapolation from tests in animals to humans.
14.1.2 REPRODUCTIVE TOXICITY This category covers the adverse effects on the reproductive system, ranging from small decrements in reproductive abil-
ity in either male or female to a situation of overall infertility. It also includes effects on the reproductive organs irrespective of the influence on fertility in the affected individual. This is particularly relevant when animals are used as models for effects in humans because fertility in rodents is often difficult to disrupt and other indicators of reproductive function may be more sensitive. Hence, fertility cannot be used as the only tool to diagnose adverse effects. The effects of pesticides on reproduction may be acute or chronic and may be directed to a single sex. Also, pesticides affecting reproduction may act on selected stages targeting the prenatal stage, the prepubertal stage, or the adult. A few examples of reproductive toxicity by one or more of these mechanisms will be reviewed herein. Furthermore, due to unique anatomical and physiological characteristics in different species, the effects noted may differ. A pesticide causing reproductive toxic effects in one species may not be toxic in another and hence the relevance to humans needs to be examined. The impact on fertility also needs to be considered because the level of sperm production differs across species and a decrease in sperm count may not have the same impact in all species.
14.1.3 EPIDEMIOLOGY The degree to which pesticide exposure mayor may not be responsible for developmental problems in humans is not known. Available epidemiological data on the developmental toxicity of occupational and environmental pesticide exposure are limited in the sense that although a number of studies have some indications of elevated risk, the epidemiological evidence on the whole is unclear. The incidence of pesticide-related adverse reproductive/developmental outcomes has been extensively reviewed (Sever et aI., 1997). Elevated risk of limb anomalies (Lin et aI., 1994; Schwartz and LoGerfo, 1988; Schwartz et aI., 1986) has been associated with ecological exposure and occupational exposure; and orofacial clefts (Nurminen, 1995) have been related to maternal environmental exposure. In several countries, it has been noted that maternal agricultural occupation and pesticide exposure may be associated with elevated risk of spontaneous abortion and stillbirth (Goulet and Theriault, 1991; Heidam, 1984; Restrepo et al., 1990a, b; Rita et aI., 1987). Nonetheless, although reports from accidental exposure as well as occupational use have documented that pesticides can be incriminated in adverse reproductive outcomes, some studies have found no indication of reproductive hazards, presenting rather inconclusive results (Nurminen, 1995). Exposure to organochlorine and organophosphate pesticides in grape gardens in India resulted in higher abortion rates (almost sixfold) in 12 exposed couples compared to 15 nonexposed couples (Rita et aI., 1987). The compounds handled in this study included dichlorodiphenyltrichloroethane (DDT), lindane, quinalphos, dithane M45, metasystox, parathion, copper sulfate, dichlorovos, and dieldrin. Similarly, women working in vineyards in the Crimea also had higher rates of miscarriage after exposure to DDT, sulfur, methyl parathion, and copper sulfate (Nikitina, 1974). In China, women exposed
14.2 Mechanisms of Action
377
to chlorophenamidine (chlordimeform), dikishuang, and kitazin were found to be at increased risk of delivering stillbirths [relative risk (RR = 1.4-1.81)] as well as spontaneous abortions (RR = 1.90-4.00, depending on gravidity). The risks would have been even higher if previous adverse pregnancy outcomes had not been controlled for in the study (Weinberg, 1993). In rural California, second-trimester occupational exposure to pesticides was associated with an odds ratio (OR) of 4.8 in a casecontrol study of stillbirths and early neonatal deaths (Pastore et aI., 1995). Male and female farmers exposed to pesticides in central Sudan had a higher odds ratio for stillbirth (>500 g) in a case-control study: OR = 5.1; 95% confidence interval (Cl) = 1.4-9.6 (Taha and Gray, 1993).
in their site of action and most of these compounds are cytotoxic, carcinogenic, mutagenic, or developmentally toxic. They may also be toxic to the reproductive system and, in fact, the disruption of reproductive function could occur at doses lower than those that cause tumors. The classic example of such a mechanism is the case where the risk of sterility following many forms of cancer chemotherapy is considerably higher than the risk of second tumors (Kay and Mattison, 1985). Other direct-acting compounds are structurally similar to endogenous molecules, such as some organochlorines that may exert their effects through interaction with estrogen receptors. Organochlorines have been implicated in abnormal menses and impaired fertility (Mattison et al., 1983).
14.1.4 EXPOSURE
14.2.1.2 Indirect-Acting
Human malformations occur in roughly 5% of live births; therefore, to demonstrate an increase in the overall rate of malformation or incidence of a specific type of malformation from a documented exposure, a much larger population is required than if the background rate were zero (Fraser, 1977). The limitations often associated with epidemiological data, such as recall bias, lack of specificity, and use of surrogates for exposure, must be considered in evaluating findings that suggest an association with pesticide exposure. One explanation for these results is the specificity of the compounds involved in the exposure. Given that exposure is often categorized as either general pesticide use or agricultural setting, exposure to specific compounds is not evaluated. However, the effects of known classes of chemicals can be studied because a number of compounds have a common mechanism of action as reviewed in the next section. The adverse effects of pharmaceutical agents have been predicted from data on laboratory animals at exposures near maternally toxic levels (John son et aI., 1990). Much of the animal data on the reproductive and developmental effects of pesticides are generated for the purpose of pesticide registration under FIFRA and do not appear in the open literature. Hence, the amount of published information is limited. This chapter will attempt to fill that void by addressing several aspects in the area of developmental and reproductive toxicity of pesticides.
14.2 MECHANISMS OF ACTION Compounds used as pesticides have different mechanisms of action and these may be independent of the species targeted. Pesticides can therefore be studied by their mechanism of action. 14.2.1 BASIC MECHANISMS OF EXPOSURE 14.2.1.1 Direct-Acting
Pesticides that are direct-acting may exert their effect by being chemically reactive; these compounds may be nonspecific
Developmental/reproductive toxicants that are metabolized to either chemically reactive products or structures similar to endogenous molecules fall into this group. The embryo and fetus as well as both the ovary and the testis have been demonstrated to have microsomal monooxygenases, epoxide hydrases, and transferases responsible for metabolizing xenobiotics (Dixon and Lee, 1980; Heinrichs and Juchau, 1980; Mattison and Thorgeirsson, 1978, 1979; Pedersen et aI., 1985). The basic mechanisms outlined previously, along with timing, influence the various developmental effects that are observed. The concept that insult prior to the beginning of "organogenesis" results only in an "all (i.e., death) or none" effect is no longer considered accurate. Abnormal development subsequent to insult at preimplantation stages suggestive of early alterations in pattern formation has been reported for retinoic acid (Rutledge et aI., 1994). We are continuing to learn more regarding developmental stage-related sensitivities and this area of pattern formation and early alterations is of concern in the area of pesticide exposure. 14.2.2 TIMING OF EXPOSURE
Just as the time of exposure determines the developmental effects of a chemical, toxicity to the reproductive system also varies with the timing of exposure. Accordingly, reproductive toxicants can be classified as follows. 14.2.2.1 Prenatal Reproductive Toxicants
These are compounds that affect the developing reproductive system in utero, resulting in prenatal ovarian or testicular toxicity in humans and animals. These include the absence of or a considerable decrease in the number of primordial oocytes (e.g., primary or secondary amenorrhea). Thus, although it is possible that prenatal exposure could affect the oocyte, current study protocols are not designed to detect subtle changes that may occur. More frequent testing for toxicity to male reproductive processes is conducted because of the premise of male sensitivity and the ease of access to gametes and gonads. Furthermore, it is often presumed that the female gamete is better
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protected from mutagenic chemicals due to the probability that chemically induced deoxyribonucleic acid (DNA) damage in a primary oocyte is repaired prior to ovulation (Preston et al., 1995). Despite the differences between males and females in terms of anatomy and biological control mechanisms for reproduction, in the absence of data to the contrary, it should be assumed that both male and female gametes are equally sensitive to reproductive toxicants.
exposure because the developing animal may be extremely sensitive to toxicants during sex differentiation, and a number of these effects are difficult to detect until late in life. Ultimately, the concern for reproductive vulnerability is noted in its impact on pesticide regulation. This might translate into the cancellation of registration for a given compound or the chemical may be placed on a list, such as California's Proposition 65, aimed at controlling human exposure to compounds capable of adverse effects.
14.2.2.2 Prepubertal Reproductive Toxicants
The modulation of the hypothalamic-pituitary-ovarian axis is influenced by higher centers in the central nervous system. Both ovulation and ovarian hormone production require the interaction of components of this axis and the effects of specific compounds on any of these levels exert their influence on reproduction. The effects can be elucidated clinically by examining the impact on menstrual cyclicity in humans and estrus cyclicity in non primate animals. Although there are no data to link the increasing trend of early menarche with pesticide use, the area of estrogenic effects and their role on the onset of puberty is receiving attention (Thigpen et al., 1999). In the rat, studies on dams consuming diets containing high concentrations of estrogenic substances, with resultant exposure of their pups in utero and prior to weaning, suggest that the estrogenimprinting metabolism of the pups or future responses to other exogenous estrogenic substances may be altered (Lamartiniere et al., 1995). Thus, the effects of exposure may be noted in subsequent generations because a number of pesticides may have estrogenic potential. The prepubertal gonad may differ from the sexually mature gonad in its sensitivity to the toxic impact of pesticides and this is an endpoint that deserves examination. Contaminants of pesticides, such as 2,3,7,8-tetrachlorodibenzop-dioxin (TCDD), may, in fact, have such an effect, but the findings are not conclusive. There is increasing attention to the latent effects of pesticides on sexual differentiation in rodents. 14.2.2.3 Adult Reproductive Toxicants
The effects on the reproductive system may be observed in adults as well as in their progeny if exposure occurs over a long period of time. These are generally detected in the multi generation studies conducted in laboratory animals. Studies submitted for regulatory purposes (fenthion, oxydemeton methyl) have demonstrated effects such as increased epididymal vacuolation and other histopathological changes (California Department of Pesticide Regulation, 2000). Gender differences in response to chemical insult must be taken into consideration; for example, unlike the male where damage to the spermatocytes may only be a temporary problem, damage to oocytes is permanent. The new guidelines now require histopathology data on ovaries to detect changes that may be occurring due to prolonged pesticide exposure over various developmental periods. Additionally, recent studies have documented adult/pubertal alterations resulting from gestational and or neonatal exposures (Gray and Ke1ce, 1996). Hence, studies should include a comprehensive assessment of reproductive function after perinatal
14.3 REGULATORY ISSUES Outlined in this section are the many issues dealing with the use of pesticides and their regulation by state, federal, and international agencies. These include the conduct and interpretation of studies as well as the application of new findings and regulations. 14.3.1 HISTORY
The effects of thalidomide and the Kefauver-Harris Act in 1962 led the U.S. Food and Drug Administration (FDA) to strengthen drug testing. Currently, the U.S. government requires manufacturers to perform hazard assessments to determine the teratogenic potential of chemicals. The U.S. Environmental Protection Agency (EPA) published teratogenicity testing requirements in 1978 under FIFRA. Essentially, it mandated how testing was to be conducted and reported, which differed little from the FDA guidelines, except that exposure was to be initiated just before implantation and concluded the day before delivery (EPA, 1978). [Whereas the FDA requires three studies covering different segments of development, the EPA requires (1) a standard teratogenicity study with exposure during the main period of organogenesis and (2) a two-generation reproduction study.] In the early 1980s, the EPA specified the kinds of data and information required under FIFRA to support the registration of pesticides (EPA, 1982), reflecting guidelines proposed in 1978 (EPA, 1978, 1984). Similar regulations also went into effect through the EPA for chemicals under the Toxic Substances Control Act (TSCA); these were revised in 1985 (EPA, 1985). In 1986, the EPA published procedures to evaluate potential developmental toxicity associated with human exposure to environmental toxic ants (EPA, 1986). Also, a screening test for developmental neurotoxicity to include behavioral and neuropathology analyses was proposed (EPA, 1986; Francis, 1987) and finalized into test rules in 1988 and 1989 (EPA, 1988, 1989). Postnatal functional assessment has been recognized as an important part of developmental toxicity testing in the United States and is required in some cases. In other countries, requirements are in place for behavioral testing as a part of developmental toxicity testing (Barlow, 1985; EEC, 1983; Tanimura, 1985; WHO, 1986). In November 1986, voters in the state of California approved an initiative to address concerns about exposure to toxic chemicals. That initiative became the Safe Drinking Water and Toxic
14.3 Regulatory Issues Enforcement Act of 1986, better known as Proposition 65. This requires the governor to publish a list of chemicals that are known to the state to cause cancer, birth defects, or other reproductive harm. The chemicals that cause birth defects or other reproductive harm are called reproductive toxicants. The Proposition 65 list contains a wide range of chemicals, including dyes, solvents, pesticides, drugs, and food additives. If a pesticide is on the list, an employer must warn the employee if the exposure levels of the pesticide present a significant health risk; the employer may also choose to provide warning simply based on the presence of the chemical, even if the risk is not significant. In the case of worker exposure to pesticides, this warning is provided through the required hazard communication procedures, and, as an agricultural crop producer, the employer is also required to keep application-specific information on the pesticides used. A number of pesticides have been listed and subsequently withdrawn from registration for use in the state. A complete list of compounds may be accessed via the Internet at http://www.oehha.ca.gov/prop65/pdf/80400LSTA. pdf. Table 14.1 lists the currently registered active ingredients on the Proposition 65 list.
be detected in reproduction (two-generation) and developmental neurotoxicity studies. These guidelines provide information on the appropriate study design and methodology for the conduct of studies and may also be accessed via the EPA Web site at http//www.epa.gov/oppts_harmonizedl870_health_effects_test _guidelines/. Figure 14.1a and b describes the protocols for developmental and reproductive toxicity testing and notes the approximate dosing and breeding schedules. Rats: Day:
0
21 Rabbits: Day: 0
29
14.3.2 PRINCIPLES OF TESTING AND EVALUATION The Office of Prevention, Pesticides, and Toxic Substances (OPPTS) recently revised the 1982 Health Effects Test Guidelines (EPA, 1998). The OPPTS-harmonized guidelines have been developed for use in the testing of pesticides and toxic substances and the development of test data that must be submitted to the agency for review under federal regulations. The purpose of harmonizing these guidelines into a single set of OPPTS guidelines is to minimize variations among the testing procedures that must be performed to meet the data requirements of both the FIFRA (7 U.S.c. 136, et seq.), as amended by the Food Quality Protection Act (FQPA) (P.L. 104-170) and the Toxic Substances Control Act (TSCA) (15 u.s.c. 2601). The Organization of Economic Cooperation and Development (OECD) guidelines 414 and 416 and the OPPT guidelines under 40 CFR 798.4900 and 40 CFR 798.4700, OPP guidelines 833 and 83-4, provided the source material for developing these harmonized OPPTS test guidelines. Changes to the previous guidelines (EPA, 1984) were considered over a period of years and involved both industry and regulatory agencies and a period of public comment. An increase in the number of animals in the developmental toxicity study conducted in rabbits has been requested. The timing of exposure has also been extended from day 19 of gestation to the day before fetuses are examined. Endpoints previously not evaluated include the status of the ovaries, sperm/semen evaluation, and examination of vaginal smears for evaluation of estrous cycles and other effects resulting from endocrine disruption. In light of the data that the critical period for inducing abnormalities may extend to the postnatal period, for example, renal development (Couture, 1990), functional deficiencies and other postnatal effects are expected to
379
20 females per dose group (Evidence of sperm/plug in female or in bedding) Begin exposure (around implantation) generally on day 0 or 6 and continue until day before parturition C-section and examine fetuses (generally C-section is on gestation day 20 if mating is on gestation day 0) 20 females per dose group (Day of artificial insemination; natural mating can be used) Begin exposure (around implantation) generally gestation day 0 or 7 and continue until day before parturition C-section and examine fetuses
At a minimum, the test substance should be administered daily from around the time of implantation to the day before Cesarean section on the day prior to the expected day of parturition. Alternatively, if preliminary studies do not indicate a high potential for preimplantation loss, treatment may be extended to include the entire period of gestation, from fertilization to approximately 1 day prior to the expected day of parturition. It is preferred that the dams are exposed from the time of mating. The timing of implantation and expected delivery may vary with the strain. Figure 14.1a Protocol for developmental toxicity testing. Age of animals: (in weeks)
5-9
15-19
21-25 15-19
FO/Pl Start of study Exposure for 10 weeks in diet (or other route, based on the most likely human exposure scenario) Mating (conducted over 2 weeks or 3 estrous cycles; remating with a different partner is not undertaken) Gestation (approximately 3 weeks) Parturition -+ F1a Weaning (approximately 3 weeks) Growth (approximately 15 weeks) Mating Gestation (approximately 3 weeks) Parturition -+ F2a Weaning
In certain instances, such as poor reproductive performance in controls, or in the event of treatment-related alterations in litter size, the adults may be remated to produce an Fib or F2b litter. If production of a second litter is deemed necessary in either generation the dams should be remated approximately 1-2 weeks following weaning of the last F la or F2a litter. Figure 14.1b Approximate dosing and breeding schedule involved in a two-generation study of effects on the reproduction process in rats.
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CHAPTER 14
Developmental and Reproductive Toxicology of Pesticides Table 14.1 Pesticides Known to the State to Cause Birth Defects or Reproductive Harm
Compound
CAS number
Altretamine
645056
August 20, 1999
Amitraz
33089611
March 30, 1999
Benomyl
17804352
July I, 1991
Bromacillithium salt
53404196
May 18, 1999
Bromoxynil
1689845
October I, 1990
Bromoxynil octanoate
1689992
May 18, 1999
Carbon disulfide
75150
July I, 1989
Chinomethionat (Oxythioquinox)
2439012
November 6,1998
Chlordecone (Kepone)
143500
January I, 1989 May 14, 1999
Arsenic (inorganic oxides)
Date listed
May I, 1997
May I, 1997
Cadmium
Chlorsulfuron
64902723
Cyanazine
21725462
April I, 1990
Cycloate
1134232
March 19, 1999
Cyhexatin
13121705
January I, 1989
2,4-D butyric acid
94826
June 18, 1999
2,4-DP (dichloroprop)
120365
April 27, 1999
1,2-Dibromo-3-chloropropane (DBCP)
96128
February 27, 1987
p'-DDT
789026
May 15, 1998
p, p'-DDT
50293
May 15, 1998
Dichlorophene
97234
April 27, 1999
Diclofop methyl
51338273
March 5, 1999
Dicumarol
66762
October I, 1992
Dinocap
39300453
April I, 1990
Dinoseb
88857
January I, 1989
Endrin
72208
May 15, 1998
Epichlorohydrin
106898
September I, 1996
Ethyl dipropylthiocarbamate
759944
April 27, 1999
Ethylene dibromide
106934
May 15, 1998
0,
Ethylene oxide
75218
February 27, 1987
Ethylene thiourea
96457
January I, 1993
Etretinate
54350480
July I, 1987
Fenoxaprop ethyl
66441234
March 26, 1999
Fluvalinate
69409945
November 6, 1998
Heptachlor
76448
August 20, 1999
Hexachlorobenzene
118741
January I, 1989
Hydramethylnon
67485294
March 5, 1999
Hydroxyurea
127071
May I, 1997
Linuron
330552
March 19, 1999
Mebendazole
31431397
February 27, 1987
Lead
August 20, 1999 July I, 1990
Mercury and mercury compounds Metham sodium
137428
May 15, 1998
Methotrexate
59052
January I, 1989
Methotrexate sodium
15475566
April I, 1990
Methyl bromide as a structural fumigant
74839
January I, 1993 July I, 1987
Methyl mercury
March 30, 1999
Metiram
9006422
Myclobutanil
88671890
April 16, 1999
Nabam
142596
March 30, 1999
Nicotine
54115
April I, 1990 (continues)
14.3 Regulatory Issues
381
Table 14.1 (continued) Compound
CAS number
Date listed
Oxadiazon
19666309
May 15, 1998
Oxydemeton methyl
301122
November 6,1998
Oxymetholone
434071
May 1, 1997 January 1, 1991
Polychlorinated biphenyls (as a contaminant) Potassium dimethyldithiocarbamate
128030
March 30, 1999
Propargite
2312358
June 15, 1999 January 29, 1999
Pyrimethamine
58140
Resmethrin
10453868
November 6, 1998
Sodium dimethyldithiocarbamate
128041
March 30, 1999
Sodium fiuoroacetate
62748
November 6,1998
Streptomycin sulfate
3810740
January 1, 1991
2,3,7,8-Tetrachlorodibenzo-para-dioxin
1746016
April 1, 1991
Thiophanate methyl
23564058
May 18, 1999
Triadimefon
43121433
March 30, 1999
Tributyltin methacrylate
2155706
December 1, 1999
Triforine
26644462,37273840
June 18, 1999
Vinclozolin
50471448
May 15, 1998
Warfarin
81812
July 1, 1987
(TCDD) (as a contaminant)
Updated November 26,1999.
14.3.3 CHOICE OF SPECIES IN TESTING The laboratory species typically used to test for developmental toxicity or for reproductive effects is the rat. Some strains are considered less suitable than others, and the rationale for the strain may vary with the compound and the effects it may cause in the species tested. The rabbit is the other species that is used as it is the one species (unlike the rat) that showed some signs for the compound thalidomide, the chemical that appeared safe in all other species tested. Additionally, because of the species specificity of teratogenic agents, the exact effects noted in laboratory animals are not necessarily those observed in humans. However, all proven human teratogens have parallel, but imperfect animal models. Determining which species is the most appropriate for extrapolation to humans for a given compound is difficult. Pesticides that involve food use are likely to have a higher potential exposure and are to be tested in two species as per FIFRA regulations. Among the species used for testing, the rat and mouse most successfully model the human reaction, but the rabbit is less likely than other species to give a falsepositive finding (EPA, 1991). The concomitant use of the rabbit with either the mouse or the rat is believed to enhance the predictive potential of the individual animal model (Schardein and Keller, 1989). Accordingly, the rat and rabbit are commonly used. Although no single species has clearly distinguished itself as being more advantageous in the detection of human teratogens over any other, it is concluded that safety decisions should be based on all reproductive and developmental toxicity data
in light of the agent's known pharmacokinetic, metabolic, and toxicologic profile.
14.3.4 CHOICE OF TESTING DOSES While reviewing studies submitted for regulatory purposes, a number of factors are taken into consideration. A compound may be embryolethal without being teratogenic; alternatively, it can also be both embryolethal and teratogenic. The teratogenic dose range, that is, the margin between the dose which will kill the fetus and that which does not have adverse effects on the fetus, is often very narrow. The wider the margin, the more potentially dangerous is the compound from a teratological perspective. Hence, it is recommended that the highest dose should only cause slight toxic effects on the pregnant animal (e.g., decreased body weight gain) such that a majority of the pregnancies reach term (Wilson, 1979). Choice of dosing regimens is critical to determining the potential of the compound to exert adverse effects. The mid-dose must not be much lower than the high dose as such a bracketing will result in a low no observed effect level (NOEL) and not provide data on the true nature of the chemical being evaluated. Although a low NOEL may appear to be more health protective than a higher NOEL, it is possible to miss the effects that the compound can cause at levels below the maximum tolerated dose. The choice of dose levels is critical to study design. Studies submitted with inappropriate doses are often unacceptable to regulatory agencies
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and the registrant has to either conduct a new study or provide justification to support the choice of doses employed. 14.3.5 INTERPRETING EFFECTS
Death of the conceptus may preclude expression of other major manifestations of developmental toxicity (i.e., structural abnormalities, altered growth, and functional deficit). Generally, the term teratogenicity refers to the observation of malformations, that is, permanent structural changes that may adversely affect survival, development, or function. Other developmental effects include variations, a term used to describe changes in the fetus that involve a divergence beyond the usual range of structural constitution that may not adversely affect survival or health. It is sometimes difficult to distinguish between malformations and variations because the responses constitute a continuum from normal to the extremely deviant. Other terms used are anomalies, deformations, and aberrations; however, they are not defined any better (EPA, 1991). To further confuse this already complex issue, these other terms may be used for either of these two categories, requiring a closer examination of the interpretation. Evaluating variations and interpreting their incidence is illustrated in the case of supernumerary ribs (SNRs). These are a common variant in some strains of mice used in standard teratology bioassays, and the increased incidence of SNRs may be induced by a wide variety of xenobiotics and/or general maternal stress. The significance of this defect in cross-species extrapolations has been problematic. In one study in mice, it was demonstrated that SNRs have a bimodal distribution composed of "rudimentary ribs" (RRs) with a mode of 0.3-0.4 mm and "extra ribs" (ERs) with a mode of 0.9-1.1 mm. The ERs and RRs were found to be morphologically distinct; the ERs were flat ended and distally joined by a cartilaginous portion, whereas the RRs were usually rounded distally and were without cartilaginous extensions. The 13th ribs were significantly longer in fetuses having SNRs than in those not having SNRs, whether treated or untreated. This relationship was present in all fetal ages examined and with both ER and RR groups, suggesting that SNRs are indicative of basic alterations in the development of the axial skeleton (Branch et aI., 1996). In the case of developmental toxicity, studies are reviewed taking into account the maternal effects observed. Developmental toxic effects in the presence of severe maternal toxicity are considered less severe than those observed in the absence of maternal effects. Generally, in order to determine whether or not the conceptus is uniquely susceptible, the developmental and maternal NOEL values are compared. The AID ratio (adult NOEL:developmental NOEL) has been advocated previously as an index of comparative teratogenic hazard (John son, 1981) and has been used to characterize the developmental effects of chemicals. The strategy of carefully characterizing the observed maternal toxicity at the individual level is also employed. To determine if the malformations observed were the result of maternal toxicity, two approaches may be adopted: consideration
of individual versus group mean data and examination of data during the specific period of gestation when the developmental malformations were likely to have occurred. Furthermore, because mere correlation of maternal toxicity with fetal effects does not imply causality (Chernoff et aI., 1987), maternal influence may not necessarily be the underlying mechanism of action. The severity of the effect on the fetus also needs to be considered; that is, the effect may be severe/life-threatening, whereas the maternal effects such as slight weight loss are minor or transient. Another confounding factor is that maternal effects may be reversible, whereas effects on the developing fetus may be permanent, underscoring the importance of characterization of the maternal effects. Examining the data in this manner leads to a more exacting interpretation of teratogenic potential. The use of a weight-of-evidence approach may help in verifying the apparent maternal toxicity at the individual level and in determining the developmental toxic potential of a chemical (lyer et aI., 1999). A weight-of-evidence approach should include the following: • Dose-response relationship • Supportive evidence in another species or related compounds • Closer scrutiny, focusing on individual data during the discrete time period(s) in which particular fetal malformations were most likely to have occurred In conducting the risk assessment, the developmental toxicity study serves as a surrogate for an acute toxicity study, based on the premise that the effects noted may be the result of a single exposure or an exposure over a short period. However, if the effects responsible for the NOAEL (no observed effect level)/RfD (reference dose) are known to result from multiple exposures, then use of the developmental toxicity study for acute effects would be inappropriate. Similar approaches are recommended for reproduction (twogeneration) and developmental neurotoxicity studies. Effects on lactation, acceptance of offspring, and sexual maturation are examined in the reproduction studies. Multigeneration studies are evaluated to determine if the effect noted is exacerbated in subsequent generations. Furthermore, spontaneous occurrence in control animals of stillborn pups and other developmental effects necessitate that evaluation of the data be subjected to rigorous statistical procedures. 14.3.6 STATISTICAL EVALUATION
One of the most important aspects of developmental toxicity analysis is that the litter is to be considered the experimental unit (EPA, 1991). Since it is the maternal unit that is exposed to the compound, the effects of the test substance on each fetus in a litter are related to the status of the animal bearing that fetus. Individual differences in maternal susceptibility can affect an entire litter, whereas others in the same dose group are unaffected. Hence, all fetuses in a single dose group are not
14.4 Toxicology Studies
equally at risk to the potential developmental effects of the test substance. Therefore, the accepted practice is to consider the litter as the experimental unit for developmental toxicity studies (Collins et aI., 1999; EPA, 1991; Gad and Weil, 1986; Gaylor, 1978). Evaluating the overall fetal effects helps further characterize the developmental toxic effects, but the litter is the preferred unit to evaluate the effects of the compound. Using model-fitting techniques and employing benchmark doses are encouraged if they are appropriate. Concurrent controls are the group of choice for comparison. Historical controls are recommended to serve as supportive evidence. 14.3.7 EXPOSURE ASSESSMENT
In evaluating the exposure for a given chemical, regulatory agencies take into consideration the amount used and the usage pattern (seasonality, etc.) and attempt to obtain the dose that reaches either the parent's germ cells or the developing conceptus. Although developmental toxicity is usually thought to be associated with maternal/embryonic exposure, there is increasing evidence for developmental effects due to male exposure (Colie, 1991; Sever, 1995). Agents associated with spontaneous abortions may also cause congenital malformations with the appropriate timing and dose; hence, the exposure pattern may determine the continuum of effects that might result. Additionally, it is thought that the steady accumulation of pesticides in the adipose tissues during a woman's lifetime may pose a risk, especially in the case of endocrine disrupting chemicals (Garcia-Rodriguez et aI., 1996). Issues associated with the importance of the timing and assessment of exposure during pregnancy have been discussed extensively (Hertz-Picciotto et aI., 1996). Knowledge of the time window of vulnerability has important implications for the assessment of risks. Along with the active ingredient, organic solvents are used extensively in pesticide formulations, and hence mixers and loaders of pesticides may be exposed to higher levels of both the active ingredient and the solvents/inerts. Effects caused by solvents could confound the issue and may impact exposure in an adverse manner. 14.3.8 IMPACT OF THE FOOD QUALITY PROTECTION ACT ON
DEVELOPMENTAL AND REPRODUCTIVE TOXIC EFFECTS OF PESTICIDES
In 1996, Congress passed landmark pesticide food safety legislation supported by the administration and a broad coalition of environmental, public health, agricultural, and industry groups. The bill was signed by the president on August 3, 1996, and the Food Quality Protection Act of 1996 became law (P.L. 104-170, formerly known as H.R. 1627). The EPA regulates pesticides under two major federal statutes. Under the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA), the EPA registers pesticides for use in the United States and prescribes labeling and other regulatory requirements to prevent
383
unreasonable adverse effects on health or the environment. Under the Federal Food, Drug, and Cosmetic Act (FFDCA), the EPA establishes tolerances (maximum legally permissible levels) for pesticide residues in food. The Food Quality Protection Act (FQPA) amendments to the FFDCA direct the EPA to consider a number of factors in making risk assessments as part of the tolerance-setting procedure. Most of these provisions originated in recommendations from the National Academy of Sciences (NAS) 1993 report "Pesticides in the Diets of Infants and Children" and reflect concerns that children may be especially susceptible to pesticide exposure. Specifically, in setting a tolerance for pesticide residues in food, the FQPA directs the EPA to consider the following: use of an extra lO-fold safety factor to account for the susceptibility of children; the special susceptibility of children, including effects of in utero exposure; cumulative effects of exposure to the pesticide and substances having a common mode of action; aggregate exposure for all consumers (i.e., other routes, such as drinking water and home and garden applications); and potential for endocrine disrupting effects. Incorporating these factors into the tolerance-setting process poses significant challenges to the agency because there are many scientific uncertainties surrounding the use of these factors in risk assessment. To this end, specific areas are brought to the Scientific Advisory Panel (SAP) for review. Current practice is generally to use a WO-fold safety factor when the toxicity data are from animal studies and to apply extra factors of 3to lO-fold only when specific case-by-case evidence seems to warrant it. A presentation to the SAP, a position paper on the subject has been released (EPA, 1996b). The 1996 law represents a breakthrough, amending both major pesticide laws to establish a more consistent, protective regulatory scheme. It mandates a single, health-based standard for all pesticides in all foods; provides special protections for infants and children; expedites approval of safer pesticides; creates incentives for the development and maintenance of effective crop protection tools for U.S. farmers; and requires periodic reevaluation of pesticide registrations and tolerances to ensure that the scientific data supporting pesticide registrations will remain up to date in the future. Additional information may be accessed via the Internet at www.epa.gov/oppfeadl/fqpa/sciissue.htm.Awider perspective on the appropriateness of extra safety factors to protect children based on the use of one pesticide (chlorpyrifos) can be gleaned from other publications on the subject (Gibson et aI., 1999; Schardein and Scialli, 1999) including the risk characterization document for chlorpyrifos (California Department of Pesticide Regulation, 2000).
14.4 TOXICOLOGY STUDIES Pesticides can be broadly classified into different categories, based on the target of use. The major classes that will be discussed include herbicides, insecticides (including insect growth regulators), fungicides, and rodenticides. In addition, compounds that have pesticidal action (antiparasitic) are used as
384
CHAPTER 14 Developmental and Reproductive Toxicology of Pesticides
animal health products and will be discussed under a miscellaneous category. 14.4.1 HERBICIDES Included in this class of chemicals are the chlorphenoxy compounds, bipyridyls, dinitrophenols, triazines, substituted ureas, some of the carbamates, plant growth inhibitors, and amides. A number of herbicides have been studied for developmentally toxic effects and adverse effects on reproduction in animals (see Table 14.2). Chlorphenoxy Compounds Of the various herbicides used, a plethora of data is available on the phenoxy defoliants 2,4-dichlorophenoxyacetic acid (2,4-D) and 2,4,5-trichlorophenoxy ace acid (2,4,5-T) which have been used worldwide in forestry and agriculture. The phenoxy herbicide, Agent Orange, composed of equal parts of 2,4-D and 2,4,5-T, has received a lot of attention after its large-scale use in Vietnam by the U.S. military during the war years of 1962-1971. The contamination of these compounds with the dioxin TCDD in commercial preparations has further confounded the effects noted. TCDD is found to be extremely teratogenic in laboratory animals: It has a very low teratogenic minimal effective dose of 1-10 !l-g/kg in susceptible strains of mice (Peam, 1985), and effects such as hydronephrosis and cleft palate have been noted (Couture et aI., 1990). Results from Agent Orange exposure are largely inconclusive and more details, including the litigation history may be obtained in previous reports (Schardein, 1993; Schuck, 1986). The phenoxy herbicide 2,4,5-T has not been manufactured since 1983. 2,4-D One case of cephalic malformations and severe mental retardation was noted in an infant whose parents received prolonged exposure via the dermal route from forest spraying (Casey and Collie, 1984). Though no birth defects were found, an increase in spontaneous abortions and premature births was noted in a case-control study examining the effects of 2,4-D (Carmelli et aI., 1981). Possible adverse effects such as gestational and neonatallosses with NOELs of 20 mg/kg/day (rats) and 500 ppm (dogs) were noted in two studies (California Department of Pesticide Regulation, 2000). Markedly reduced gestational and neonatal survival was not accompanied by a commensurate degree of maternal toxicity. In the Ontario Farm Family Health Study exposure to phenoxy herbicides during the first trimester was generally not associated with increased risk of spontaneous abortion (Arbuckle et aI., 1999). However, the results suggest a possible role of preconception (possibly paternal) exposures to phenoxy herbicides in the risk of early spontaneous abortions. Of other phenoxy herbicides that have been studied, 4-chloro-2-methylphenoxyacetic acid ethyl ester caused (31 % incidence) cleft palate and anomalies of the heart and kidney in rats (Schardein, 1993). 4-chloro-o-toloxy acetic acid was also found to be teratogenic in mice and rats at high oral doses (Roll and Matthiaschk, 1983; Schardein, 1993).
Amitrole This nonselective postemergence herbicide is also an antithyroid agent. Although structural malformations were not noted when tested in animal studies, fetal thyroid lesions were observed in rats exposed to amitrole via drinking water (Schardein, 1993). Bromoxynil The developmental toxicity of the wide-spectrum herbicide bromoxynil (bromoxynil phenol; 3,5-dibromo-4-hydroxyphenyl cyanide) and its octanoate ester (2,6-dibromo-4-cyanophenyl octanoate) was evaluated in Sprague-Dawley rats and Swiss-Webster mice. The highest doses of both compounds increased the incidence of supernumerary ribs (SNRs) in the fetuses of treated rats, but did not induce other anomalies (Rogers et aI., 1991). In the teratogenicity study submitted to the California Department of Pesticide Regulation (CDPR), rats were dermally exposed to Buctril (containing 33.8% bromoxynil octanoate) diluted with water. Based on a dose-dependent increase in the incidence of extra thoracic ribs in the fetuses, at the 15 mg/kg/day level and above, the developmental NOEL was determined to be 10 mg/kg/day. Dinoseb Dinoseb (2-sec-butyl-4,6-dinitrophenol) has been shown to produce substantial spermatotoxicity after 1-5 doses in short-duration tests (Linder et aI., 1992). In mice at 17.7 mg/kg/day subcutaneous or intraperitoneal administration of dinoseb during organogenesis resulted in skeletal defects, cleft palate, hydrocephalus, and adrenal agenesis. Maternal toxicity, however, was noted at doses between 17.7-20 mg/kg/day (Gibson, 1973). Dinoseb has also been reported to produce a high incidence of dilated renal pelvis in the term rat fetus (McCormack et aI., 1980) as well as SNRs in mice (Kavlock et aI., 1985). Teratogenic effects such as an increased incidence of microphthalmia were also reported in the rat fed dinoseb in the diet (Giavini et aI., 1986). Eye defects and neural malformations were noted in the rabbit, leading to dinoseb's being banned by the EPA in 1986. Bipyridyl Compounds (Paraquat, Diquat) The herbicide paraquat has resulted in at least eight fetal deaths when taken during pregnancy as a result of maternal poisoning (Talbot et aI., 1988). However, no adverse effects were reported in animal developmental and reproductive toxicity studies submitted to the CDPR. On the other hand, for the herbicide diquat, adverse systemic effects were noted in the rat in parents and offspring (cataracts and eye pathologies in both sexes of Fo and Fl at more than 240 ppm; an increase of hypertrophy and hyperplasia of collecting duct epithelium and tubular dilatation in the renal papilla in both sexes of Fl at 240 ppm; Fl and F2 pups showed hydronephrosis at and above 240 ppm). In data submitted to the CDPR for the rat, the developmental NOEL and NOAEL were equal to 12 mg/kg/day with intrauterine growth retardation, as measured by decreased weight and delayed skeletal ossification, and hemorrhagic kidneys as the main effects observed (CD PR, 2000). Mice appear to be more sensitive to diquat than were rats. The NOEL in the mouse
14.4 Toxicology Studies
385
Table 14.2 Developmental and Reproductive Toxicity Profile of Herbicidesa
Dose Chemical
Species
Toxicity profile
Acrolein
Rat
Reproduction
(mg/kg)b
3:pup l:parent
Alachlor (ethane
Rat
Comments
References
Decreased body
CDPR Database, 2000
weight in pups Heydens et ai., 1996
1000
sulfonate) Ametryn
Rat
Increased skeletal
50
Infuma et ai., 1987
variants
As cited in Schardein, 1993
Reproduction Rabbit Amitraz Amitrole
Rat
60 Reproduction
Mouse
Altered estrous cycles
Rat
Thyroid effects
Rabbit
Abortions, reduced weight gain
Infuma et aI., 1987
10.5 ppm
Pup mortality
CDPR Database, 2000
0.0004%
Via drinking
As cited in Schardein, 1993
DPN #287,1994 water (3-G study) 4
DPN#0216
CDPR Database, 2000 (worksheet 033 45711)
Arsenic
Teratogenic: malformations
Arsenic acid
Mouse
Sodium arsenate
RatIMouse
Asu1am
Rat
CDPR Database, 2000 7.5
DPN #180 Parenterally (ip)
Reproduction: decreased
1000 ppm
CDPR Worksheet (360) 010
number of live births Atrazine
Rat
Fetal toxicity
As cited in Schardein, 1993 25257; 19
70
2-G study
Disruption of ovarian
Infuma et ai., 1987 Cooper et al., 1996
cycle, induced pseudopregnancy Reproduction Balagrin
Rabbit
Fetal toxicity
75
Infuma et ai., 1986
Mouse
Teratogenic
22
As cited in Schardein, 1993
Rat
Teratogenic
1/50 LDsO 75
Mirkova, 1980 C1opyra1id
Rat
Delayed ossification Reproduction
Rabbit
Selectively toxic
Hayes et aI., 1984
(2-G study)
Dietz et aI., 1986
250
Cyanazine
Rat
Developmentally toxic
2,4-D
Mouse
Developmentally toxic
Hayes et ai., 1984 Anophthalmia!
Lu et ai., 1982
microphthalmia 221
As cited in Schardein, 1993
Teratogenic Rat
Fetal death
50
Teratogenic
2-G study
Reproduction 2,4-D + picloram
Hamster
Teratogenic
Mouse
Developmentally toxic
20 0.2%
Teratogenic
Drinking water
Blakley et ai., 1989
route
2,4-D + 2,4,5-T
Rat
Behavioral effects
2,4-D butoxyethanol
Rat
Teratogenic
150
Rat
Teratogenic
150
As cited in Schardein, 1993
Sheep
Teratogenic
3kg
As cited in Schardein, 1993
50/125
Mohammad and St. Omer, 1988 As cited in Schardein, 1993
ester Rabbit 2,4-D butyiester
75
Sterility, fetal death 2,4-D diethylamine
Rat
Bifenox
Mouse
Teratogenic
Liberacki et aI., 1994
3%/10 ha pasture 0.5 LDSO 100
As cited in Schardein, 1993 Francis, 1986
(continues)
386
CHAPTER 14
Developmental and Reproductive Toxicology of Pesticides
Table 14.2 (continued) Dose Chemical
Species
Bromacil
Rat Rabbit
Bromoxynil
Toxicity profile
(mg/kg)b
Comments
References
250 ppm
Diet
As cited in Schardein, 1993
250ppm
Diet
Mouse
Developmentally toxicC
96.4
Rat
Skeletal variation
15
Butiphos
Rat
Developmentally toxic
Buturon
Mouse
Increased mortality
Chloramben
Rabbit
12.5
Rogers et aI., 1991 As cited in Schardein, 1993
100
Teratogenic Intrauterine growth
500
retardation
CDPR Database, 2000 (Worksheet 266 011 36993)
Chloridazon
Rat
Resorption
1/50 LDSO
As cited in Schardein, 1993
4-Chloro-2-methyl
Rat
Teratogenic
100
As cited in Schardein, 1993
phenoxyacetic acid ethyl ester 2-Chlorophenyl-4' -
Mouse
1000
Francis, 1990
Mouse
1000
Francis, 1990
3000
Tanaka et aI., 1997
nitropheny I ether 2-Chlorophenyl-4' nitrophenyl ether Chloroprophan
Mouse
2,4-D isooctyl ester
Rat
2,4-D
Rabbit
Developmentally toxic Teratogenic
750
Teratogenic
150
As cited in Schardein, 1993
75
Liberacki et aI., 1994
87.5
As cited in Schardein, 1993
75
Liberacki et aI., 1994
500
CDPR Database, 2000
isopropylamine 2,4-D propylene
Rat
Developmentally toxic
glycol butyl ether ester 2,4-D
Rabbit
triisopropanolamine Dalapon
Rat
Daminozide
Rat
Diallate
Rabbit
Skeletal effects
(Worksheet 006 036526) 1000 10
Khera et aI., 1979b As cited in Schardein, 1993
2,4-DM
Rat
2,5-Dichlorophenyl-
Mouse
1000
Francis, 1990
Mouse
1000
Francis, 1990
Mouse
500
Francis, 1990
Mouse
1000
Francis, 1990
Mouse
400
Francis, 1990
400
Roll and Matthiaschk, 1983
Developmentally toxic
3.4
4' -nitrophenyl ether 3,4 dichlorophenyl4' -nitrophenyl ether 2,6Dichlorophenyl-4' nitropheny I ether 2,5Dichlorophenyl-4' nitrophenyl ether 2,3Dichlorophenyl-4' nitropheny I ether Dichloroprop
Mouse
Teratogenic
Rat
Postnatal behavioral
5
Buschmann et aI., 1986
effects Dicotex
Rat
Dicuran
Rat
Teratogenic
20
As cited in Schardein, 1993
5000 (continues)
14.4 Toxicology Studies
38',
Table 14.2 (continued)
Dose Chemical
Species
Toxicity profile
Dinoseb
Rat
Teratogenic
(mg/kg)b
200ppm
Reproductive system
Comments
References
Diet, only
Giavini et aI., 1986
developmental
As cited in Schardein, 1993
toxic by po route Diuron
Rat
500
Endothall
Rat
25
EPTC
Rat
1/20 LDso
Ethalfluralin
Rat
1000
Hamster
Teratogenic
Byrd et aI., 1990a Minta and Biernacki, 1981
20 40
Rabbit Rabbit
Increased resorptions,
30
CDPR Toxicology Summary,
delayed ossification Fiuoxypyrmethyl-
As cited in Schardein, 1993
400
Rat Ethofumesate
1/ 2OLDSO
300
Rabbit Ethephon
As cited in Schardein, 1993 Trutter et aI., 1995
Rat
Skeletal variations
Rat
Reproductive system
1993 600
Carney et aI., 1995
hepty I ester Hexazinone
5000 ppm
Ioxynil octanoate
Mouse
100 ppm
Lenacil
Dog
500 ppm
Linuron
Diet; 3-G study
Rat
Reproductive system
Rat
Teratogenic
As cited in Schardein, 1993 Diet 3-G study
200
Reproductive system
Khera et aI., 1978 3-G study
100ppm 125 ppm
Rabbit Maleic hydrazide MCPA
Mecoprop
Embryotoxic
1/2 LDso
Teratogenic
1/2 LDso
Mouse
Teratogenic
200
Mouse
Teratogenic
400
Rat
Postnatal toxicity
Meturin
Rat
Molinate
Rat
Teratogenic: increased
Diet
Reproduction: sperm abnormalities, detached
As cited in Schardein, 1993 Khera et aI., 1979b As cited in Schardein, 1993 Roll and Matthiaschk, 1983 Roll and Matthiaschk, 1983 Buschmann et aI., 1986
13
As cited in Schardein, 1993
1/10 LDsO 35
CDPR Toxicology Summary,
resorptions, intrauterine growth retardation Rat
As cited in Schardein, 1993 CDPR,1993
1600
Rat Rat
Kennedy and Kaplan, 1984
125
Rabbit
1998 DPN #228 5 ppm males
Jewell and Miller, 1998
20 ppm females
heads, ovarian interstitial tissue vacuolation 3-Monochlorophenyl-
Mouse
1000
Francis, 1990
4'-nitrophenyl ether Monolinuran
Mouse
Naphoxyacetic acid
Rat
Nitrofen
Mouse
Mortality
25
Teratogenic
25
Growth retardation
As cited in Schardein, 1993
250
Henwood et aI., 1990
250
Nakao et aI., 1981
Teratogenic Rat
Teratogenic
121
dermal/oral
Costlow et aI., 1983
3-G & fertility Reproductive system Hamster
Teratogenic
studies 400
Gray et aI., 1985 (continues)
388
CHAPTER 14
Developmental and Reproductive Toxicology of Pesticides
Table 14.2 (continued) Dose Chemical
Species
Toxicity profile
Oxadiazon
Rat
Teratogenic:
(mg/kg)b
Comments
References CDPR,2000
12
postimplantation loss incomplete ossification Rabbit
Teratogenic: resorptions,
60
constraint-related arthrogryposis Oxyfluorfen
Rat
Teratogenic: early
CDPR,2000
15
resorptions, decreased fetal weight, skeletal malformations and variations As cited in Schardein, 1993
Paraquat
Mouse
1,2,3,7,8-Penta-
Mouse
Teratogenic
4000
100
Bimbaum et aI., 1991
Mouse
Teratogenic
2400
Bimbaum et aI., 1991
Rat
Fetotoxicity
250
Rabbit
Fetotoxicity
20
bromodibenzofuran 1,3,4,7,8-Pentabromodibenzofuran Phosphinothricin Pichloram Pichloram
Ebert et aI., 1990 Breslin et aI., 1991
1000
Rat Rabbit
400
lohn-Greene et aI., 1985
Rabbit
500
Breslin et aI., 1994
1000
Breslin et aI., 1994
ethylhexyl ester Pichloram
Rabbit
Abortion
Prometryn
Rat
Teratogenic
Propach10r
Rat
Equivocally teratogenic
1/5 LDSO 1/5 LDSO 0.2 mgim 3
triisopropanolamine
Propazine
Rat
Decreased fetal weight
Simazine
Rat
Teratogenic
SLA 3992
Rat
Teratogenic
20
Rabbit
Teratogenic
20
2,4,5-T
Mouse
Teratogenic
15
Rat
Teratogenic
50
Mirkova and Ivanov, 1981 Machemer et aI., 1992 As cited in Schardein, 1993
40 Teratogenic
20 113
Sheep Primate
Inhalation route
3-G study
Reproductive system Rabbit Hamster
As cited in Schardein, 1993
25
Growth retardation
40
Wilson, 1971 As cited in Schardein, 1993
Abortion 2,4,5-T butyl ester
Rat
Teratogenic
50
Mouse
Teratogenic
74
2,4,5-T phenol
Mouse
2,4,5-T propylene
Sheep
9 100
glycol butyl ether ester Tebuthiuton
Rat
Triallate
Rabbit
Trichloroacetic acid
Rat
1800 ppm
Diet
10 Developmentally toxic
330
Teratogenic
330
Smith et aI., 1988
(continues)
14.4 Toxicology Studies
389
Table 14.2 (continued)
Dose Chemical
Species
Toxicity profile
Trichloropyr
Rat
Reproductive system,
(mg/kg)h
250
fetotoxicity (FI and Fz gen) Trichloropyr
Comments
References
reduced litter size
CDPR,2000
and pup weight
Rat
Developmentally toxic
300
Breslin et aI., 1996a
Rat
Developmentally toxic
300
Breslin et aI., 1996a
Resorption
250
Hanley et aI., 1987
butoxyethyl ester Trichloropyr triethylamine Tridiphane
Mouse
Teratogenic Rat
Skeletal variants
Hanley et aI., 1987
100
Reproductive system
2-G and repro
Rao et aI., 1986
studies Trifluralin
Triisopropanolamine
Mouse
Skeletal variation
Rat
Depressed fetal weight
1000
Rabbit
Developmentally toxic
500
Rat
Beck, 1977
1000
Byrd and Markham, 1990 Breslin et aI., 1991
a Includes
plant growth regulators. bDose is oral unless stated otherwise; dose is the LOEL wherever effects were observed and the NOEL when there were no effects. C Developmental toxicity includes reduced fetal weight, increased embryo/fetal mortality, and/or increased developmental variations, but does not include malformations. Teratogenic is the term used to indicate the presence of malformations.
for both maternal toxicity (clinical signs, death) and developmental toxicity (skeletal anomalies, exencephaly, and umbilical hernia) was 1.0 mg/kg/day. However, the rabbit appeared to be the most sensitive laboratory animal to diquat in developmental toxicity studies. The NOEL for maternal toxicity (histopathological changes in the liver, intestine, and vasulature; mortality) was 3.0 mg/kg/day, but the developmental NOEL was below 1.0 mg/kg/day. Delayed ossification of the ventral tubercle of the cervical vertebrae was noted in all treatment groups compared to the controls. The incidence of fetal malformations was significantly greater in the low dose (1.0 mg/kg/day) and the high dose (10 mg/kg/day) compared to the controls; although the mid-dose lacked statistical significance, it appeared biologically significant (more than a two-fold increase over controls) and hence supportive of a treatment-related effect. Ethyl dipropylthiocarbamate (EPTC) The thiocarbamate class of pesticides has been shown to cause a wide range of effects. (EPTC) was determined to cause adverse developmental effects in the rat with a NOEL of 30 mg/kg/day, based on increased resorptions at levels below the maternal NOEL (100 mg/kg) (CDPR, 2000). Ethofumesate In a developmental tOXICIty study in rabbits, ethofumesate was determined to have adverse effects with a developmental NOEL of 30 mg/kg/day, based on increased resorptions and delayed ossification noted at the higher doses (300 and 3000 mg/kg/day) tested. Maternal toxic effects such as abortions and death were noted at the high
dose of 3000 mg/kg/day, resulting in a maternal NOEL of 300 mg/kg/day (CD PR, 2000). Molinate Results from several studies on the herbicide molinate have consistently demonstrated that exposure of male laboratory animals to the compound via the oral/inhalation route causes a decrease in fertility, abnormal sperm morphology, decreased epididymal sperm number, and/or testicular degeneration. Unexposed females mated to exposed males (rabbits/mice/rats) had significant (p < 0.05) preimplantation loss, possibly a result of the inability of the sperm to fertilize the ova. Female rats and mice exposed to molinate in the diet also exhibited significantly (p < 0.05) reduced litter sizes, along with histopathological abnormalities in the ovaries such as vacuolation and hypertrophy of the thecal/interstitial cells (CD PR, 2000). Recent data do suggest that humans are probably less sensitive and less likely than rats to experience the reproductive toxicity of molinate largely due to unequal rates of metabolism of molinate to molinate sulfoxide (Jewell and Miller, 1998). However, the relative degree of risk cannot be quantified at this time. In the rat, molinate demonstrated adverse effects such as increased resorptions and intrauterine growth prior to the onset of maternal toxicity with a developmental NOEL of 35 mg/kg/day. Nitrofen This pre- or postemergence herbicide induced a high incidence of diaphragmatic hernia and harderian gland alterations in mice fetuses subsequent to maternal oral exposure (Gray et al., 1983; Nakao et al., 1981). In rat studies, hydronephrosis and respiratory problems were noted (Costlow
390
CHAPTER 14
Developmental and Reproductive Toxicology of Pesticides
and Manson, 1980), whereas eye abnormalities were noted following percutaneous exposure to the dams (Francis and Metcalf, 1982). Exposure during only two gestational days altered the development of the para- and mesonephric ducts, resulting in renal malformations in females and agenesis of the vas, epididymis, and seminal vesicles in males (Gray et aI., 1985). The teratogenic activity of nitrofen has been attributed to alterations in maternal and fetal thyroid hormone status (Manson et aI., 1984). Triazines This class of compounds is heavily used throughout the world. They include herbicides such as atrazine, cyanazine, propazine, and simazine and the insecticide cyromazine. Eye defects such as anophthalmia, cryptophthalmia, microphthalmia, and cyclopia have been noted in animal studies for some of these compounds (CD PR, 2000). In a reproduction study in rats exposed to cyanazine, the significant toxicological finding was decreased pup viability: F1a pups at 250 ppm on day 21 and F2a pups at 150 and 250 ppm on day 4 (NOEL = 75 ppm). A possible adverse effect was indicated (pup NOEL < adult NOEL). In the rat teratogenicity study, the developmental toxicity NOEL was 5 mg/kg/day (increased number of fetuses and pups with microphthalmia or anophthalmia at 25 and 75 mg, decreased litter size and weight at 75 mg, increased total litter resorptions at 25 and 75 mg/kg, and decreased live litter size and survival to day 21 of lactation at 75 mg). Because developmental toxicity was seen at levels of cyanazine causing only slight maternal toxicity, the effects were considered adverse (CDPR, 2000; Iyer et al., 1999). Urea Herbicides Several urea herbicides induce genetic abnormalities in standard tests for genotoxic potential. They are generally the phenylureas and the effects of some of these compounds are detailed next. Diuron A widely used substituted urea herbicide, diuron induced wavy ribs at doses of 250 and 500 mg/kg/day (mid- and high dose) in rats. Ossification of the calvarium was delayed in fetuses of dams that received 125 mg/kg with the study yielding no NOEL (Khera et aI., 1979c). However, diuron did not produce any adverse effects for either reproduction or teratogenicity in studies submitted for registration (CDPR, 2000). Isoproturon Reproductive abnormalities, particularly those affecting sperm morphology and function, were noted in rats exposed to isoproturon (Behera and Bhunya, 1990). Maturational malformation of sperm and retarded spermatogenesis were also observed (Sarkar et aI., 1997). Linuron An antiandrogenic pesticide, linuron has been shown to induce a level of external effects consistent with its low affinity for the androgen receptor (AR), resulting in reduced anogenital distance, retained nipples, and a low incidence of hypospadias as well as malformed epididymides and testis atrophy (Gray et aI., 1999b). Additionally, linuron may produce Leydig cell tumors via an antiandrogenic mechanism where sustained
hypersecretion of luteinizing hormone (LH) appears to be responsible for the development of Leydig cell hyperplasia and adenomas (Cook et aI., 1993). Linuron may display several mechanisms of endocrine toxicity, one of which involves AR binding (Gray et aI., 1999b). Linuron produced malformations in the rat at 100 mg/kg/day but did not demonstrate teratogenic potential in the rabbit. Monolinuron, a related compound, has been shown to cause cleft palate in the mouse (Schardein, 1993). 14.4.2 INSECTICIDES
Included in this class of chemicals are the organophosphates, organochlorines, chlorinated cyclodienes, and carbamate esters. Many insecticides have demonstrated developmentally toxic effects and adverse effects on reproduction in animals (see Table 14.3). 14.4.2.1 Organochlorines, Including Chlorinated Cyclodier Aldrin Prenatal exposure to aldrin induced developmental changes (a decrease in the median effective time for incisor teeth eruption and an increase in the median effective time for testes descent) in rat pups and persistent behavioral alterations (the locomotor frequency of the experimental rats was higher than that of the controls at 21 and 90 days old) in adults after pregnant rats were subcutaneously treated with aldrin (1.0 mg/kg) or with its vehicle (0.9% NaCI solution plus Tween-80) from day 1 of pregnancy until delivery (Castro et aI., 1992). Aldrin may also have a direct inhibitory influence on gonadotrophin release and may exert a direct action on the testes (Chatterjee et aI., 1988). A review of the developmental toxicity of aldrin concludes that aldrin induces malformations (eye and digit defects and cleft palate) in mice and hamsters, a low frequency of malformations in rats, but was not teratogenic in dogs or swine. Aldrin is readily converted to dieldrin, and similar results were noted for dieldrin in these species as well as in the rabbit (Schardein, 1993). In the studies submitted for registration, possible adverse effects were noted in both the developmental toxicity studies (mouse and hamster at high doses) and the reproduction study in rats at the chronic toxicity dose range (CD PR, 2000). Amitraz In rats, adverse effects on reproduction were reduced litter size and substantial neonatal mortality at 200 ppm and slight to moderate neonatal mortality at 50 ppm, leading to a NOEL of 10.5 ppm. In the mouse, prolongation of the proestrus phase, a trend toward shortening of the diestrus phase, and depressed serum prolactin and progesterone levels with a NOEL of 25 ppm were observed (CD PR, 2000). In a developmental neurotoxicity study, rats were administered 20 mg/kg every third day and pups born were cross fostered. Open-field behavior (locomotion and rearing frequencies or immobility time) showed no significant differences, other than some transient delays (Palermo-Neto et aI., 1994). Postnatal exposure to
14.4 Toxicology Studies
391
Table 14.3 Developmental and Reproductive Toxicity Profile of Insecticidesa Dose Chemical
Species
Toxicity profileb
Aldicarb
Rat
Acutely toxic; hence not
Rabbit Aldrin
Apholate Bendiocarb 1,3-Bis(carba-
considered teratogenic
(mg/kg)C 0.04
Teratogenic Reproduction
Hamster
Embryotoxic
50
Teratogenic
50
Rat
Equivocally teratogenic
Rat
Reproduction
Risher et aI., 1987 As cited in Schardein, 1993
Mouse
Teratogenic
25 3-G study
Hodge et aI., 1967 As cited in Schardein, 1993
1
As cited in Schardein, 1993
10 2
DPN#50094
100
Mouse
References
0.1
Rat
Sheep
Comments
CDPR Database, 2000 Schardein, 1993
moylthio)-2-N,N-
dimethy lamino propane 100
Rat Hamster Bromophos
Mouse
Carbaryl
Mouse
Equivocally teratogenic
100 183
Nehez et aI., 1986
150 po or
As cited in Schardein, 1993
5660 ppm (diet) Rat
Reproduction
Dog
Resorption, teratogenic
Pig
2000
2-G study
500
3-G study
As cited in Schardein, 1993
6.25 30
Hamster
Fetal mortality
Rabbit
Teratogenic
150
Guinea
Teratogenic
300
125
pig Sheep
Teratogenic
Primate
Abortion
Cow Carbofuran
Mouse
250ppm
Diet
2 5.5
Liver histopathology
Rat
0.05 0.2
Adults
0.4
In utero, lactation
3 Rabbit Chlordane
As cited in Schardein, 1993
50ppm
Dog
Pant et aI., 1995, 1997
As cited in Schardein, 1993
0.5
Rat Mouse
Cell-mediated immune
Rat
(Temporary) tremors
80 g/kg
Days 7-17
Usami et aI., 1986
8
Throughout
Cranmer et aI., 1979
response
pregnancy 150-300 ppm
during and after
Ingle, 1952
gestation Chlordecone
Mouse
Reproductive failure
Rat
Fetotoxic
40ppm
Fertility study
10
As cited in Schardein, 1993
Reproduction Chlordimeform
Rat
Postnatal behavioral deficit
Chlorfenvinphos
Rat
Ossification disorders
Huber, 1965 Canon and Kimbrough, 1979
100
~g
Diet
As cited in Schardein, 1993
1/20 LDso
and carbaryl Chlormequat chloride
Rat Hamster
1000 ppm Teratogenic
Diet
100 (continues)
392
CHAPTER 14
Developmental and Reproductive Toxicology of Pesticides
Table 14.3 (continued)
Dose Toxicity profileb
Chemical
Species
Chlorpyrifos
Mouse
Fetotoxic
Rat
Reproduction
(mg/kg)C
Comments
References
2-0 study
Breslin et aI., 1996b
Deacon et aI., 1980
I
15
Ciafos
Rat
10kg
Coumaphos
Cow
28 g/45 kg
Crotoxyphos
Cow
Cyfiuthrin
Rat
As cited in Schardein, 1993 Topical route
BW 3.1 Reproduction: reduced pup viability; fetal malformations Rabbit
Postimplantation loss
Cypermethrin
Rat
Developmentally toxic
DDT
Mouse
CDPR Database
50ppm 0.46mg/m 3
Toxicology profile DPN #50317
20
Shawky et aI., 1984
1/40 LDso
Teratogenic Reproduction
7ppm
diet
As cited in Schardein, 1993
6-0 study
DEET
Rat
Reproduction
Rabbit
Developmentally toxic
10
Rabbit
Incomplete ossification, other
30
CDPR Database, 1999
Mouse
86
Nehez et al., 1986
Mouse
85
Nehez et aI., 1986
1/50 LDso
3-0 study
skeletal effects Demethylbromophos sodium Demethylbromophostetramethylammonium Dialifor
Hamster
Teratogenic
100
Robens, 1970a
Diazinon
Rat
Equivocally teratogenic
95-100
Campbell et aI., 1985 Dobbins, 1967
Reproduction: decreased pup survival, decreased ovary
CDPR Toxicology
< IOppm
Summary, 1999
weights Rabbit
0.25
Hamster Mouse
As cited in Schardein, 1993
30 Reduced postnatal growtb
0.18
Spyker and Avery, 1977
6.6
As cited in Schardein, 1993
Behavioral deficits Cow m-Dichlobenzene Dichlorvos
200
Rat Mouse
60
Rat
25
Rabbit
62
As cited in Schardein, 1993
8.5
Pig Reproduction
6.2
Cow Dieldrin
Mouse
Teratogenic
6-0 study
Reproduction 6
Rat
3-0 study
Reproduction Hamster
Embryotoxic, teratogenic
sulfonamide
30
Rabbit
6
Sheep
25ppm 0.2
Dog N,N -diethyl-benzene-
As cited in Schardein, 1993
15
Rat
Teratogenic
300
Rabbit
Resorption
25
Leland et al., 1992
(continues)
14.4 Toxicology Studies
393
Table 14.3 (continued)
Dose Chemical
Species
Toxicity profileb
N ,N -Diethyl-n-
Rat
Reduced fetal weight
Rabbit
O,O-Dimethyl-S-
750
References Schoenig et aI., 1994 Wright et aI., 1992 Schoenig et aI., 1994
325
Pig Sheep
Dimethoate
Comments
Reproductively neurotoxic
toluamide Diflubenzuron
(mg/kg)C
100ppm
Fertility study
100ppm
Fertility study
As cited in Schardein, 1993
Rat
Teratogenic
3
Khera et aI., 1979c
Cat
Teratogenic
12
Khera et aI., 1979a
Mouse
Reproduction
Mouse
Developmentally toxic
5-G study
Budreau and Singh, 1973 As cited in Schardein, 1993
8
(2-acetylaminoethyl) dithiophosphate Empenthrin
Rat
Endosulfan
Rat
Delayed ossification
Endrin
Mouse
Teratogenic
2.5
As cited in Shepard, 1998
Hamster
Embryotoxic
5
As cited in Shepard, 1998
Behavior deficits
5
Gray et aI., 1979
Teratogenic
0.75
Ethohexadiol
Rat
500
Developmental study
Fertility study
2 ml/kg
Kaneko et aI., 1992 Gupta et al., 1972
5
Occlusive
As cited in Schardein, 1993
cutaneous Fenamiphos Fenbutatin oxide
Rat
3
Rabbit
2.5
Rat
Reproduction: decreased pup
Machemer et aI., 1992
250ppm
CDPR Database, 1999
weight gain in F 1 and F2 Fenitrothion
Mouse
DPN#214 80ppm
Rat
Postnatal behavioral deficits
Fensulfothion
Rabbit
Teratogenicity: malfonnations,
Fenthion
Rat
Diet
Lehotzky et aI., 1989
10 0.1
CDPR Worksheet 234-
incomlpete ossification Rat Fluvalinate
Rabbit
084054352, 1987 18
Astroff et aI., 1996
Epididymal cytoplasmic vacuolation Teratogenicity: malfonnations
CDPR Database, 1997 14ppm
2-G study
25
Fonnothion
Rabbit
Heptachlor
Rat
Heptachlor and
Rat
7ppm
Imidazolidinone
Rat
240
Isobenzan
Mouse
1 ppm
lOppm
(Worksheet 093 90460) CDPR Database, 1999
30 Cataracts in both generations
Benes et aI., 1973
As cited in Schardein, 1993 2-G study
Ruttkay-Nedecka et aI., 1972 Eisler, 1970
heptachlor epoxide As cited in Schardein, 1993 Diet
Isofenphos
Rat
Leptophos
Rat
Developmentally toxic
Lindane
Mouse
Reduced fetal growth
Rat
Decreased fertility
0.5
Naishtein and Leibovich, 1971
Mortality and developmental
0.5
As cited in Schardein, 1993
10 12.5 ppm
Mast et aI., 1985 Diet
Kanoh et aI., 1981 As cited in Schardein, 1993
delay 100ppm
Malathion
Rabbit
Inhibited development
40
Hamster
Inhibited development
20
Rat
Palmer et aI., 1978 As cited in Schardein, 1993
300 240
Rabbit
3-G study
Khera et aI., 1978 2-G study
As cited in Shepard, 1998
100 (continues)
394
CHAPTER 14
Developmental and Reproductive Toxicology of Pesticides
Table 14.3 (continued)
Dose Chemical
Species
Toxicity profileb
Metam sodium
Rat
Teratogenicity: decreased fetal
(mg/kgY
Comments
CDPR Database, 1999
5
weights, severe malformations Rabbit
Postimplantation loss, severe
References
DPN#50150 5
defects (cleft palate, meningiocele) Methamidaphos
Rabbit Rat
Methomyl Methyl demeton
Rabbit Rat
Reproduction
Rat
Reproductive toxicity (altered
(oxydemeton
Methoxychlor
2-G study
CDPR Database, 2000
100 ppm
Diet
As cited in Schardein, 1993 As cited in Schardein, 1993 CDPR Database
and decreased fertility) Mouse
2.5 mg/gBW
Rat
2.5 mg/g BW
Rat
Fetopathic
Cow Methyl parathion
10ppm
ovarian, epididymal histology,
methyl) MethylISP
CDPR Database, 2000
2.5 Decreased pup weight gain
Rat
Developmentally toxic
Wu et al., 1989 Khera et al., 1978
100 9.9
As cited in Schardein, 1993
5
Frosch, 1990
Teratogenic Naled
Rat
Naphthylisothio-
Rat
Liver histopathology
100
Khera et al., 1979b
100
As cited in Schardein, 1993
cyanate Nicotine
Mouse Rat
As cited in Schardein, 1993
Resorption Teratogenic Developmentally toxic Postnatal behavioral deficits
0.008 I-lg/g BW 2
sc route
25mg
iv route Mini-osmotic pump
Hydrocephalus Toxic action in oocytes Rabbit Oxamyl
0.4
Antenatal
5
sc route
N-Octyl-
Reproduction
4 CDPR Database, 2000
10,000ppm
Rat
Vara and Kinnunen, 1951
1- and 3-G studies
Reproduction Rabbit
As cited Shepard, 1998
Kennedy, 1986
100
Rat
As cited in Schardein, 1993
bicycloheptene dicarboximide Pentachlorophenol
Courtney et al., 1970a
75
Mouse Hamster
1.25
Rat
4
Embryotoxic
As cited in Schardein, 1993 Less toxic by po route
Developmentally toxic
Welsh et aI., 1985 As cited in Schardein, 1993
Reproduction Phorate
1.94 mg/m 3
Rat
Phosalone Phosfolan
1/10 LDso Shen, 1983
Rat
Teratogenic
Rat
Embryotoxic
0.3
Teratogenic
0.3
Rabbit Photodieldrin
3-G study
Reproduction
Rat
As cited in Schardein, 1993
3-G study
Reproduction Mouse
Inhalation route
Selectively toxic
As cited in Schardein, 1993
35
Mouse
0.6
Rat
0.6
Chemoff etal., 1975
(continues)
14.4 Toxicology Studies
395
Table 14.3 (continued) Dose Chemical
Species
Photomirex
Rabbit
Toxicity profileb
(mglkg)C
Cataracts and reduced survival
Mouse
Teratogenic
20ppm
Reproduction
As cited in Schardein, 1993
As cited in Schardein, 1993 study Ogata et aI., 1993
660
As cited in Schardein, 1993
3,000
Rat Rabbit
References
10
Rat
in generational study Piperonyl butoxide
Comments
Equivocally teratogenic
Potassium arsenate
Sheep
Propoxur
Rat
Neonatal CNS impairment
Ronnel
Rat
Developmentally toxic
100 0.75
Rabbit
Teratogenic
Fox
Teratogenic Embryotoxic
1,000 ppm
Diet
2.5 Nafstad et aI., 1983 Berge and Nafstad, 1983
100
Khera et aI., 1981
2.5
Rotenone
Rat
Sarin
Rat
380 J.l-g/kg
Rabbit
15 J.l-glkg
Sodium arsenite
Hamster
25
Lu et aI., 1984 Bates and LaBorde, 1986 Teratogenic by iv route
Sodium selenite
Hamster
Soman
Rat
Teratogenic
Rabbit Sulfluramid
Rabbit
Spinosad
Rat
Hood and Harrison, 1982 WiIIhite, 1981
90
Ferm et aI., 1990
165 J.l-g/kg
Bates et aI., 1990
15 J.l-glkg Neonatal mortality
Stump et aI., 1997
0.3 200 10 (2-0)
Maternal NOEL
CDPR Database
= developmental NOEL
Rabbit 2,3,5,6-Tetra-
50
Rat
Zielke et aI., 1993
150
chloropyridine Thiometon
Rabbit
5
As cited in Schardein, 1993
Toxaphene
Mouse
Teratogenic
15
As cited in Schardein, 1993
Rat
Decreased skeletal ossification
35
Tribufos
Rat
Trichlorfan
Mouse
28
Rat
Teratogenic
400
Hamster
Teratogenic
400
Pig
Teratogenic
60
Trichloro-
Astroff et aI., 1996
300
Teratogenic
As cited in Schardein, 1993
7.5
acetonitrile 1,2,3-
Rat
600
Black et aI., 1983
Trichlorobenzene 1,2,4-
Rat
300
Trichlorobenzene 1,3,5-
Rat
600
Trichlorobenzene Triphenyltin
Rat
hydroxide
20 Equivocally reduced fertility
100ppm
As cited in Schardein, 1993 Reproduction study
Valexon
Rat
Teratogenic Testicular toxin
a Includes
90 0.7
Fertility study
chemosterilants, repellants, and growth regulators. bDevelopmental toxicity includes reduced fetal weight, increased embryo/fetal mortality, and/or increased developmental variations, but does not include malformations. Teratogenic is the term used to indicate the presence of malformations. cDose is oral unless stated otherwise; is dose the LOEL wherever effects were observed and the NOEL when there were no effects.
396
CHAPTER 14
Developmental and Reproductive Toxicology of Pesticides
amitraz caused transient developmental and behavioral changes in the exposed offspring in a subsequent study (Palermo-Neto et aI., 1997). Results showed that the median effective time (ET50) for fur development, eye opening, testis descent, and onset of the startle response was increased in rats postnatally exposed to amitraz compared to those of the control. However, the age as incisor eruption, total unfolding of the external ears, vaginal and ear opening and the time taken to perform the grasping hind-limb reflex were not affected by amitraz exposure. Dibromochloropropane Dibromochloropropane (DBCP) is a brominated organochlorine that was used as a nematocide from the mid-1950s until its ban in the United States in the late 1970s (Whorton and Foliart, 1983). It was used in the United States, mostly in Hawaii and along the southern Atlantic and Pacific coasts, to protect citrus, grapes, peaches, pineapples, soybeans, and tomatoes. Of the pesticides studied to date, DBCP is the most toxic to the human male reproductive system. Torkelson et al. (1961) reported that DBCP caused testicular atrophy in rats, guinea pigs, and rabbits. Azoospermia and oligospermia were reported in DBCP production workers and these were linked with the length of time the persons worked with the chemical (Whorton et aI., 1977). Higher levels of follicle-stimulating hormone (FSH) were noted in a number of these individuals as well as in those who did not revert to normal spermatogenic levels long after exposure. Data from the animal studies revealed severe testicular insults, including degenerative changes in the semeniferous tubules, increase in Sertoli cells, reduction in the number of sperm, and increased abnormalities in the sperm cells. Epididymal (Kluwe, 1981), posttesticular effects (Kluwe et aI., 1983), and in vitro effects (Bartoov et al., 1987) were also noted. The mechanism of action was determined to be at the level of mitochondrial respiration, and DBCP (the parent compound) was demonstrated to inhibit carbohydrate metabolism at the reduced nicotinamide adenine dinucleotide (NADH) dehydrogenase step in the mitochondrial electron transport chain of rat sperm (Greenwell et aI., 1987). Workers on banana crops documented convincing evidence of increased spontaneous abortion in their family histories; follow-up studies among production workers in Israel showed that some recovered testicular function, but among their offspring, there was a predominance of females. Those who did not recover from azospermia were those with high levels of FSH (Goldsmith, 1997). No other studies have shown increased birth defects or increased infant mortality. Dichlorodiphenyltrichloroethane (DDT) Several investigators have examined the effects of DDT on both development and the reproductive system. DDT and its metabolite dichlorodiphenyldichloroethylene (DDE) have resulted in eggshell thinning in birds of different species (Porter and Wiemeyer, 1969). Thinner shells are associated with a higher disappearance and/or destruction of eggs, and it is this phenomenon, along with the huge public outcry subsequent to the publishing of Rachel Carson's book Silent Spring that led to
the ban on the use of DDT in the United States. The inconsistent effects of DDT and DDE on eggshell thickness might reflect differential sensitivities among species of birds. The thinning is thought to be similar to the way estrogen inhibits the formation of eggshells, although the calcium adenosinetriphosphatase (ATPase) inhibition by DDT demonstrated by Matsumura and Ghiasuddin (1979) may also be responsible. The interference (direct or indirect) with fertility and reproduction is thought, therefore, to be related to steroid metabolism and the inability of the bird to mobilize sufficient calcium to produce a strong eggshell to withstand the rigors of being buffeted around the nest; the resultant cracking allows the entry of bacteria, causing the developing embryo to die (Carson, 1962; Peakall, 1970). In male cockerels and rats, DDT (20% o,p'-DDT and 80% p,p'-DDT) reduced testicular size; in females o,p'-DDT administration yielded estrogenic effects such as edematous, blood-engorged uteri (Ecobichon and MacKenzie, 1974; Hayes, 1959). DDT has resulted in uterotrophic effects; o,p'-DDT has demonstrated increased weight in ovaries and uteri and an advanced time of vaginal opening (Gellert et aI., 1972). Other estrogenic effects such as increases in cellular progesterone receptors, tissue mass, DNA synthesis, and cell division have also been reported (Ireland et aI., 1980; Mason and Shulte, 1980; Nelson, 1974; Robison and Stancel, 1982). Kupfer and Bulger (1976) have shown that the o,p'isomer competes with estradiol for binding with estogen receptors in rat uterine cytosol. A review of studies conducted in Israel, India, and the Ukraine suggested that maternal and fetal tissue levels of DDT and metabolites were higher in fetal deaths than in other pregnancies; however, studies from Poland, Italy, and Florida observed no significant differences in the levels of organochlorine pesticides in samples from normal versus aborted pregnancies (Sever, 1988). DDT is known to be lipophilic and a potent bioaccumulator and can be detected in fatty tisues in the food chain long after its use. This and the fact that it is still widely used as an efficient agent in the control of mosquitoes causing malaria can result in marked long-term ecological impact. Methoxychlor Methoxychlor is a chlorinated hydrocarbon insecticide that has a much lower bioaccumulation potential than that of DDT. Early studies in pregnant rats demonstrated maternal and fetal toxicity (wavy ribs) at exposures of 200 and 400 mg/kg (Khera et al., 1978). In rats, methoxychlor is metabolized in vivo to 2,2-bis(p-hydroxyphenyl)-I, 1,1trichloroethane (HPTE), the active estrogenic form. It has direct estrogenic effects on the rat uterus and also inhibits the decidual cell response, which is an accepted model for implantation-associated effects. It has adverse effects on fertility, early pregnancy, and in utero development in females; in adult males, adverse effects such as altered social behavior following prenatal exposure to methoxychlor were noted (Cummings, 1997; Cummings and Gray, 1989). Recent work in mice concluded that neonatal exposure to methoxychlor at doses of 0.1, 0.5, and 1.0 mg/kg/day did not interfere with mating, but significant alterations were seen in initiating and/or
14.4 Toxicology Studies
maintaining pregnancy. The deleterious effects on pregnancy may be due to the influence of neonatal methoxychlor treatments on the hypothalamic-pituitary-ovarian axis as well as on possible alteration of the uterine environment (Swartz and Eroschenko, 1998). The significance of this toxicity with respect to human health remains to be determined. In LongEvans hooded rats, methoxychlor affects the central nervous system (CNS), the epididymal sperm numbers, and the accessory sex glands and delays mating without significantly affecting the secretion of LH, prolactin, or testosterone. These data indicate that methoxychlor did not alter pituitary endocrine function in either an estrogenic or an antiandrogenic manner (200-400 mg/kg/day) and demonstrate a pronounced degree of target tissue selectivity (Gray et aI., 1999a). Chlordecone Chlordecone was sold as an insecticide and fungicide between 1958 and 1975. Better known by its trade name Kepone™, it was widely used and caused contamination of the lames River near the plant where it was manufactured in Virginia. Mirex, an insecticide that photodegraded to Kepone, has also been extensively studied and the toxicity of both compounds will be summarized here. Production workers exposed to chlordecone were noted to be oligospermic and had reduced sperm motility (Taylor et aI., 1978). Chlordecone primarily affects sperm motility and viability via mechanisms that are not completely understood. One study in CD-l mice documented that the pool of potentially ovulatory follicles was reduced subsequent to prolonged exposure to Kepone (Swartz and Mall, 1989). It is a potent inducer of the mixed-function oxidase system and may affect fertility by stimulating hepatic degradation of steroids. Although chlordecone (probably as the hydrate) has a binding affinity for estrogenic sites (Hammond et aI., 1979), other studies have concluded that the reproductive toxicity was not by a mimicry of estrogen (Cochran and Wiedow, 1984). Instead, it appears that chlordecone can act on the hypothalamic-pituitary axis (Hong et aI., 1985). Exposure to 50 and 75 mg/kg of chlordecone in the female rat before or after mating substantially reduced fertility (Uphouse, 1986). A review of the developmental toxicity of chlordecone exposure during gestation indicated fetotoxicity in mice and rats and some CNS impairment in fetuses in the rat (Schardein, 1993). Lindane (Hexachlorocyclohexane-HCH) Lindane, a nonaromatic chlorinated cyclic hydrocarbon, has shown some effects on the female reproductive system (Welch et aI., 1971), but, on the whole, the results are variable. In male rats fed 75 mg/kg/day of lindane for 90 days, testicular atrophy, degeneration of semeniferous tubules, and disruption of spermatogenesis were reported (Shivanandappa and Krishnakumari, 1983). Reduced epididymal sperm concentrations were also noted in other studies in rats given a single dose of 30 mg/kg/day (Dalsenter et aI., 1996). As with DDT, the estrogenic effects of lindane might occur via estrogen receptors, but studies in female Long-Evans rats dosed at 40 mg/kg/day for 7 days or in ovariectomized rats for 5 days did not show altered serum
397
estradiol concentrations or changes in the number of estrogen receptors. No changes in estrogen-dependent induction of progesterone in the hypothalamus, pituitary, or uterus were observed. Hence, it is thought that lindane may act via altering multiple processes such as the gamma-aminobutyric acid (GABA)-nergic system or via altered growth factors (Laws et aI., 1994). In reproduction studies, fertility was not reduced, but most of the F 1 pups died shortly after birth. Erratic estrous cycles were also noted with exposure to lindane (Gray et aI., 1988). The adverse liver effects in the rat reproduction study do not appear to be important to human safety because they were reversible after much higher exposures in the combined study (CDPR, 2000), and the kidney effects (apparent hydronephrosis) were species- and sex-specific. The reproductive NOEL was 20 ppm, based on reduced neonatal pup survival (largely due to total litter losses), slightly reduced pup growth rate, and slightly slower pup development such as delays in hair growth and tooth eruption (CD PR , 2000). In a suicidal poisoning, maternal ingestion of lindane resulted in the death of twin fetuses (Konje et aI., 1992). In addition to these and other endocrine effects, a dose-related increase in the incidence of fetuses with an extra 14th rib in CFY rats and an extra 13th rib in rabbits has been reported at 15 mg/kg/day but a lack of teratogenicity was determined (Palmer et aI., 1978). This was consistent with the negative teratogenicity results in mice, mutagenicity studies, and 3-generation rat reproduction studies. Regional changes in brain noradrenaline and serotonin levels have also been reported as developmental effects (Rivera et aI., 1991). Adverse effects, however, were not noted in the developmental toxicity studies submitted to the CDPR, and lindane and related HCH isomers are not listed as chemicals known to the state to cause reproductive toxicity under Proposition 65 or the Safe Drinking Water and Toxic Enforcement Act of 1986. 14.4.2.2 Organophosphates These compounds cause a combination of a reduction in brain acetylcholinesterase activity and altered reproductive behavior in a number of species. The reduced acetylcholinesterase has been associated with decreased egg production and serum LH and serum progesterone (Rattner et aI., 1982). The standard dominant lethal test in mice was negative for dichlorvos (Dean and Blair, 1976). Possible mechanisms of toxicity from studies on trichlorfon and parathion in the rat are thought to involve interference with steroid hormone binding to receptors in the liver, adrenal, uteri, and testes (Trajkovic et aI., 1981). In a case report, the organophosphate pesticide mercarbam crossed the placental barrier and caused the death of a 5-month fetus (Schardein, 1993). There has been some indication that organophosphates (OPs), in general, may affect the menstrual cycle and cause an early menopause in humans. Reproductive effects from exposures to mixtures of OPs have been documented by Mattison et al. (1983) and Nakazawa (1974) among women in agriculture. These effects included abnormal menstruation (e.g., hypermenorrhea, oligomenorrhea, amenorrhea) and early menopause. On the other hand, Willis
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et al. (1993) found no effects of pesticide exposures (including methyl parathion) on the pregnancy outcome among 535 women enrolled in a Southern California community clinic perinatal program.
Chlorpyrifos The developmental toxicity and reproductive toxicity of this widely used compound have been extensively studied. The studies submitted for registration under FIFRA did not show adverse developmental or reproductive effects. However, some recent studies on the role of cholinesterase in morphogenesis have used chlorpyrifos as the model compound and intimated an influence in learning disabilities among children who were exposed in utero or during the early postnatal period (Roy et aI., 1998). Chlorpyrifos, a phosphorothioate, undergoes oxidative desulfuration to form chlorpyrifos-oxon, which can then phosphorylate acetylcholinesterase, rendering it incapable of metabolizing the neurotransmitter acetylcholine to choline and acetate. The oxon, however, either may be detoxified by combining with carboxylesterase or may be hydrolyzed by oxonase to metabolites not capable of combining with acetylcholinesterase. It has been suggested that inhibition of DNA and protein synthesis may be direct noncholinergic effects of chlorpyrifos, but other mechanisms such as alterations in blood flow patterns may also be involved. In reviewing several studies as part of the risk characterization of the compound, it appears that newborn and juvenile rodents are more susceptible to the toxic effects of chlorpyrifos than adults. The increased susceptibility of young rats appeared not to be due to a difference in the affinity of the oxon for the acetylcholinesterase, but rather incomplete development of the enzyme systems that detoxify the oxon (CDPR, 2000). In vivo studies in rats have used newborns to explore the effects of chlorpyrifos on the ontogeny of the mammalian nervous system during the equivalent last trimester of human fetal development. However, sublethal concentrations were administered to postnatal day 1 rats via intraperitoneal or subcutaneous routes. In the human, in utero exposures are likely to be mitigated by the mother's metabolism and the availability of chlorpyrifos to the fetus may therefore not be comparable to newborn rodents being directly dosed. Aggregate exposure estimates, acute illness reports, and dermal exposure from surface wipes also point to low levels of chlorpyrifos (CDPR, 2000; Lewis et aI., 1994; Lu and Fenske, 1999). Higher exposures to children have been reported for oral nondietary and dermal exposure (Gurunathan et aI., 1998); the assumptions used, however, may not be appropriate. Thus, although the fetus (independent of maternal metabolism) may be theoretically more susceptible than the adult, the evidence that chlorpyrifos causes developmental neurotoxicity under physiologically relevant conditions is not compelling. The inhibition of cholinesterase activity (brain and liver) in pups was detected at doses nearly lethal to the rat dams (Tang et aI., 1999). Thus, taken together, the studies suggest that maternal effects will be observed prior to levels causing developmental toxicity. Chlorpyrifos has been banned in the United States for indoor application.
Diazinon This organophosphate insecticide has been tested extensively and yielded variable results for reproductive and developmental endpoints. Spyker and Avery (1977) exposed pregnant mice (9 mg/kg) and observed behavioral effects and functional impairments in overtly normal offspring along with neuropathology in the forebrain. The standard developmental toxicity studies in rats and rabbits did not demonstrate adverse effects (CD PR, 2000). In another review, malformations were not reported in hamsters and rabbits; renal, rib, and limb malformations and anomalies of the CNS and digits were noted in rats, and skull and teeth abnormalities were noted in puppies (Schardein, 1993). In both one-generation and two-generation reproduction studies, a variety of adverse effects were observed, including a decrease in the gestation index (number of litters with live offspring/number pregnant), a reduction in ovarian weights, and a prolonged gestation length. The reproduction NOEL was less than 10 ppm (1 mg/kg/day) or the lowest observed effect level (LOEL) was 10 ppm (CDPR, 2000). Behavioral effects were confirmed in neurotoxicity studies in rats (CD PR, 2000); hence, the developmental neurotoxicity of diazonon may need evaluation. Dimethoate Developmentally toxic effects were not noted in either the rat or the rabbit studies submitted under FIFRA (CDPR, 2000); however, rib defects in rats and polydactyly in cats were noted (Khera et aI., 1979c). In the mouse, variable results were noted with embryotoxicity without teratogenicity in earlier studies (Scheufler, 1975, 1976) and no adverse effects were found in another study (Courtney et aI., 1985). Reduced number of pregnant females and lowered pup weights and reduced litter sizes were noted in a second-generation reproduction study in rats (CDPR, 2000). Fenthion A reduction in fetal weights at 80 mg/kg was noted in a study in mice with an increase in malformations in 14.5% of the offspring (Budreau and Singh, 1973). In rats, however, a marginal increase in resorptions was noted at 18 mg/kg/day, demonstrating no other adverse developmental effects at lower doses and yielding a NOAEL of 4.2 mg/kg/day. Exposure to fenthion in the diet at 14 and 100 ppm in the reproduction study demonstrated epididymal cytoplasmic vacuolation (ECV) associated with decreased fertility, reduced survivability, and postnatal growth retardation resulting in a reproductive NOEL of2 ppm (CDPR, 2000). Parathion and Methyl Parathion Three multigeneration reproductive toxicity studies in rats have been submitted to regulatory agencies (CDPR, 2000), and decreased pup survival was consistently found in all three studies. A search of the literature revealed one study showing ovarian toxicity in rats (Dhondup and Kaliwal, 1997) and one study showing possible sperm abnormalities in mice (Mathew et aI., 1992). Testicular and reproductive effects have also been reported in avian species (Maitra and Sarkar, 1996; Solecki et aI., 1996). In the absence of clinical symptoms and behavior changes, reproductive effects such as reduction in the number of eggs laid (~ 20% reduction), egg
14.4 Toxicology Studies weight (~9% reduction), and eggshell thickness (7-10% reduction) were noted in the Japanese quail. Suppression of growth and ossification in both mice and rats were observed subsequent to methyl parathion exposure; in the mouse, high mortality and cleft palate were noted (Tanimura et aI., 1967). No other epidemiological data specific to methyl parathion are available. Malathion Malathion does not appear to cause adverse developmental or reproductive effects (CD PR, 2000). However, malathion administered to mice at 250 mg/kg (corresponding to 1/12 LDso) and examined 4, 14, 18, and 26 days after injection induced teratozoospermia. Sperm count at different time intervals was significantly increased compared to the controls, and there was a parallel increase in depletion of the semeniferous tubules; all germinal cell populations studied were affected by malathion, especially mice spermatid differentiation, which compromises mostly the flagella, perhaps due to an alkylating effect that disturbs the normal assembling of tail structural protein components (Contreras and Bustos-Obregon, 1999). No evidence for histopathological alteration or teratogenic anomalies in the fetuses was demonstrated, although placental transfer of malathion was indicated by the presence of the insecticide residues in fetuses from rats fed wheat material containing bound residues of malathion S-I,2-di(ethoxycarbonyl)-ethylO,O-dimethyl phosphorodithioate (Bitsi et al., 1994). The reproductive effects of the aerial spraying of the organophosphate insecticide malathion in California have been examined in a case-control study of spontaneous abortions «28 weeks) and stillbirths; relative risks were 1.21 (95% Cl = 0.94-1.52) for spontaneous abortions and 1.51 (95% Cl = 0.21-11.3) for stillbirths. A cohort of 7450 pregnancies identified through three Kaiser-Permanente facilities in the San Francisco Bay area, in relation to exposure to the pesticide malathion, applied aerially to control an infestation by the Mediterranean fruit fly, was examined for reproductive outcomes. No important association was found between malathion exposure and spontaneous abortion, intrauterine growth retardation, stillbirth, or most categories of congenital anomalies. Gastrointestinal anomalies noted were related to second-trimester exposure (OR = 2.6), based on 13 cases, and were not specific to any particular International Classification of Diseases code (Thomas et aI., 1992). Trichlorfon Trichlorfon, an organophosphate insecticide, has been associated with a cluster of babies born with Down's syndrome in Hungary (Czeizel et al., 1993). A case-control study and environmental investigations reported excessive use of trichlorfon at local fish farms. The high content of trichlorfon in the diet of pregnant women, including all of the mothers with affected offspring, along with an absence of known teratogenic factors such as familial inheritance and consanguinity, was supportive of the association. Trichlorfon or chlorophos, marketed under the brand name Dipterex, displayed embryotoxic and teratogenic effects in the Wistar rat after an oral dose of 80 mg/kg during a critical period of embryogenesis, but was negative at the low dose of 8 mg/kg (Martson and Voronina, 1976). A review of this insecticide and anthelmintic demonstrated a potent
399
developmental toxicity profile in laboratory animals such as the rat (oral and inhalation), mouse (oral), and hamster (oral), but teratogenicity was not noted if the exposure was via the intraperitoneal route (Schardein, 1993). Recent work from Norway characterized the teratogenic effects of trichlorfon on the guinea pig brain by determining the effective dose and sensitive period (Hjelde et al., 1998). Following oral or subcutaneous exposure, almost all regions of the brain were reduced in weight. The cerebellum was the most vulnerable region, but the medulla and hypothalamus were also greatly reduced in weight. While the mechanism behind the teratogenic effect is not known, alkylation of DNA or altered DNA repair may be involved. 14.4.2.3 Carbamates
Similar to organophosphates, carbamates result in the inhibition of cholinesterase and also exert an anesthetic effect. The dithiocarbamates have been used as fungicides and will be discussed under that category. Carbofuran Most studies with carbofuran have been negative for teratogenicity. However, one study in mice resulted in fine-structure abnormalities in mice (Schardein, 1993); in the FIFRA reproduction study, reduced body weight gain in adults and birthweights in offspring worsening to 15% by weaning were noted (CDPR, 2000). Decreased weights of the epididymides, seminal vesicles, ventral prostate, and coagulating glands were also noted in rats, along with decreased sperm motility, reduced epididymal sperm count, and increased morphological abnormalities in the head, neck, and tail regions of spermatozoa (Pant et al., 1995). Testicular and spermatotoxic effects were also noted at levels higher than 0.2 mg/kg in rats exposed to carbofuran in utero or via lactation (Pant et aI., 1997). Studies from Sri Lanka in rats have concluded that carbofuran administered orally 0.2, 0.4 and 0.8 mg/kg during early gestation is detrimental to pregnancy (enhanced preimplantation losses) and possibly harmful to neonatal development (Jayatunga et aI., 1998a). Similarly post-implantation losses were noted after exposure to carbofuran during mid-gestation (Jayatunga et al., 1998b). Bendiocarb This compound is a residual insecticide and appears to cause adverse effects on reproduction in the rat. A decrease in the number of pups and reduced survival at dose levels of 200 and 250 ppm resulted in a NOEL of 50 ppm (CD PR, 2000). Thiodicarb Adverse effects were documented in the rat reproduction study submitted for registration. Decreased pup weight gain at 100 ppm and above and decreased viability at 900 ppm were noted. The parental NOEL was 100 ppm (decreased Fo and Fl body weights); the reproductive LOEL was 100 ppm (reduced pup weight gain). A statistically based estimate of the NOEL provided an NEEL (no expected effect level) of 81 ppm in males and 80 ppm in females (CD PR, 2000).
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Zineb and Thiram The exposure of rats to zineb and thiram has documented an alteration (prolongation) of the estrous cycle in association with a reduction in ovarian and uterine weights (Ghizelea and Czeranschi, 1973). Direct gonadal effects have been noted in mouse and rabbit oocyctes, resulting in inhibition of oocyte meiotic maturation and prevention of germinal vesicle breakdown. In the mouse oocyte exposed to isopropyl-N-phenylcarbamate, the formation of nuclear condensates (macromolecules) has also been observed by Crozet and Szollosi (1979). Carbaryl Carbaryl (N -methyl-I-naphthyl carbamate; Sevin®) is a carbamate insecticide that is widely used, resulting in exposure during its use as well as via consumption of treated food. It is not metabolized to an active intermediate; the parent compound itself is thought to be the active agent. Carbaryl acts via inhibition of acetylcholinesterase by carbamylation of the active-site serine residue. Adverse effects on rodent spermatogenesis at 0.4-5 mg/kg ip (intraperitoneal) or orally have been reported (Weil et al., 1972), but several studies do not support this finding. Hence, although the data suggest that carbaryl does not produce a testicular effect similar to DBCP, human personnel may be affected as demonstrated by some studies (Wyrobek et al., 1981). The teratogenicity of carbaryl has been reviewed extensively, but the results can best be summarized as equivocal (Schardein, 1993). Data are available in a number of species; early studies in the rat reported terata but subsequent studies were negative. Studies in mice were variable, and the eye defects seen previously were not observed in later studies. In rabbits, the data are contradictory: Omphalocele and skeletal variations were noted at the high dose; however, malformations were not seen in the other study. A recent study in rabbits submitted to the CDPR documented agenesis of the gall bladder at the high dose in two fetuses of different litters, along with reduction in the size of the gallbladder and a missing bile duct, presenting the case for a continuum of effects (CD PR, 2000). Sheep that were fed carbaryl in the diet demonstrated heart defects; in the dog, where pregnant females were dosed with 6.25-50 mg/kg/day in the diet, malformations included abdominothoracic fissures, intestinal agenesis and displacement, brachygnathia, failure of skeletal formation, anurous (no tail), and superfluous phalanges. In addition to these malformations that were seen in several puppies, resorption was noted in 21 of 181 pups (11.6% fetal incidence; 21.1 % litter incidence). However in primates, minature swine, and cattle, exposure to carbaryl during gestation resulted in no malformations, although abortions were noted in the primate study (Schardein, 1993). 14.4.2.4 Pyrethroids
Another class of insecticide that is widely used is the pyrethroid group of compounds. These are considered to be relatively safe and are perceived to be innocuous because of the origin of the natural pyrethrins from the chrysanthemum family of plants. Earlier studies on the metabolism and toxicity of synthetic pyrethroids (fenothrin, furamethrin, proparthrin, resmethrin,
tetramethrin, and allemethrin) indicate that neither the cis- nor the trans-isomer of chrysanthemumate is teratogenic in rats, mice, and/or rabbits (Miyamoto, 1976). Toxicity studies with decamethrin, a synthetic pyrethroid, found no evidence of teratogenic activity in rats or mice at dose levels that produced marked maternal toxicity (Kavlock et al., 1979). However, numerous studies on the genetic toxicity potential of this group of compounds (cypermethrin and deltamethrin) have demonstrated a wide range of effects, including mitotic/chromosomal abnormalities and the induction of sister chromatid exchanges (Chauhan et al., 1997). In the studies submitted in support of registration for deltamethrin, no significant developmental toxicity was reported in rats; delayed ossification was noted in the high dose in rabbits, aIong with maternal effects (CDPR, 2000). However, a 5% deltamethrin formulation in Wistar-derived albino rats resulted in dose-dependent early embryonic death, retardation of fetal growth, hypoplasia of the lungs, and dilation of the renal pelvis with no skeletal abnormalities (Abd El-Khalik et al., 1993). Significant increases over respective controls were evident for chromosome aberrations, micronuclei, and sperm abnormalities (Bhunya and Pati, 1990). A number of effects of exposure to pyrethroids during early development have been described in rats and mice. Cypermethrin caused an apparent increase in blood-brain barrier permeability in 1O-day-old rat pups after a single or repeated doses of about 15% of the LDso but had no effect on the adult barrier (Gupta et al., 1999). A slightly lower daily dose of 4% of the LDso over postnatal days 10-16 caused an increase in renal D 1 receptor density in rats, which persisted at least until day 90 (Cantalamessa et al., 1998). Similarly low doses of bioallethrin to mice over postnatal days 10-16 decreased muscarinic receptor density in adult mouse neocortex and produced lasting changes in adult behavior (TaIts et al., 1998). These very interesting results have, however, proved impossible to reproduce in other laboratories using only slightly different protocols (unpublished results; D. E. Ray, personal communication), and their applicability to humans is at present uncertain. Similar results with DDT have been reproduced in rats as well as mice (P. Eriksson, personal communication) and in a second strain of mice (unpublished results; D. E. Ray, personal communication). The pyrethroids are also capable of producing gross effects on brain maturation and morphology, but only if given at dose levels that cause reduced body weight in the offspring (Patro et al., 1997), probably via a nonspecific developmental delay due to undernutrition. 14.4.2.5 Insect Growth Regnlators Methoprene A review of the developmental effects of methoprene indicated that a high incidence of multiple malformations was induced in mice, but not in rats (Schardein, 1993). In 1995, middle-school students reported (on the Internet) a high incidence of malformed frogs from a southern Minnesota farm pond. Consequently, increased rates of congenital anomalies in regions of Minnesota associated with pesticide use have heightened awareness of the possible effects of pesticides
14.4 Toxicology Studies
(Garry et aI., 1996). Another group has implicated agricultural contaminants in the hindlimb deformities in frogs from a number of ponds in Quebec (Ouellet et aI., 1997). Additionally, some degradation products of the insect growth regulator S-methoprene have been reported to alter early frog embryo development in the laboratory (La Clair et aI., 1998). However, confirmation of these effects in mammlian species is lacking. The standard teratogenic studies conducted under FIFRA requirements for methoprene do not demonstrate similar results. Furthermore, it is not known if this compound, a juvenile growth hormone agonist, is used in quantities high enough to be a cause for concern. Recent findings linking the limb defects in frogs to a trematode parasite have shifted suspicion away from methoprene, but in the interest of providing the reader with the putative effects of this compound, the aforementioned information is provided. Diflubenzuron Other insect growth regulators such as diflubenzuron (Dimilin, TH 6040; N-[[(4-chlorophenyl)amino]carbonyl]-2,6-difluorobenzamide) have been tested in male and female layer-breed chickens from 1 day of age through a laying cycle at levels of 1, 2.5, 25, and 250 ppm in the feed. Feeding diflubenzuron at levels up to 250 ppm did not affect the characteristics measured such as egg production, egg weight, eggshell weight, fertility, hatchability, and effects on the progeny (Kubena, 1982). 14.4.3 FUNGICIDES Most fungicides tend to produce positive results in standard in vitro microbial mutagenicity tests. This is because the microorganisms used in such test systems are similar to the fungi. However, given the predictive possibility of the mutagenicity tests for teratogenic and carcinogenic potential, there is mounting concern about the developmental toxicity of these compounds. Several fungicides have documented developmental toxicity and details are given later in text (also see Table 14.4). There is evidence to suggest that fluconazole, a bis-triazole antifungal agent, exhibits dose-dependent teratogenic effects; however, it appears to be safe at lower doses (150 mg/day). Ketoconazole, flucytosine, and griseofulvin have been shown to be teratogenic and/or embryotoxic in animals. Iodides have been associated with congenital goiter and should not be used during pregnancy. Benomyl In reproduction studies in the rat, a reduction in epididymal sperm counts in pubertal animals was observed. Postpubertal animals showed a wide variation in susceptibility of sperm counts. Histological exams of testicular tissue showed an increased incidence of diffuse hypo-spermatocytogenesis in pubertal and postpubertal males (Carter et aI., 1984). In the 3000- and 1O,000-ppm males, lower sperm counts were noted. In addition, testicular atrophy and degeneration (4/30 and 29/30 in PI and 9/30 and 21125 in FI 3000- and 1O,OOO-ppm groups respectively) and oligospermia in the epididymides (unilateral and bilateral with 1130 at 3000 ppm and 26/30 at 10,000 ppm
401
in PI, 9/30 and 20125 in FI respectively) were observed. For the reproduction study, the NOEL was 500 ppm in males and 3000 ppm in females (decreased body weights). The NOEL for developmental toxicity was 31.2 mg/kg/day (dose-related reduction in fetal weight, hydrocephaly, microphthalmia, fused ribs, fused vertebrae, and decreased ossification in tail and in vertebral centra) in rats. Findings at the highest dose tested of 125 mg/kg/day included full litter resorptions in 6 of 11 surviving pregnant dams, enlarged lateral ventricles, enlarged renal pelves, and delayed ossification (more widespread than at 62.5 mg/kg/day). Fetotoxicity and teratogenicity findings in the absence of obvious maternal toxicity indicated possible adverse effects. MBC, a metabolite of benomyl, appeared to cause significant effects (postimplantation loss) in rabbits at the midand high-dose levels and resulted in a developmental NOEL of 10 mg/kg/day versus a maternal NOEL of 20 mg/kg/day. However, a teratology study in rabbits exposed to benomyl that was acceptable to the CDPR did not demonstrate possible adverse effects (CDPR, 2000). Dinocap Technical-grade dinocap, a complex-mixture fungicide, has been demonstrated to be teratogenic in the CD-l mouse, causing cleft palate, a dose-related increase in supernumerary ribs, a low frequency of exencephaly, umbilical hernia at high doses, otolith defects, weight deficits in fetuses at term, increased neonatal mortality, abnormal swimming behavior, and torticollis (Rogers et aI., 1986). Neither of the purified isomers, 2,4-dinitro-6-(1-methylheptyl)phenyl crotonate and 2,6-dinitro-4-( 1-methylheptyl)phenyl crotonate, exhibited any developmental toxicity when administered under identical conditions (Rogers et aI., 1987). Similar developmental defects were not seen in the rat and hamster (Rogers et aI., 1988). Hexachlorobenzene Hexachlorobenzene (HCB) is a preemergent fungicide and is ubiquitous in the environment. It has been isolated in the repoductive tract in several species, including humans (Jarrell et aI., 1998; Trapp et aI., 1984). Although HCB was not mutagenic in microbial test systems and was negative in dominant lethal mutation tests, it did cause terata in mice (renal and palate malformations) and in rats (increased incidence of the 14th rib). HCB was also found to be particularly toxic to the developing perinatal animal, transplacentally and via the milk, causing enlarged kidneys, hydronephrosis, hepatomegaly, and possible effects on the immune system (Ecobichon, 1996). It is commonly present in fat because of its lipophilicity and tendency to bioaccumulate (Mes et aI., 1982). The adverse reproductive effects of HCB have been reported in rats, minks, ferrets, and monkeys (Bleavins et aI., 1984; Iatropoulos et aI., 1976). These effects include decreased fertility, fecundity, and impaired cyclicity. In a recent study in Germany, HCB concentration correlated with maternal age (r = 0.249, p < 0.01), with 2.7-fold higher serum levels in offspring of 40-year-old as compared with 20-year-old women, concluding that the neonatal burden depends on maternal age and duration of pregnancy. This reflected the increase in body accumulation with these substances during human life as well as a continuous
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Table 14.4 Developmental and Reproductive Toxicity Profile of Fungicides Dose Chemical
Species
Toxicity profile"
Alkyldithiocarbamic
Rat
Teratogenic
(mg/kg)b
Comments
References As cited in Schardein, 1993
1/20 LDso
acid Benomyl
Mouse
Developmentally toxic
Kavlock et aI., 1982
100
Teratogenic
Selectively toxic
Rat
Fetal mortality
200
Teratogenic
200
Reproduction
Teratogenic po; negative
Kavlock et aI., 1982 As cited in Schardein, 1993
via diet 3-G study
Bis(tri-N -butyltin) oxide
Rabbit
Visceral variations
Mouse
DevelopmentaIIy toxic
23.4
Teratogenic
11.7
Rat Bitertanol
Rat
Captafol
Rat
Developmentally toxic
10
Teratogenic
10
Teratogenic
1/10 LDso 500
Primate
Carbendazim
200 2000
Teratogenic
Primate
37.5 75
Hamster
Teratogenic
Dog
Teratogenic
Rat
Vergieva, 1990 As cited in Schardein, 1993
100
Rat Rabbit
Crofton et aI., 1989
25 Teratogenic
Mouse Captan
As cited in Schardein, 1993
ISO
Rabbit Hamster
Munley and Hurtt, 1996
180
Two species
300
Embryotoxic
100
Teratogenic
200
Robens, 1974 As cited in Schardein, 1993
160
Rabbit Rat
1.5 J.lg/100 g
Rabbit
47J.lg/animal
sc route
Janardhan et aI., 1984
iv route
As cited in Schardein, 1993
BW Cupric acetate
Rat
Developmentally toxic
0.01
Teratogenic
0.01
Toxic to testis Cycloheximide
Mouse
Teratogenic
Rabbit
Developmentally toxic
Rat Cymoxanil
Rabbit
As cited in CDFA, 1987
30
As cited in Schardein, 1993
0.05
As cited in CDFA, 1987
400
As cited in Schardein, 1993
Teratogenic: early resorptions, skeletal variations
CDPR Database, 1999
Dazomet
Rat
Dinocap
Mouse
DevelopmentaIIy toxic
12
Selectively
Rat
DevelopmentaIIy toxic
100
By oral and
Gray et aI., 1996
toxic Rogers et aI., 1988
dermal routes Rabbit
Fetotoxic
48
By oral and
Costlow et aI., 1986
dermal routes Hamster Ethy lenebisisothiocyanate sulfide
DevelopmentaIIy toxic
Rat
12.5 200
Mouse Functional alterations
Rogers et aI., 1988 Chernoff et aI., 1979b
30
(continues)
14.4 Toxicology Studies
403
Table 14.4 (continued)
Dose Chemical
Species
Ethy lthiuram
Rat
Toxicity profilea
(mglkg)b
Comments
References As cited in Schardein, 1993
60
monosulfide Ferbam
Mouse Rat
300 Teratogenic
150
As cited in Schardein, 1993
Reproduction Flusilazole
Rat
Folpet
Mouse
Teratogenic
lOO
Rat
500
Rabbit Hamster
As cited in Schardein, 1993
80 Teratogenic
Primate Hexachlorobenzene
Vergieva, 1990
1/5 LDso
500 Two species
Mouse
Teratogenic
Rat
DevelopmentaIIy toxic
100 10
As cited in Schardein, 1993
DevelopmentaIIy neurotoxic Reproduction Rabbit Imazalil
Mouse
Teratogenic: increased
Rat
Teratogenic: resorptions,
Goldey and Taylor, 1992 10
As cited in Schardein, 1993
10
CDPR database, 1999 DPN#413
skeletal defects NOEL =40
reduced fetal weights Reproduction: decreased
Pup NOEL = 20
litters and litter size Imidazolidinethione
Mouse Rat
800 DevelopmentaIIy toxic
10
Teratogenic
20
Resorption
10
Hamster
Teratogenic
270
Cat
Teratogenic
As cited in Schardein, 1993 Selectively toxic
Rabbit
Guinea
5 100
pig Isoprothiolane
Mouse
Mancozeb
Mouse Rat
DevelopmentaIIy toxic Teratogenic
Rabbit Maneb
Mouse
lOO 1330 1330 80
Increase in variations
375
As cited in Schardein, 1993
Altered behavior Rat
Teratogenic
480
Reproduction Metiram
Rat
Teratogenic: decreased live
80
CDPR Database, 1999
56
CostIow et aI., 1983
litter size Ochthilinone
Rat Rabbit
Embryotoxic
DPN #217 1.5
Phenoxyacetic acid
Mouse
900
Phenylphenol
Mouse
2100
Rat
DevelopmentaIIy toxic
Rabbit Polycarbacin
Rat
Embryotoxic
As cited in Schardein, 1993
150
As cited in Schardein, 1993
250
Zablotny et aI., 1992
610
As cited in Schardein, 1993
Teratogenic (continues)
404
CHAPTER 14
Developmental and Reproductive Toxicology of Pesticides
Table 14.4 (continued) Dose Chemical
Species
Toxicity profilea
(mg/kg)b
Propamocarb
Rat
Teratogenic: skeletal
0.31 mllkg
Comments
CDPR Database, 1999
changes
DPN#50308
Propineb
Rat
Teratogenic
1000
Quintozene
Mouse
Teratogenic
500
Rat
500
Sodium phenylphenol
Mouse
400
Terrazole
Rabbit
Increased resorptions,
References
As cited in Schardein, 1993
IS
CDPR Database, 1999
30
As cited in Schardein, 1993
malformations 2,3,4,6-Tetra
Rat
Delayed ossification
Mouse
Retarded growth
chlorophenol Thiophanate ethyl
200
Reproduction Thiram
Triadimefon
Rabbit
3-G study
Mortality
0.01 LDso
Teratogenic
0.01 LDSO
Hamster
Teratogenic
Rat
Reduced fertility, viability,
250 1800ppm
CDPR Database, 1999
and lactation indices Rat
Decreased body weight gain,
lOO
cleft palate Rabbit
Malformations, variations,
50
delayed ossification 2,4,5-trichlorophenol
Mouse
Tridemorph
Mouse Rat
Teratogenic
900
As cited in Schardein, 1993
245
Merkle et aI., 1984
Developmentally toxic
60.2
Teratogenic
60.2
Selectively toxic
Vinclozolin
Rat
Reproductive malformations
100
Typical of
Gray et aI., 1994
endocrine disruption Zineb
Rat
Ziram
Rat
Teratogenic
2 g/kg 250
As cited in Schardein, 1993 Nakaura et aI., 1984
a Developmental toxicity includes reduced fetal weight, increased embryo/fetal mortality, and/or increased developmental variations, but does not include malformations. Teratogenic is the term used to indicate the presence of malformations. bDose is oral unless stated otherwise; is dose the LOEL wherever effects were observed and the NOEL when there were no effects.
transplacental transfer from mother to fetus during pregnancy (Lackmann et aI., 1999).
decrease fertility. Although it is a reproductive toxin, it does not appear to be teratogenic.
Ethylene Dibromide Ethylene dibromide (EDB; 1,2-dibromoethane), primarily a scavenger of lead compounds in gasoline, has also been used extensively as a fumigant for its chemical and biocidal properties as a soil sterilant and a spot fumigant or control agent in grain milling machinery, in the grain itself, and in fruit and vegetable infestations. In addition to its tumor-causing capabilities in rats and mice, it has been documented to cause changes in sperm morphology in bulls (Amir and Vo1cani, 1965). Spermatids appear to be the target for this compound and it has been shown to affect spermatogenesis in rat bulls and rams and to affect fertility in fowl (Alexeeff et aI., 1990). Human studies indicate that EDB may harm sperm and
Ketoconazole Ketoconazole, an imidazole antifungal agent, can compromise early pregnancy and may affect P450 enzymes of the mammalian steroidogenic system and inhibit progesterone synthesis in the ovary (Cummings et aI., 1997). It is a potential antiandrogenic agent and has displayed antihormonal activities, apparently by inhibiting ovarian hormone synthesis, resulting in delayed delivery and whole litter loss (Gray et aI., 1999a). Maneb Maneb produced fetal hydrocephalus in litters of rats receiving 480 mg/kg/day (Chemoff et aI., 1979a). In FIFRA studies, adverse developmental effects apparently occurred
14.4 Toxicology Studies
because of contamination of maneb with ethylene thiourea (CDPR, 2000). The teratogenicity of a commercial formulation of the fungicide maneb (Maneb 80, containing 80% manganese ethylenebisdithiocarbamate and 20% inert ingredients) was evaluated in chick embryos. It was found to be teratogenic at all concentrations tested (0.5, 1.5, 4.5, and 13.5 g/l maneb aqueous solutions for 30 s), producing mainly unilateral lower limb deformities. No adverse effects on development were noticed after exposure to the inert ingredients (Maci and Arias, 1987). Metam Sodium Adverse effects on the reproductive system were not observed in a rat study; however, histopathology in the nasal cavity demonstrated the following: Bowman's duct hypertrophy with loss of alveolar cells, disorganization/degeneration/atrophy of olfactory epithelium, hyperplasia of olfactory epithelium, and dilatation of ducts of Bowman's glands at the high dose in females of both Fo and Fl generations (0.1 mg/ml in the drinking water). Developmental effects were noted in several studies and included severe malformations (meningocele, anophthalmia, hydrocephaly) at 60 mg/kg in the rat. The rat developmental NOEL was 5 mg/kg, based on a decrease in fetal weights, numerous skeletal developmental delays at 20 mg/kg, and higher levels and delayed ossification in hand and foot bones at 60 mg/kg. In the rabbit, postimplantation loss, early intrauterine deaths, and total litter resorptions were increased at 60 mg/kg with a developmental NOEL of 5 mg/kg/day (based on a decrease in mean live litter size, mean litter and fetal weights, and proportion of males/females at 60 mg/kg). Increases in severe defects (cleft palate and meningocele) at 60 mg/kg and skeletal variations at 20 mg/kg/day and above were also noted. In another study of Himalayan rabbits, embryotoxicity in the form of a statistically significant dose-related increase in dead implantations per pregnant animal at the mid- and high doses was noted. At the high dose, two fetuses with a neural tube closure defect (meningocele+spina bifida) were observed. This study was also acceptable under FIFRA guidelines: The developmental NOEL was 10 mg/kg/day, based on the increase in dead implantations at 30 mg/kg/day; the maternal NOEL was 30 mg/kg/day, based on decreased food consumption at 100 mg/kg/day (CD PR, 2000). Methyl Thiophanate The fungicide methyl thiophanate, widely used to control some of the most common fungal diseases in crops, is metabolized in animals into benzimidazole compounds, including the reproductive toxic ant carbendazim. However, standard toxicological tests did not indicate that methyl thiophanate may cause testicular toxicity and/or embryotoxicity, which are typical effects of many benzimidazoles. In the B6C3F1 mouse, in spite of the high doses administered, none of the testicular parameters examined (sperm head count, specific enzyme activities, histopathology on days 335 postdosing) showed significant alterations as compared to the controls at any time postdosing. Pregnant CD rat dams administered orally the limit dose of 650 mgkg- 1 body weight
405
day-l during the preimplantation (gestational day or GD 2-5) or peri-implantation (GD 6-9) phase showed maternal toxicity, with only marginal reductions of the growth of embryos and adnexa (Traina et aI., 1998). Earlier studies submitted by Atochem North America, Inc. (1985) for methyl thiophanate showed for the rat a teratogenic NOEL of 2500 ppm (125 mg/kg/day), based on the high dose tested (HDT), with a maternal NOEL of 250 ppm (12.5 mg/kg/day), an LEL of 1200 ppm (60 mg/kg/day), and a fetotoxic NOEL of 2500 ppm at HDT (EPA-IRIS database). For mice [study by Pennwalt Corp. (1970a)], the 1000mg/kg/day dose caused a decreased number of implantations; other details were unavailable because fetal examinations did not appear to include soft-tissue examinations. A threegeneration reproduction study in the rat [study by Pennwalt Corp. (1970b)] demonstrated a reproductive NOEL of 160 ppm (8 mg/kg/day) and an LEL of 640 ppm (32 mg/kg/day; HDT), based on reduced litter weights (EPA, IRIS Database). Pentachlorophenol Pentachlorophenol (PCP) is used primarily as a wood preservative. It has been shown to be fetotoxic and teratogenic during early gestation. Commerical PCP is contaminated with chlorinated dioxins and dibenzofurans, tetrachlorophenols, and hydroxychlorodiphenyl ethers (Williams, 1982), which can exert their own effects. Additionally, PCP was reported to be a contaminant in commercial creosote preparations used in wood preservation and may have contributed to its early fetotoxicity. Pentachlorophenol was not teratogenic in rats (Schwetz et aI., 1974). In studies submitted for registration, PCP was found to have adverse effects in the rat developmental toxicity study due to fetal resorptions, decreased fetal weights, ossification delays, and malformations, which findings cannot be assured to result strictly from maternal toxicity (CDPR, 2000). The maternal NOEL was 30 mg/kg/day (body weight and food consumption decrements) and the developmental NOEL was 30 mg/kg/day (increased fetal resorptions, decreased fetal weights, a modest incidence of malformations such as gastroschisis, hydrocephaly, and diaphragmatic hernia judged to be treatment-related, although not statistically significant; significant increase in incidence of dilated pelves delayed ossification in several areas and increased mean numbers of thoracic vertebrae and associated increased incidence of 14th ribs). Another study found a possible adverse effect in that a relatively low developmental toxicity NOEL was observed in the absence of maternal toxicity. A maternal NOEL of 200 ppm (13 mg/kg/day), based on reduced weight gain, clinical signs such as ringed eye, and possible vaginal hemorrhaging, and a developmental NOEL of 60 ppm (4 mg/kg/day), based on reduced fetal weights, misshapen centra, and a possible treatment-related increase in resorptions (significant increase in females with more than two resorptions), were noted (FDA, 1987), indicating a possible adverse effect. Terrazole Decreased live litter size, fetal weight and pup survival (24 h); increased resorptions; and malformations were noted at 45 mg/kg/day in rabbits. The developmental NOEL
406
CHAPTER 14 Developmental and Reproductive Toxicology of Pesticides
was 15 mg/kg/day; Adverse effects were indicated; even though the NOELs for maternal toxicity and developmental toxicity were equivalent, the developmental effects were quite marked (total resorptions equaled 31 in the high dose compared to 7 in the control; 24 h survival of 80% in the high dose compared to 99% in the control). The increased incidence of malformations in the high dose included tail defects, underdeveloped hind limbs, and crossed hind legs (CDPR, 2000).
VincIozoIin This fungicide has the unique claim of being a compound that results in abnormal rodent sex differentiation following exposure during critical stages of life. Effects such as hypospadias, ectopic testes, vaginal pouches, agenesis of the ventral prostate, and nipple retention in male rats were commonly observed (Gray et aI., 1994). In the FIFRA reproduction study, failure of F] males to acquire normal anatomical and functional male characteristics, marked retardation in neonatal growth and survival at dose levels not commensurately toxic to adults, and lenticular degeneration were the principal possible adverse effects. In the FIFRA developmental toxicity study in rats, decreased anogenital distance in males, a finding that was repeated in all the studies conducted and interpreted as feminization of male fetuses, was observed. Similar findings were not noted in the mouse or rabbit (CD PR, 2000). 14.4.4 RODENTICIDES Reviews on the effects of warfarin exposure indicate an uncommon, but strikingly similar pattern of congenital anomalies in children born to women exposed to the compound. The syndrome consists of nasal hypoplasia, stipled epiphyses and growth, retinal-optic atrophy and central nervous system anomalies (Friedman and Polifka, 1994; Schardein, 1993). Although warfarin is used as a rodenticide, it may also be administered to women with heart valve prosthesis. Ginsburg and Barron (1994) recommended not giving warfarin to women between 6 and 12 weeks of gestation.
also noted for commonly used compounds such as piperazine and ivermectin exposures.
Benzamidazole Family of Compounds Variable, but teratogenic potential has been noted for the benzamidazoles as a group. As reviewed by Schardein (1993), skeletal abnormalities were reported in sheep and rats exposed to parbendazole; however, teratogenicity was not reported at comparable or higher doses in hamsters, rabbits, cattle, and swine. Cambendazole was noted to have induced multiple defects in rats and sheep; flubendazole was found to be developmentally toxic in rats, producing multiple malformations; and mebendazole induced malformations in rats with up to 100% incidence in a dosedependent manner but was not teratogenic in rabbits even at high doses. Oxyfenbendazole induced multiple abnormalities in rats and sheep (Schardein, 1993) and swine (Morgan, 1982). Parbendazole has a safety index of over 30 times the recommended dose in healthy animals, but may be teratogenic at doses only slightly higher than the recommended one. It was parbendazole that first alerted scientists to the embryotoxicity of benzamidazoles (Manger, 1991). Thiabendazole is used as a veterinary anthelmintic and fungicide and has variable teratogenic effects in animals. Teratogenicity has been noted in some laboratory animal species (mice and rats), but other studies and reviews have claimed it to be relatively safe (Manger, 1991; Schardein, 1993). In acceptable FIFRA studies submitted for registration, adverse effects were not noted (CD PR, 2000). Another benzamidazole, benomyl (as reported earlier in this chapter), has also demonstrated fetotoxicity and teratogenicity findings in the absence of obvious maternal toxicity, indicating possible adverse effects.
See Tables 14.5 and 14.6 for the toxicity profiles of some miscellaneous pesticides and animal health pesticides.
Antimalarials Chloroquine and its congeners, which are inhibitors of dihydrofolate reductase, and primaquine are known to exert teratogenic effects, but because they are under the category of prescription drugs, the likelihood of exposure during pregnancy is low. Defects noted after quinine exposure include deafness due to auditory nerve hypoplasia, optidisc problems, limb anomalies, and visceral malformations as well as fetal deaths (Schardein, 1993). Rates of spontaneous abortion and birth defects were comparable in pregnant women taking mefloquine, compared with chloroquine-proguanil, or pyrimethamine-sulfadoxine prophylaxis in the first trimester of pregnancy (Phillips-Howard and Wood, 1996). Teratogenic effects for mefloquine were observed in animals but data from humans are lacking (Vanhauwere et aI., 1998).
Anthelmintics Several case reports have been published associating anthelmintic drugs with the induction of birth defects in humans. Ectromelia in infants whose mothers were treated with a tin-based tenifuge; multiple malformations (brain, jaw, ear, limb, and heart defects) subsequent to the intake of mebendazole during the first month of pregnancy; and the incidence of spina bifida and renal anomalies along with hydrocephalus due to quinacrine administration in the first trimester are noted in a review of the data (Schardein, 1993). Negative reports are
ImidacIoprid Imidacloprid is widely used against fleas in dogs and cats and also as an insecticide for use on soil, seed, or foliar treatment in rice, cereal, vegetables, cotton, and turf to control ricehoppers, thrips, termites, turf and soil insects, and some beetle species. In a rat developmental toxicity study, a high percentage of male fetuses and increased incidence of wavy ribs were noted at 94.1 mg/kg/day, indicating a possible adverse effect. The maternal NOEL was 25.9 mg/kg/d, based on decreased body weight gain and reduced food consumption of
14.4.5 ANIMAL HEALTH PRODUCTS, FUMIGANTS, AND MISCELLANEOUS PESTICIDES
14.4 Toxicology Studies
407
Table 14.5 Developmental and Reproductive Toxicity Profile of Miscellaneous Pesticidesa Dose Chemical
Species
Toxicity profileb
Acrylonitrile
Rat
Teratogenic
(mg/kg)"
Comments
As cited in Schardein, 1993
25
Testicular toxin Arsenic trioxide
WiIIhite, 1981 10
Rat
References
Teratogenic by
Stump et aI., 1998a---{;
ip route Reproduction Benzenesulfonic acid
Mouse
Embryo/fetal mortality
Four generations As cited in Schardein, 1993
5.5
hydrazide Diet
Morita et aI., 1981
Benzylbenzoate
Rat
1%
Biphenyl
Rat
500
As cited in Schardein, 1993
Busan 77
Rabbit
125
Drake et aI., 1990
Chlorofebrifugine
Rat
9.3 (po) or
As cited in Schardein, 1993
6ppm (diet) Chloropicrin
Rat
Reduced fetal weight
3.5 ppm
Inhalation
York et aI., 1994
route Rabbit
Reduced fetal weight
2ppm
Inhalation route
Chlorosil
Rat
100
CycIonite
Rat
50
Boikova et aI., 1981 Minor et aI., 1982
Reproduction Rabbit Dikurin
Rat
Diphenylamine
Rat
20 20 Renal lesions
1.5%
Reproduction Ethylene oxide
Mouse
Shepelskaya, 1988 Diet
As cited in Schardein, 1993
2-G study
DevelopmentaIly toxic
150
iv route: also
Teratogenic
150
findings by
As cited in Schardein, 1993 RutJedge and Generoso, 1989
inhalation route Rat
DevelopmentaIIy toxic
10ppm
Reproduction Rabbit
Embryotoxicity
Gliftor
Mouse
Reduced fertility
Guanylthiourea
Rat
Teratogenic
Methyl bromide
Mouse
As cited in Schardein, 1993
route 9
iv route
300
Kimmel et aI., 1982 As cited in Schardein, 1993
33 250
Rat
70 Testicular toxicity in Agenesis of gall bladder
Inhalation
As cited in Schardein, 1993
route
reproductive studies Rabbit
Inhalation
2-G studies 70
Inhalation
Hardin et aI., 198 I
route Methylisothiozolinone
Rat
ScialIi et aI., 1995
15 Reproduction
Rabbit
N-methyl-N-(l-
Mouse
naphtyl)-fluoro
1.5 Fetal growth retardation
<20
Teratogenic
<20
As cited in Schardein, 1993
acetamide DevelopmentalIy toxic
4
PCA
Rat
Peropal
Rat
30
Potassium cyanide
Rat
500ppm
Welsh et aI., 1985 King, 1981 Diet
As cited in Schardein, 1993
(continues)
408
CHAPTER 14
Developmental and Reproductive Toxicology of Pesticides
Table 14.5 (continued) Dose Chemical
Species
Toxicity profileb
Potassium
Rat
Fetotoxic
150
Rabbit
Fetotoxic
76
dimethylthiocarbamate Sulfuryl fluoride
Rat Rabbit
Tetramethyl
Decreased fetal weight
Rabbit
(mg/kg)"
Comments
References Drake et aI., 1989
225 ppm
Inhalation
225 ppm
route
164
Or dermal
Hanley et aI., 1989 As cited in Schardein, 1993
thiodipheny lene phosphorothioate a Includes fumigants, miticides, rodenticides, pediculicides, coccidiostats, and molluscicides.
b Developmental toxicity includes reduced fetal weight, increased embryo/fetal mortality, and/or increased developmental variations, but does not include malformations. Teratogenic is the term used to indicate the presence of malformations. cDose is oral unless stated otherwise; dose is the LOEL wherever effects were observed or the NOEL when there were no effects.
Table 14.6 Animal Health Pesticides (Veterinary Antiparasiticals)a Dose Chemical
Species
Toxicity profileb
Amitraz
Rat
Developmental and behavioral
(mg/kg)"
Comments
References
20
Palermo-Neto et aI., 1994
Developmentally toxic
50
As cited in Schardein, 1993
Teratogenic
50
changes Bromofenofos Cambendazole
Rat Rat Sheep
Dibromochloropropane
Embryotoxic
7.6
Teratogenic
7.6
Embryotoxic
50
Teratogenic
50
Horse
Teratogenic
20
Rat
Developmentally toxic
25
Rabbit
Reproductively toxic
>0.1 ppm
As cited in Schardein, 1993
Drudge et aI., 1983 As cited in Schardein, 1993 Reproductively toxic in man
Diethylcarbamazine
Rat
100
Rabbit
200
2 x use
Dog
level Diethylcarbamazine and
13.2/10
Dog
Fenbendazole
Rat
Flubendazole
Rat
Over two
Rat
120 Teratogenic
40 40
Neonatal mortality and
::::0.4
Developmentally toxic
decreased pup growth
Rodwell et aI., 1987
18/15
Dog
As cited in Schardein, 1993
Muitigeneration
As cited in Schardein, 1993
study 100 flglkg
Primate Ivermectin and pyrantel
As cited in Schardein, 1993
generations generations
oxibendazole
Ivermectin
Over two
As cited in Schardein, 1993 Whorton and Foliart, 1988
In infants One-generation study
Mebendazole
Rat
Embryotoxic
10
Teratogenic
10
As cited in Schardein, 1993
Rabbit
40
As cited in Shepard, 1986
Naftalofos
Rat
15
As cited in Schardein, 1993
Netobimin
Rat
Nitroxynil
Sheep
Teratogenic
71 34
(continues)
14.4 Toxicology Studies
409
Table 14.6 (continued) Dose (mg/kg)C
IChemical
Species
Toxicity profileb
Oxfendazole
Rat
Teratogenic
16
Sheep
Teratogenic
23
Oxibendazole
14
13.5
Morgan, 1982
Mouse
30
As cited in Schardein, 1993
Rat
149 Embryotoxic
10
Teratogenic
10
Abortion
60
Cow Pig Sheep
10 100
Hamster
90 Teratogenic
60
Mouse
400
Rat
400
Piperazine
Pig
PW 16
Rat
Pyrante1
Rat
400 15,000 Skeletal variations Reproductive effects
Rabbit
As cited in Schardein, 1993 3-0 study
Reproduction Rabbit
Ziborov et aI., 1982
440
As cited in Schardein, 1993
440 3,000
Miscarriage
Horse Sodium arsenate
As cited in Schardein, 1993
Pig
Rabbit
Permethrin
References
Cow
Rat Parbendazole
Comments
1,000 12.5
Mouse
120
ip/iv doses
Hood et aI., 1982
teratogenic in other species Terrazo1e
Rabbit
Increased resorptions and
15
CDPR Database, 1999
malformations Thiabendazole
Mouse
Teratogenic
700
Rat
Teratogenic
500
Sheep Tribendimin
Rat
Triclabendazole
Rat
As cited in Schardein, 1993
200 200 Reduced fetal growth
100
a Includes
acaracides and anthelmintics. bDevelopmental toxicity includes reduced fetal weight, increased embryo/fetal mortality, and/or increased developmental variations, but does not include malformations. Teratogenic is the term used to indicate the presence of malformations. cDose is oral unless stated otherwise; dose is the LOEL wherever effects were observed or the NOEL when there were no effects.
the 94.1-mg/kg/day treatment group; the developmental NOEL was 25.9 mg/kg/day, based on increased incidence of wavy ribs in the fetuses of the 94.1-mg/kg/day treatment group (CDPR, 2000).
Methyl Bromide This gas has been used extensively as a fumigant to combat nematodes in strawberries and tomatoes. Alternatives to its use are needed due to its ozone-depleting properties and it is slated for replacement as per the Montreal protocol. Chemically, it is an alkylating agent and capable of neurotoxicity. The compound appears to demonstrate extreme differences among species, dogs being unable to tolerate doses
severalfold lower than those in the rat. Genetic polymorphism for the metabolism of this compound has also been noted. Exposure to methyl bromide in a two-generation reproduction study in Sprague-Dawley rats by inhalation affected fertility (the fertility index decreased from 90.9% in the controls to less than 68% in the 30- and 90-ppm groups) and decreased the body weights of parental and reduced the growth of neonatal rats. Pregnant animals were only exposed 5 days/week (for a total of 14-15 days) during their pregnancy and the pups were not directly exposed until after weaning on postnatal day 28. The parental NOAEL was 3 ppm (reduced fertility). The progeny NOAEL was 3 ppm, based on decreased pup body
410
CHAPTER 14
Developmental and Reproductive Toxicology of Pesticides
weight and reduced organ weights, including reduced F 1 brain weight/reduced width of the cerebral cortex. Data submitted for registration purposes were found to be marginally acceptable, but did not conclusively demonstrate the absence of neurotoxic potential. The developmental study in New Zealand white rabbits demonstrated maternal toxicity at 80 ppm (311 mg/m 3 ) such as reduced body weight and weight gain and clinical signs of central nervous system toxicity. Fetal effects that were not statistically significant but quite rare (i.e., considered biologically significant) were noted. These included omphalocele, hemorrhaging with or without hydrops, and retroesophageal right subclavian artery. Also gall bladder agenesis, fused sternebrae, and decreased fetal body weight were statistically significant at 80 ppm resulting, in a NOAEL for maternal toxicity and developmental effects of 40 ppm (155 mg/m3 ). In rats, a NOEL of 20 ppm for a developmental toxicity study based on delayed skeletal ossification and a maternal NOEL of 70 ppm or more were noted, but the study did not test at a high enough dose level. Reports of other studies via oral exposure in rats and another strain of rabbit demonstrated microphthalmia in rats and some skeletal malformations in the rabbit though not in a doseresponsive pattern (Kaneda et aI., 1998). The reports did not meet FIFRA specifications and historical negative-control data for the rabbit strain employed (Kbl:JW) are not generally available in the open literature. The oral route provides accurate dosage, but because it is more likely to be metabolized prior to reaching the brain than the inhalation route, it may be argued that a higher dosage may be needed to compare the oral route with the inhalation route. Although the pharmacokinetics of transplacental transfer of methyl bromide gas is not available because methyl bromide is known to have neurotoxic potential, and human exposure is most likely via inhalation or skin, inhalation may be the preferred route to detect neurotoxic damage. Hence, adverse effects to development were not observed in the oral studies, but the inhalation studies did demonstrate adverse effects in both reproduction and developmental toxicity studies.
than that for developmental toxicity, and one-quarter of these have been documented to be reproductive toxicants to humans and laboratory animals. There is greater concordance between laboratory animal models and humans for adverse effects on fertility than in the area of developmental effects; for example, male reproductive toxic ants acting on the testes in laboratory animals have the same site of action in humans (Schwetz, 1994). Developmental toxicity in animals, however, does not translate to the same kind of developmental defects in humans (Kimmel et aI., 1993; Schardein, 1989). This lack of concordance, noted for both drugs and other chemicals of commerce, has led to the interpretation that some adverse developmental effect in an animal study is potentially predictive of some adverse developmental effect in humans. Further confounding the issue are accounts that around 6070% of pesticides registered for use have not been adequately tested at the laboratory or clinical level (Mott and Snyder, 1987) and that 25% of the compounds exported from the United States are, in fact, banned or unregistered in the United States (Schardein, 1993). In evaluating global exposure patterns, the data submitted to regulatory agencies become more valuable. Regulatory agencies in the European Community are moving to reduce the number of experimental animals that are being sacrificed for studies. Although such a trend is helping to reduce unnecessary wastage of experimental animals, in vivo data submitted to agencies in the United States often serve as critical studies for a specific compound. Regulations in Japan (Ministry for Agriculture, Forests and Fisheries) have some similarities to U.S. reguations and so duplication of studies can occur, but the benefits of such studies serve to reduce the likelihood of a thalidomide disaster. The aim of testing and regulation is thus to minimize the liability to the manufacturers of chemicals, while assuring that the public will be exposed to a safe set of pesticides. This chapter discussed the majority of pesticides that have been used and reviewed the available data on their reproductive and developmental toxicity. Details for a specific compound or new active ingredient may be obtained from the databases that are accessible to the public at state and federal agencies or from publications.
14.5 CONCLUSIONS Over 4500 chemical tests have been reported using the FDA Segment 11 protocol and more than one-third have come up as positive for developmental toxicity (J. L. Schardein, personal communication). Approximately 25-30 chemicals or families of chemicals are considered human teratogens, based either on a Segment 11 or FIFRA study or on other animal models that have been optimized to detect a toxic effect (Schwetz, 1994). The pilot range-finding studies, in addition to helping to narrow the dose level, have also served to be predictive of the definitive study. Although much is known about the mode of action of some chemicals, the complexity of development suggests that there may be multiple mechanisms of interference with normal development. These mechanisms are not known even for known human teratogens. Similarly, hundreds of chemicals have been tested using protocols for reproductive toxicity; lO-fold less
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CHAPTER
15 Worker Exposure: Methods and Techniques Graham Chester Syngenta
15.1 INTRODUCTION
15.2.1 DERMAL EXPOSURE
Pesticides are biologically active compounds, which may pose a health risk to agricultural workers during or after their use. Operators involved in handling, dispensing, and applying pesticides and postapplication crop reentry workers will be exposed to these compounds through different routes and to varying extents. It is essential both for stewardship and regulatory approval purposes that the possible health risk associated with this exposure is assessed using quantitative information on the toxicological hazard and the amount of exposure. This chapter presents methods by which exposure can be determined. It is not intended to be a complete review of the literature or to provide detailed guidance on how to measure operator exposure or conduct a field exposure study but rather a summary of current "state of the art" principles and methodology involved in measuring exposure of agricultural workers to pesticides. The processes by which such exposure data are used and interpreted are dealt with elsewhere in this book.
In practice, two measurements or estimations are usually made for all work activities associated with the use of pesticides:
1. Potential dermal exposure-the total amount of pesticide coming into contact with the protective clothing, work clothing, and skin. 2. Actual dermal exposure-the amount of pesticide coming into contact with the bare (uncovered) skin and the fraction transferring through protective and work clothing or via seams to the underlying skin, which is therefore available for percutaneous absorption. The biological availability or absorption of a pesticide via the dermal route of exposure is a property of the formulated product and the diluted material and is a separate subject in its own right. Given the significance of the dermal route, precise determinations of percutaneous absorption are key components of the overall assessment of the absorbed dose of the pesticide for risk assessment. 15.2.2 EXPOSURE BY INHALATION
15.2 ROUTES OF EXPOSURE Agricultural workers involved in the use of pesticides and postapplication crop reentry activities may be exposed to pesticides via the skin, by inhalation, or by accidental oral ingestion. Exposures via the first two routes are usually determined separately, but little attention is paid to oral ingestion because it is difficult to estimate or measure. Exposure is usually greatest by the dermal route, although inhalation can be an important route for pesticides that have significant vapor pressures, are applied in confined spaces, or have an application technique which generates a significant proportion of respirable or inhalable particles. Handbook of Pesticide Toxicology Volume 1. Principles
Fundamentally, as far as possible health effects are concerned and setting aside volatile pesticides for the moment, the only spray droplets or particles that pose a potential risk comprise the so-called inhalable or inspirable fraction, which is the mass fraction of airborne particulate capable of entering the respiratory tract via the nose and the mouth, so providing a source of absorption into the body, either from direct inhalation or from subsequent oral absorption. This is considered to be the most important indicator of potential inhalation exposure (ACGIH, 1985; Vincent and Mark, 1987). The inhalable fraction depends on the speed and direction of the air movement, on the rate of breathing, and on other factors. For sampling purposes, inhalable particles can be considered to have a mass median diameter of ~1 00 Jl.m diameter or less. The respirable fraction is the mass
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CHAPTER 15
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fraction of inhaled particles, which penetrates to the unciliated airways. For sampling purposes, respirable particles can be considered to have a diameter between 0 and 15 J.l.m (ISO, 1995). In risk assessment it is common to assume that volatile airborne pesticides are completely retained and absorbed via the respiratory tract, unless there are specific data to demonstrate otherwise. A reasonable default value is 50% retention and absorption of vapors as adopted by the Worker Health and Safety Branch of the California EPA Department of Pesticide Regulation (Thongsinthusak et aI., 1993). Inhalation exposure is usually a small fraction of the total exposure and can, in some cases, be ignored, for example, the mixing and loading of liquid formulations, particularly if a closed loading system is involved. Conditions under which exposure by the inhalation route becomes important usually involve the use of volatile pesticides or of dusts, fumigants, and sprays, especially in enclosed spaces. It should, however, be borne in mind that a higher proportion (up to 75%; Ross et aI., 2000) of the inhaled dose may be retained systemically, compared with the proportion absorbed after dermal exposure, which could be as low as 1% or less of the available dermal dose. 15.2.3 ORAL EXPOSURE
Some of the larger airborne particulates may be trapped in the mouth or nasal passages and subjected to oral ingestion. Some of the exposure, which is measured as inhalation, may indeed be trapped and absorbed in this way. No serious attempts have been made to measure separately the amount of exposure by this route because of the obvious difficulties involved. Biological monitoring takes into account all routes of absorption, but it is usually unable to distinguish between their relative contributions.
15.3 PREVIOUS REVIEWS AND GUIDANCE ON METHODOLOGY Durham and Wolfe (1962), Wolfe (1976), and Davis (1980) were responsible for the earliest reviews of methodology. These reviews and particularly the methodology of Durham and Wolfe (1962) were used to develop the World Health Organization (WHO) standard protocol for the measurement of exposure (WHO, 1975, 1982). The first WHO protocol (1975) advocated the Durham and Wolfe "patch" method to estimate dermal exposure and included a reference to biological monitoring through the use of cholinesterase activity measurement for organophosphorus insecticides. The revised protocol of 1982 proposed an alternative method for the measurement of dermal exposure, the "whole body" sampling technique, and gave an overview of biological monitoring as a means of measuring absorption arising from all routes of exposure. This revised protocol was used to develop the U.S. National Agricultural Chemicals Association (NACA) guidelines for mixerloader-applicator exposure studies (Mull and McCarthy, 1986).
These guidelines placed primary emphasis on the use of the Durham and Wolfe (1962) patch methodology rather than the whole body technique. NACA also published guidelines for conducting field biological monitoring studies as a means of measuring the absorption of pesticides (NACA, 1985). Both the United States Environmental Protection Agency (U.S. EPA, 1987) and NACA guidelines contained detailed reviews of the advantages and limitations of the methods for the measurement of exposure to and absorption of pesticides. The International Group of National Associations of Manufacturers of Agrochemical Products (GIFAP), now known as the Global Crop Protection Federation, published a position paper on general aspects of monitoring studies for the assessment of worker exposure to pesticides (GIFAP, 1990). These guidelines were intended to inform the nonspecialist of the various approaches to exposure/absorption evaluation and their significance. Curry and Iyengar (1992) reviewed and compared the currently available guidelines (published and unpublished) for the evaluation of exposure to individuals using pesticides and those exposed to residues in indoor and outdoor environments. Harmonized exposure and biological monitoring guidelines were proposed in a workshop held in the Netherlands (Chester, 1993; Henderson et aI., 1993). The guidelines agreed on at this workshop were further discussed at a Health CanadaINorth Atlantic Treaty Organization-sponsored "Workshop on Methods of Pesticide Exposure Assessment" held in Canada in 1993, resulting in a draft guidance document to be submitted to the Organization for Economic Co-operation and Development (OECD). A designated peer review group established at the workshop revised the guidance document, which was submitted to the OECD and published as an OECD Guidance Document (OECD,1997). The U.S. EPA meanwhile published a series of test guidelines in 1996 describing their preferred approaches to passive dosimetry and biological monitoring for exposure during indoor and outdoor occupational and residential use of pesticides (U.S. EPA, 1996). The guidance in these OECD and U.S. EPA documents represents the most up to date, harmonized approaches to the assessment of exposure to pesticides and the reader is referred to them for more detailed information.
15.4 DESIGN OF AGRICULTURAL WORKER EXPOSURE STUDIES The purpose of a worker exposure and/or biological monitoring study is to generate data for use in a risk assessment. Good study design is therefore a key consideration in ensuring that relevant and useful exposure data are obtained. In deciding on the basic methodological approach, reference can be made to the tiered approach to exposure and risk evaluation to determine if passive dosimetry will suffice or whether the use of biological monitoring is warranted to give the most accurate
15.4 Design of Agricultural Worker Exposure Studies
determination of the dose absorbed by the worker (Henderson et aI., 1993). Agricultural worker exposure studies can be regarded as being of two types: pre- and reregistration studies and postregistration surveillance studies (OECD, 1997). Studies of the first type involve the test subjects complying fully with the requirements of the product label, in particular, use of protective clothing and equipment, application rates, and clean-up procedures. By ensuring compliance, certain constraints are imposed on the activities of the workers that may influence the amount and variability of their exposures. Studies done according to these criteria are also appropriate for inclusion in generic exposure data bases since they will have been conducted according to a set of standardized principles. Studies of the second type are done primarily in support of product stewardship and postregistration evaluation of actual pesticide use conditions and practices. Adverse effects on the health of workers might be reported, or there might be a possible need to study the extent of compliance with product label precautions and recommendations. Therefore, the study design should take into account the need to measure exposure under the actual conditions of use and would be free of the constraints imposed on pre- or reregistration studies. Other factors that may influence the sampling strategy include possible concern about specific work activities during pesticide use, including the use of nonstandard application equipment. Passive dosimetry enables separate measurements of the respective contributions of these activities to the total exposure. Identification of the differences in the magnitude of exposure attributable to these activities permits the use of different regulatory proposals for reduction of exposure to acceptable levels. An example is the recommended use of additional protective equipment during procedures with greater potential for exposure, for example, handling and mixing of the concentrated formulation. In designing a study consideration should be given to whether passive dosimetry and biological monitoring should be conducted concurrently. Both may be justified to provide data for inclusion in generic data bases, to examine the relationship between exposure and absorption, and to provide a second measurement should one fail. The V.S. EPA, in their latest test guidelines, however, is strongly opposed to this idea (U.S. EPA, 1996). Their concern stems from the reasonable perception that the dermal passive dosimeters, in particular, intercept pesticide as it comes into contact with the worker, thus interfering with the process of dermal contamination and absorption and reducing the biological monitoring. However, if clothing dosimeters are representative of what workers normally use then this disadvantage can be overcome. This matter is discussed further in Section 15.6. Concerning the number of measurements of exposure and/or absorbed dose in the field study, as a general guide, OECD (1997) proposes a minimum of 10 different subjects. The V.S. EPA (1996) requires a minimum of 15 replicated measurements of exposure, not necessarily in different test subjects. Factors to consider in making the decision on the number of subjects are:
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the likely end use of the data; the nature of any identified toxicological endpoint; the required level of statistical confidence; and the overall manageability of the study.
Where feasible, subjects should be randomly selected from the worker population. It is recommended that a sufficient number of measurements be made in different locations to cover the range of use procedures, conditions, and application equipment for which exposure data are required. Variability in exposure can be addressed by increasing the number of subjects rather than repeatedly monitoring the same individuals. Variability between workers is typically greater than that within the same worker (Kromhout et aI., 1993; Rappaport, 1991). In addition, as many sites as possible should be included rather than having subjects use the same equipment under the same conditions. This applies particularly if location is believed to have a significant impact on the variability of the exposure measurements. Certain types of pesticide application procedures render management of the study difficult, such as those involving aerial application. In such cases, repeated monitoring of the same individuals is a possible option, although the limitations of such a choice should be recognized. Should biological monitoring be necessary, a further limitation is introduced in that interindividual variation in metabolism of the pesticide would be less well evaluated. The duration of the study will be prolonged owing to the possible need to collect urine samples over several days for each individual. The consequence is that repeat monitoring cannot commence until urine collection is complete. These difficulties should be considered when using biological monitoring in such circumstances. Ideally, the duration of a single measurement of exposure or absorbed dose should be representative of the typical working day, so that all the work activities that contribute to total exposure, such as equipment repair and clean up, are assessed. This criterion applies particularly to studies involving biological monitoring in which the skin is the predominant, if not only, route of exposure and absorption. For many pesticides that are not well absorbed dermally, absorbed dose data should not be linearly extrapolated on the basis of time or amount of active ingredient used in the same way as passive dosimetry exposure data. The percutaneous absorption of a pesticide depends on the rate at which it is absorbed, the area of skin contaminated, and the duration of skin contact. The rate will increase up to the maximum steady state rate. Once this is achieved, further increases in deposition of pesticide on the skin will have no further impact on the absorption process as it is saturated. Certainly, an increase in deposition can result in increased absorption up to maximum rate, but not in direct proportion (Chester, 1988). In studies involving passive dosimetry and volatile or unstable pesticides, a shorter monitoring period should be considered. This would be based on a consideration of the physicalchemical properties of the compound. The choice of monitoring duration should account for the possibility of dosimeter saturation. Ideally, a single set of dosimeters should be used per
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worker; however, a change of dosimeters during the workday may be necessary if different tasks are to be monitored separately. This can, however, be difficult to manage from the standpoint of practicality. The choice of use pattern (including application equipment) should account for factors such as whether it is the predominant one for the product or a minor one for which no generic data are available to the investigator or regulatory authority to enable a risk assessment. In pre- and reregistration studies the product should be used in the study at a representative recommended rate of application and on the likely maximum area of crop treatable in a working day under local conditions. It should also be applied in accordance with all the label recommendations for use. These principles also apply to studies involving postapplication crop reentry in which exposure should be determined after the shortest permissible reentry period, if known. In postregistration surveillance studies these criteria should not be enforced; this ensures maximum representativeness of actual use conditions and exposure variability. Where the product label recommends the use of protective clothing and/or equipment, these items should be provided to the subjects in pre- and reregistration studies by the supervisory team to ensure standardization. This will benefit the scientific interpretation of the data, because the variable of differing standards of protection by different types and conditions of protective equipment will have been removed. Inclusion of criteria such as these ensures exposure and risk assessment for the product in accordance with the label recommendations for regulatory use. However, efforts should be made to ensure that the recommended protective clothing and equipment are practical and realistic to use under local conditions. In postregistration surveillance studies, the study team should not mandate use of protective clothing and equipment, although ethical or legal viewpoints on product label recommendations in this respect should be considered.
15.5 TEST SUBJECTS Agricultural workers should be the test subjects in a field study, rather than inexperienced volunteers. If this is not possible, use of nonprofessional personnel may have to be considered, provided that they are given the requisite training in the handling and use of the pesticide and equipment. The disadvantage of this choice is that the subjects would not necessarily be representative of the worker population. Males or females may be considered for inclusion in passive dosimetry studies. However, in studies involving biological monitoring, the decision is more difficult owing to the potential impact on the interpretation of the metabolite excretion data due to possible sex differences in metabolism and kinetics. If it is likely that the product will be used by both sexes then it is important to know if there are differences. All subjects should be asked to provide written, informed consent to participate in a study after they are provided with the requisite information on the pesticide. Potential subjects must
be informed that they are free to withdraw from the study at any time. Subjects should be screened for any preexisting medical conditions that may be affected by use of the pesticide, depending upon its toxicological profile. Depending upon the circumstances and local custom, it may be appropriate to provide the subjects with information on their individual results.
15.6 METHODS FOR MEASURING EXPOSURE (PASSIVE DOSIMETRY) 15.6.1 PATCH METHOD FOR DERMAL EXPOSURE
In this method, the potential contamination of the workers' skin and clothing is measured using a variable number of absorbent cloth or paper patches, attached to body regions inside and outside clothing. The surface area covered by the patches represents less than lO% of the total body surface area. After a defined or measured period of exposure, the patches are removed and analyzed for pesticide content. The quantity of a pesticide on a patch of known area is then related to the area of limb or other body part on the assumption that deposition is uniform over the body parts. Body part surface areas can be obtained from standard reference texts and exposure guidance documents, the most recent of which are those of the V.S. EPA (1996,1999) and the OECD (1997). The assumption of uniform deposition is perhaps the principal disadvantage of the patch technique. This is illustrated by the extrapolation of the value given by half the limit of quantification to the total body part; this may give a substantial underor overestimate of exposure. The principal disadvantage can be mitigated to a certain extent by increasing the number of patches located on body parts likely to receive significant exposure. Individual body part exposure values are then added to give a total potential exposure expressed in mg/hour, mg/day, or mg/kg product handled or applied. In the WHO protocol fewer patches are recommended, representing only 3% of the body area. In both cases standard body part surface areas are used in correcting individual patch values (WHO, 1982; U.S. EPA, 1996). According to the WHO protocol, only the pesticide contacting the normally unclothed area of skin, for example, head, neck, hands, and forearms, is used to calculate actual exposure. In temperate climates about lO% of the total body area is normally unprotected during use of pesticides and other agricultural activities. Normal work clothing, such as cotton trousers or shirts, is absorbent and may retain and allow penetration of a proportion of the pesticide contamination. Therefore, it is still necessary to estimate exposure to the covered areas of the body. Consequently the amount of pesticide penetration is often measured using patches attached beneath the clothing. Without such measurements, an estimate of penetration of normal clothing must be used. However, penetration of clothing by pesticides is a highly variable process, which is influenced by factors such as the type of formulation (liquid or solid) and the amount or volume of deposition on the clothing, dampness of the clothing, pesticide vapor pressure,
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the location of deposition (e.g., seams), and the type of fabric. Further uncertainties are introduced by the method of sampling for clothing penetration. The patch method may give significant under- or overestimates of exposure, depending on whether the patches have captured the nonuniform, random deposition of concentrate splashes or spray droplets. This limitation applies equally to crop reentry procedures, such as harvesting, where contact with the crop is not a uniform process. Despite the readily apparent limitations, the patch method remains a useful method for exposure evaluation.
acceptability applicable to all pesticides, the surrogate compound should not significantly alter the physical properties of the formulation or spray mixture. The key question that determines the utility of a tracer or dye is whether it affects, is retained by, or penetrates clothing similar to or in parallel with the pesticide of interest. The technique can only be used in exposure studies and not in biological monitoring (although it can be used with concurrent biological monitoring of a pesticide). Its greatest utility probably lies in substituting for pesticides that are particularly unstable during the sampling and analytical phases.
15.6.2 USE OF FLUORESCENT TRACERS AND VISIBLE DYES: QUANTIFICATION BY ANALYSIS OR VIDEO IMAGING
15.6.3 WHOLE BODY METHOD
Dermal exposure can be quantified directly by measuring deposition of fluorescent materials or visible dyes on the clothing and/or skin. The fluorescent tracer or dye can be substituted and extracted from passive dosimeters and analyzed in the same way as the pesticide. By adjustment for concentration differences, an estimate of exposure to the pesticide can be obtained, similar to extrapolation from one monitored pesticide to another used in the same way. A recent technical advance has been the development of a video imaging/fluorescent tracer technique (Fenske et aI., 1986a, b; Fenske, 1990). This method involves the incorporation of a fluorescent tracer in a pesticide formulation and subsequent visual and quantitative analysis using a video imaging technique. It reveals non uniform patterns of exposure that escape detection by the patch method. It also demonstrates that exposure can occur beneath protective clothing. An important advantage of this technique is that the skin serves as a collection medium rather than dosimeter patches or clothing. The main limitations of this method are the assumptions that the relative transfer of the tracer and the pesticide in the field and their permeation of the clothing are equivalent. However, these assumptions are analogous to those involving use of generic exposure databases. That is, the exposure to a pesticide measured under a given set of conditions is assumed to represent the exposure associated with a second pesticide under the same conditions. Ancillary studies can assess possible differences in the relative transfer of the tracer and the pesticide. The techniques are particularly useful for the training of operators by demonstrating the extent of their contamination, thus enabling modification of their working practices to reduce exposure. Roff (1994) developed the technique a stage further by using a dodecahedrallighting system to illuminate the contaminated worker. Measurement errors inherent in the Fenske technique caused by body surface morphology are apparently reduced. This rather large piece of equipment has been used under field conditions. If a tracer compound or visible dye is chosen, its performance and suitability as a surrogate should be validated before the field study. Apart from the usual criteria of quality control
The whole body method came into use during the late 1970s1 early 1980s (WHO, 1982; Abbot et aI., 1987). The method involves the use of clothing, usually two layers of cotton or cottonlblend material that act as the pesticide collection media. The outer layer of clothing should be representative of what the workers might wear under normal circumstances. The inner layer, usually a set of combination "long johns," represents the skin. This method overcame one of the inherent problems of the patch method, i.e., the assumption of uniformity of pesticide deposition on the skin and clothing. Exposure of the head is assessed by use of a hood or hat, preferably made of the same material. A face wipe technique can also be used, in which the skin of the face and anterior and posterior neck is wiped with cotton swabs containing a suitable detergent to remove the pesticide contaminant. Any additional protective clothing and equipment recommended for the product under study are worn over the sampling clothing, thus enabling an evaluation of their protection efficiency. The use of the whole body method overcomes the perceived problem of non uniformity of deposition. Furthermore, extrapolation from small target areas to larger body regions is not necessary. For these reasons, the method is believed to give a more accurate estimate of potential and actual dermal exposure. The whole body method can be adapted for concurrent use with biological monitoring by use of work clothing as dermal dosimeters, which represents what the workers would normally wear under the prevailing conditions. This is contrary to the view expressed by the U.S. EPA (1996). Whereas the patch and standard whole body methods place sampling media between the pesticide and the clothing or skin thus acting as a barrier interfering with the normal process of skin contamination and percutaneous absorption-the U.S. EPA's main concern-the advantage of this method is that the capture, retention, and penetration properties of the normal work clothing are mimicked as closely as possible. It is important, therefore, to have an understanding of the range of normal work clothing worn by the worker population under study. Clothing for sampling should be selected cautiously, using the minimum that might be worn under the prevailing conditions. Therefore, the assessment of residual clothing contamination and transfer to the skin
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beneath the clothing is as realistic as possible. This method is particularly relevant for North European and North American temperate countries where the typical work clothing consists of a "T" shirt, long sleeved shirt, socks and long trousers, and/or coveralls. Actual exposure of the skin beneath the clothing can be estimated by detemining the ratio of outer to inner clothing penetration or transfer of the pesticide. An obvious limitation is that the permeation and transfer properties of the outer and inner clothing are assumed to be the same. For analytical considerations, it may be necessary to use noncolored, white materials such as cotton or cotton/polyester mixtures. As in the standard whole body method the clothing is sectioned into individual body parts and analyzed separately to determine the regional distribution of total potential and actual exposure. Few attempts have been made to validate methods for the monitoring of dermal exposure, for example, by using biological monitoring to compare the derivations of absorbed dose. Until this is done, all methods should be viewed as providing only an approximate estimate of the dermal exposure. Dermal exposure methods will still be needed, as biological monitoring cannot be applied to all pesticides. Exposure method development is therefore a continuing need. 15.6.4 HAND EXPOSURE
Measurement of hand exposure is one of the most important aspects of a study to monitor worker exposure. The contribution of the hands to total exposure is well documented and was originally recognized in the seminal works of Batchelor and Walker (1954) and Durham and Wolfe (1962). The U.S. EPA (1996) reviewed the literature on studies that had included hand exposure measurements and concluded that its contribution to total exposure ranged from around 40 to 98%, depending upon the application method. The methods for measuring hand exposure include using lightweight absorbent gloves or sections cut from gloves and swabbing or rinsing the hands in various solvents, for example, 95% ethanol (U.S. EPA, 1996). Mild detergent solutions can be used in the hand-wash technique, for example, Aerosol OT. All these methods have advantages and limitations and it is difficult to evaluate the accuracy of any procedure. This is the reason for one of the major issues with developing generic hand exposure data bases-lack of standardization in measurement technique. It is feasible to evaluate the efficiency of recovery of a pesticide for a hand-wash or hand-rinse method under laboratory conditions using human subjects. This has been investigated by Fenske and Lu (1994) for the insecticide, chlorpyrifos, using a solvent hand-rinse method. Their findings suggest that exposure to pesticides such as chlorpyrifos that are well adsorbed to the skin cannot be estimated accurately by the hand-rinse method. However, Geno et al. (1996) showed for chlorpyrifos and pyrethrin that hand wipes with isopropanol removed in excess of 90% of the material applied to hands. Fenske et al. (1999) compared the hand exposures of orchard apple thinners
to azinphos-methyl using three methods: glove, hand-wash, and wipe. Hand exposure estimates derived from the three methods differed significantly. Based upon the hand-wash measurements and the laboratory recovery/efficiency study, it was concluded that the glove method gave a 2.4-fold overestimate whereas the wipe method gave a lO-fold underestimate. The authors also concluded that methods should be validated and standardized to enable the development of more accurate hand exposure estimates. The generally held view is that the use of gloves results in a significant overestimation of total dermal exposure, owing to the retention of more of the pesticide than would otherwise be retained by the skin. Gloves also contain foreign materials such as sizing, which may be coextracted with the pesticide. At low levels of contamination this may cause analytical difficulties. However, glove contamination with dirt and grease arising from the worker's activities are a more likely cause of analytical problems. Washing the hands with a solvent such as ethanol might cause skin damage or disrupt skin barrier function and enhance percutaneous absorption of the pesticide. In the many studies that have used this technique there has been little evidence that these effects occurred. In studies involving concurrent passive dosimetry and biological monitoring, a hand-washing procedure involving standardized detergent and water can be used. This procedure is identical to that described earlier as the hand-wash method, except that the measurement is taken only when the workers would normally wash their hands. The basis is that when the total absorbed dose is determined with biological monitoring there should be no interference with the normal process of dermal contamination and percutaneous absorption. Therefore the use of gloves or a solvent washing or rinsing technique is inappropriate, because these methods would retain pesticide otherwise available for absorption or potentially disrupt the barrier function of the skin. Inevitably there is a loss of standardization of the intervals at which samples are taken. However, it does give some information on the extent of hand exposure that might be of value in overall data interpretation. 15.6.5 INHALATION EXPOSURE
Exposure by inhalation is usually a minor route of absorption in comparison with the dermal route. There are exceptions to this, for example, when dusts, fine aerosols, and fumigants are applied or when materials are applied indoors. The U.S. EPA (1996) reviewed several exposure studies and found that the inhalation route contributed negligible amounts (nondetectable) to about 9% of total exposure. In most cases, the contribution was less than 1%. The extent of the contribution depends upon the method of application, whether used outdoors or indoors and on factors such as the volatility of the pesticide. Significantly, for pesticides that are poorly absorbed via the skin, the inhalation route can become the most important route of absorption.
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The important reviews of methodology for field monitor- particulates and vapor can be achieved by mounting the filter ing of airborne pesticides were provided by Van Dyk and sampling head in front of the vapor trap in a "sampling train." Visweswariah (1975) and Lewis (1976). The former reviewed This train allows retention of any vapor stripped off the filter the sampling media available for collection of pesticides but on the resin. The material on the filter can be analyzed both with particular emphasis on static environmental sampling gravimetrically and/or chemically and an estimate made of the rather than personal sampling. pesticide content of the particulate sample. If use of such a A personal air sampling method is the most appropriate for sampling train is needed, laboratory validation of the sampling the determination of potential inhalation exposure of workers. efficacy, particularly of the adsorbent resin, is necessary owing Several techniques are available such as gauze pads in place of to the possibility of stripping material from the resin by the relfilters in respirators for agricultural use, pioneered by Durham atively high flow rate of 2 lImin. The measurement of the inhalable fraction with use of adsorand Wolfe (1962), midget impingers, solid adsorbents, and filter cassettes attached to battery-powered personal sampling bent resins for the vapor phase is recommended as the method pumps. A personal sampling technique involving sampling de- of first choice. vices located in the breathing zone and sampling pumps is preferred for reasons of practicability and representativeness. 15.7 METHODS TO MEASURE THE Breathing rates for the calculation of inhalation exposure from airborne concentration data can be obtained from standard refABSORBED DOSE erence texts such as the D.S. EPA's Exposure Factors Handbook (1999). 15.7.1 BIOLOGICAL MONITORING The advantage of using the modified respirator is that the subject produces the airflow so that breathing rate and total vol- Biological monitoring of pesticide workers was first used as a ume of air inhaled do not have to be estimated. However, the means of assessing health effects or modification of biochemgauze pads must be capable of trapping the pesticide efficiently. ical parameters as a consequence of exposure to organophosThe respirator must also fit properly to the face. Perhaps the phorus compounds by measurement of plasma cholinesterase main disadvantage of this technique is that the subject must levels (for example, Peoples and Knaak, 1982). This type of aswear the respirator for the duration of the monitoring period, sessment, used in the chemical industry for many years, can be which ideally should be a complete working day, which may termed biological effect monitoring and must be distinguished cause discomfort. from the type of monitoring that determines the absorption of Midget impingers, which traditionally have used ethylene chemicals by measuring the chemical or its metabolites in body glycol as the trapping medium, have a long history of use in fluids, usually urine, blood, or exhaled breath. measuring agricultural worker inhalation exposure. The techAnalysis of body fluids and excreta, usually urine, for parent nique suffers from the major drawbacks of spillage of the trap- compound or metabolites can provide both a qualitative and a ping liquid and inefficient trapping and retention of some pes- quantitative measurement of absorbed dose for pesticides that ticides. Microimpingers have been developed, which overcome lend themselves to this form of monitoring. The technique has the first of these drawbacks. a distinct advantage over passive dosimetry because it evaluates Personal air samplers allow the use of respirators or dust actual, rather then potential, absorption. It integrates absorption masks for protection, if required by the product label. How- from all routes of exposure: dermal, inhalation, and primary and ever, they do not measure the true exposure of workers wearing secondary oral ingestion. However, it is difficult to differentiate respiratory protection. The choice of sampling medium is de- the contributions to the absorbed dose from different aspects termined by the nature of the pesticide. A filter cassette or of the work procedures or to distinguish between the relative sampling head should be used for spray particulates and a solid contributions of the different routes of exposure to the total adsorbent material for volatile compounds. absorbed dose. Van Heemstra-Lequin and Van Sittert (1986), The inhalable fraction (all material capable of being drawn Wang et al. (1989), and Henderson et al. (1993) reviewed biinto the nose and mouth) is the most biologically relevant frac- ological monitoring in the context of pesticides. The OECD tion to measure. An example of a suitable device is the Institute (1997) provided detailed guidance on how to conduct biologof Occupational Medicine personal sampling head designed ical monitoring studies. specifically to collect this fraction (Vincent and Mark, 1987). Early biological monitoring studies on pesticides were able For use of this device, a sampling flow rate of 2 lImin is a spe- to demonstrate absorption without quantifying the amount of cific requirement. There is now a commercially available device absorption. For example, Durham and Wolfe (1962) measured manufactured by SKC that collects, as it separate fractions, the p-nitrophenol, a metabolite of parathion, in the urine of workinhalable and respirable components. ers; Swan (1969) measured paraquat in the urine of backpack Examples of suitable adsorbent materials for some volatile spray operators. The absorbed dose of a pesticide can only compounds are activated charcoal and Tenax and XAD-2 resins be quantified accurately if the metabolism and pharmacokimounted in stainless steel or glass tubes. The choice of ma- netics of the compound are understood, ideally from human terial should be determined by analytical retention (trapping studies. Woollen (1993) reviewed the specific requirements for efficiency) and extractability studies. Concurrent sampling for this type of biological monitoring study. Studies that have met
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these requirements include those on the herbicide fluazifopbutyl (Chester and Hart, 1986) and 2,4-dichlorophenoxyacetic acid amine (Grover et aI., 1986; Ritter and Franklin, 1989). Data from human dosing studies facilitate the design of a field sampling strategy and secondly define the body fluid matrix of choice. Ideally, urine is the matrix of choice as its collection is noninvasive and the collection of 24-hour urine samples is practicable. Complete 24-hour urine collections are essential, and this can be checked in a number of ways, for example, by measuring the concentration of creatinine. Substantially incomplete collections are readily apparent and these samples are either excluded or an allowance is made, for example, by use of a correction factor based upon the average daily urine volume for the individual concerned (Woollen, 1993). Specific gravity and osmalarity are alternative means of checking for completeness of urine collection (Allesio et aI., 1985). Unstable or highly volatile pesticides are not good candidates for passive dosimetry, despite the efforts to accurately assess field, storage, and transit losses. Biological monitoring should be considered for these pesticides and may be the only means of obtaining adequate quantitative data from which the absorbed dose can be derived. An example of the successful use of biomonitoring to estimate exposure occurring primarily by the inhalation route is urinary monitoring for a metabolite of the fumigant 1,3-dichloropropene (Osterloh et al., 1989; Van Welie et aI., 1991). If a human metabolism study were impracticable, then animal metabolism data might be used, if metabolism and excretion kinetics are similar in several animal species, then it could be assumed that humans will metabolize and excrete the compound in a similar manner. This carries a degree of uncertainty. There are examples, described in detail by Woollen (1993), that demonstrate the limitations ofthis approach. Apart from interspecies differences in metabolism, there is the possibility of dose-dependent differences, which might necessitate metabolism studies in animals and humans at doses similar to the anticipated worker exposures. Pesticides that are extensively metabolized to a large number of metabolites are not good candidates for biological monitoring. The absorbed dose cannot be determined accurately using data on a minor metabolite, particularly if there is wide interindividual variation in the proportion of the parent compound excreted as this metabolite. However, a minor metabolite might provide some useful information as a "biological indicator" in the absence of more abundant metabolites. Overall, it can be concluded that if the requisite human metabolism data are available for a pesticide, biological monitoring provides the most accurate means of estimating the absorbed dose for quantitative risk assessment.
REFERENCES Abbott, 1. M., Bonsall, J. L., Chester, G., Hart, T. B., and Turnbull, G. J. (1987). Worker exposure to a herbicide applied with ground sprayers in the United Kingdom. Am. Ind. Hyg. Assoc. 1. 48,167-175.
Alessio, L., Bolin, A., Dell'Orto, A., Toffoletto, E, and Ghezzio, I. (1985). Reliability of urinary creatinine as a parameter used to adjust values of urinary biological indicators. Int. Arch. Occup. Environ. Health 55, 99-106. American Conference of Government Industrial Hygienists (ACGIH)fTechnical Committee on Air Sampling Procedure (1985). "Particle Size Selective Sampling in the Workplace." ACGIH, Cincinnati. Batchelor, G. S., and Walker, K. C. (1954). Health hazards involved in the use of parathion in fruit orchards of North Central Washington. Am. Med. Assoc. Arch. Ind. Hyg. 10,522-529. Chester, G. (1988). Pesticide applicator exposure-towards a predictive model for the assessment of hazard. Aspects Appl. Bio!. 18,331-343. Chester, G. (1993). Evaluation of agricultural worker exposure to, and absorption of pesticides. Ann. Occup. Hyg. 37, 509-523. Chester, G., and Hart, T. B. (1986). Biological monitoring of a herbicide applied through backpack and vehicle sprayers. Toxicol. Lett. 33, 137-149. Curry, P., and Iyengar, S. (1992). Comparison of exposure assessment guidelines for pesticides. Rev. Environ. Contam. Toxico!. 129, 79-93. Davis, J. E. (1980). Minimizing occupational exposure to pesticides: personal monitoring. Residue Rev. 75, 35-50. Durham, W. E, and Wolfe, H.T. (1962). Measurement of the exposure of workers to pesticides. Bull. WHO 26, 75-91. Fenske, R. A. (1990). Non-uniform dermal deposition patterns during occupational exposure to pesticides. Arch. Environ. Contam. Toxicol. 19,332-337. Fenske, R. A., Leffingwell, J. T., and Spear, R. C. (1986a). A video imaging technique for assessing dermal exposure-I. Instrument design and testing. Am. Ind. Hyg. Assoc. 1. 47, 764-770. Fenske, R. A., and Lu, C. (1994). Determination of handwash removal efficiency; incomplete removal of the pesticide chlorpyrifos from skin by standard handwash techniques. Am. Ind. Hyg. Assoc. 1. 55, 425-432. Fenske, R. A., Simcox, N. J., Camp, J. E., and Hines, C. J. (1999). Comparison of three methods for assessment of hand exposure to azinphos-methyl (Guthion) during apple thinning. App. Occup. Environ. Hyg. 14,618-623. Fenske, R. A., Wong, S. M., Leffingwell, J. T., and Spear, R. C. (1986b). A video imaging technique for assessing dermal exposure-H. Fluorescent tracer testing. Am. Ind. Hyg. Assoc. 1. 47, 771-775. Geno, P. w., Camann, D. E., Harding, H. J., Villabos, K., and Lewis, R. G. (1996). Handwipe sampling and analysis procedure for the measurement of dermal contact with pesticides. Arch. Environ. Contam. Toxicol. 30, 132138. Groupement International des Associations NationaIes de Fabricants de Produits Agrochemiques (GIFAP). (1990). "Monitoring Studies in the Assessment of Field Worker Exposure to Pesticides," Technical Monograph No. 14. GIFAP, Brussels. Grover, R., Franklin, C. A., Muir, N. I., Cessna, A. J., and Riedel, D. (1986). Dermal exposure and urinary metabolite excretion in farmers repeatedly exposed to 2,4-D amine. Toxicol. Lett. 33, 73-83. Henderson, P. Th., Brouwer, D. H., Opdam, J. J. G., Stevenson, H., and Stouten, J. Th. J. (1993). Proceedings of workshop on: risk assessment for worker exposure to agricultural pesticides. Ann. Occup. Hyg. 37, 499-507. International Organization for Standardization (ISO). (1995). "Air QualityParticle Size Fraction Definitions for Health-Related Sampling." ISO 7708: 1995(E). Kromhout, H., Symanski, E., and Rappaport, S. M. (1993). A comprehensive evaluation of within and between worker components of occupational exposure to chemical agents. Ann. Occup. Hyg. 37, 253-270. Lewis, R. G. (1976). Sampling and analysis of airborne pesticides. In "Air Pollution from Pesticides and Agricultural Processes" (R. E. Lee, Jr., ed.). CRC Press, Cleveland. Mull, R., and McCarthy, J. E (1986). Guidelines for conducting mixer-Ioaderapplicator studies. Vet. Hum. Toxicol. 28, 328-336. National Agricultural Chemicals Association (NACA). (1985). "Guidelines for Conducting Biological Monitoring-Applicator Exposure Studies." NACA, Washington, DC.
References
Osterloh, J. D., Wang, R., Schneider, F., and Maddy, K. (1989). Biological monitoring of dichloropropene: air concentrations, urinary metabolite, and renal enzyme excretion. Arch. Environ. Health 44, 207-213. OECD (1997). Guidance Document for the Conduct of Studies of Occupational Exposure to Pesticides During Agricultural Application. OECD Environmental Health and Safety Publications Series on Testing and Assessment No. 9. Environment Directorate, OECD Paris. Peoples, S. A., and Knaak, J. (1982). Monitoring pesticide blood cholinesterase and analysing blood and urine for pesticides and their metabolites. In "Pesticide Residues and Exposure" (J. R. Plimmer, ed.), ACS Symposium Series, Vol. 182, pp. 41-57. Am. Chem. Soc., Washington, DC. Rappaport, S. M. (1991). Assessment of long-term exposures to toxic substances in air. Ann. Occup. Hyg. 35, 61-121. Ritter, L., and Franklin, C. A. (1989). Use of biological monitoring in the regulatory process. In "Biological Monitoring for Pesticide Exposure" (R. G. M. Wang, C. A. Franklin, R. C. Honeycutt, and J. C. Reinert, eds.), ACS Symposium Series, Vol. 382, pp. 354-367. Am. Chem. Soc., Washington, DC. Roff, M. W. (1994). A novel lighting system for the measurement of dermal exposure using a fluorescent dye and an image processor. Ann. Occup. Hyg. 38,903-919. Ross, J. H., Driver, J. H., Cochran, R. C., Thongsinthusak, T., and Krieger, R. 1. (2000). Could pesticide toxicology studies be more relevant to occupational risk assessment? 1. Occup. Hyg. 45(Suppll), 5-17. Swan, A. A. B. (1969). Exposure of spray workers to paraquat. Br. 1. Ind. Med. 26,322-329. Thongsinthusak, T., Ross, J. H., and Meinders, D. (1993). "Guidance for the Preparation of Human Pesticide Exposure Documents, HS-1612, May 4." California Environmental Protection Agency, Worker Health and Safety Branch. U.S. Environmental Protection Agency. (1987). Pesticide Assessment Guidelines, Subdivision U, Applicator Exposure Monitoring. U.S. EPA, Washington, DC.
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U.S. Environmental Protection Agency. (1996). Occupational and Residential Exposure Test Guidelines, OPPTS 875.1000, EPA 712-C-96-261. U.S. EPA, Washington, DC. U.S. Environmental Protection Agency. (1999). Exposure Factors Handbook, EPN600/C-99/001, February. U.S. EPA, Office of Research and Development, Washington, DC. Van Dyk, L. P., and Visweswariah, K. (1975). Pesticides in air: sampling methods. Residue Rev. 55, 91-134. Van Heemstra-Lequin, E. A. H., and Van Sittert, N. J., eds. (1986). Biological monitoring of workers manufacturing, formulating and applying pesticides. Toxicol. Lett. 33, 1-236. Van Welie, R. T., van Duyn, P., Brouwer, D. H., van Hemmen, J. J., Brouwer, E. J., and Vermuelen, N. P. (1991). Inhalation exposure to 1,3dichloropropene in the Dutch flower-bulb culture. Part n. Biological monitoring by measurement of urinary excretion of two mercapturic acid metabolites. Arch. Environ. Contam. Toxicol. 20, 6-12. Vincent, J. H., and Mark, D. (1987). Comparison of criteria for defining inspirable aerosol and the development of appropriate samplers. Am. Ind. Hyg. Assoc. 1. 48, 451-457. Wang, R. G. M., Franklin, C. A., Honeycutt, R. c., and Reinert, J. C. (1989). "Biological Monitoring for Pesticide Exposure: Measurement, Estimation and Risk Reduction," ACS Symposium Series, Vol. 382. Am. Chem. Soc., Washington, DC. Wolfe, H. R. (1976). Field exposure to airborne pesticides. In: "Air Pollution from Pesticides and Agricultural Processes" (R. E. Lee, Jr. ed.). CRC Press, Cleveland, Ohio. Woollen, B. H. (1993) Biological monitoring for pesticide absorption. Ann. Occup. Hyg. 37, 525-540. World Health Organization. (1975). "Survey of Exposure to Organophosphorus Pesticides in Agriculture," Standard Protocol, VBCI75.9. WHO, Geneva. World Health Organization. (1982). "Field Surveys of Exposure to Pesticides," Standard Protocol, VBC/82.1. WHO, Geneva.
CHAPTER
16 Residential Exposure Assessment: An Overview Jeffrey H. Driver, John H. Ross, Muhilan D. Pandian infoscientific.com & risksciences.net
Jeffrey B. Evans D.S. Environmental Protection Agency, Office of Pesticide Programs
Gary K. Whitmyre risksciences, LLC
16.1 INTRODUCTION Following the use of products in and around the home, postapplication chemical exposures to consumers may occur in a variety of microenvironments that correspond to the daily activities in which adults and children engage. These activity patterns may place individuals in contact with a variety of chemicals including pesticides (e.g., dislodgeable foliar residue exposures during gardening, lawn chemical exposures after reentry onto treated turf; and chemical emissions from treated surfaces inside the residence). To understand the potential health significance of these exposures it is necessary to characterize their sources and estimate their magnitude. In response to these needs, efforts have been undertaken to develop methodologies for quantifying pesticide and other chemical exposures in soil, air, food, and water (Cal-EPA, 1994; McKone, 1991, 1993; Ott, 1985; Thompson et aI., 1984; Vaccaro et al., 1996; Wallace, 1987, 1989, 1990, 1991; Wall ace et aI., 1982, 1984, 1985, 1986, 1987a, b, c, 1988, 1989, 1991a, b). The U.S. Environmental Protection Agency (U.S. EPA), for example, in response to the the Food Quality Protection Act of 1996 (FQPA; http://www.epa.gov/docs/oppfeadsllfqpa),has been revising exposure monitoring guidelines that emphasize nonoccupational, residential exposure to pesticides; these guidelines are referred to as "Series 875, Occupational and Residential Exposure Test Guidelines Group B: Post-Application Monitoring Test Guidelines" (http://www.epa.gov/docs/opptsfrs/OPPTS_ Harmonized/87 5_Occupational_and_ResidentiaLExposure_ TesCGuidelines; Whitford et al., 1999). The series 875 guidelines provide information and protocols relevant for persons required to submit postapplication exposure data under 40 CFR 158.390 (Fig. 16.1). Generally, these data are required under the Handbook of Pesticide Toxicology Volume 1. Principles
Federal Insecticide, Fungicide and Rodenticide Act (FIFRA) when certain toxicity and/or exposure criteria have been met for a given pesticide (Driver and Wilkinson, 1996). Although at a low level relative to occupational exposures, the major source of chemical exposures for the general population, appears to result from the use of products in and around the home (Hill et aI., 1995; U.S. EPA, 1999a; Whitmore et aI., 1994). For example, the National Academy of Sciences Committee on Urban Pest Management noted that 5000 health-related incidents involving pesticides were reported as occurring in homes in the United States from 1966 to 1979 (NRC, 1980). More recent data regarding pesticide healthrelated incidents can be obtained from the American Association of Poison Control Centers (http://www.aapcc.org), the U.S. EPA (http://www.epa.gov/pesticides), or state regulatory agencies, (e.g., California Department of Pesticide Programs (http://www.cdpr.ca.gov). It is perhaps not surprising, therefore, that the potential health risks associated with exposure to chemicals such as pesticides occurring in and around the home (in air and from surfaces) and from other sources (e.g., consumer products, combustion appliances, and outdoors) are being evaluated much more carefully now than in the past (Driver and Wilkinson, 1996). Pesticides, of course, are just one of many types of chemicals to which humans are exposed in the home. During the past decade and a half, a number of studies, most notably the Total Exposure Assessment Methodology (TEAM) studies sponsored by the EPA, have demonstrated that, for a variety of contaminants, indoor air and other residential pathways are often a more significant source of exposure than corresponding outdoor pathways (Curry et aI., 1994; Furtaw et aI., 1993; Pellizzari et aI.,
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CHAPTER 16
Residential Exposure Assessment: An Overview
Dissipation Studies Dislodgeable Foliar Residue (DFR) Dissipation Study Soil Residue Dissipation (SDR) Study Indoor Surface Residue (ISR) Dissipation Study
SOURCES
Measurements ofHuman Exposure Dermal Exposure (passive dosimetry) Inhalation Exposure Biological Monitoring Other Relevant Data Descriptions of Human Activity Data Data Reporting and Calculations Detailed Product Use Information
ACTIVITIESI PATHWAYSI ROUTES
ApplicatIon
Post-Application
//j
Figure 16.1 U.S. EPNOPPTS Series 875, Group B: description of required studies (adapted from ILSI, 1998). RECEPTORS
1987,1993; Thomas et aI., 1993; Wallace, 1993). Indeed, some of the studies found that indoor concentrations of some chemicals were higher than those outdoors and raised serious questions about the relative contribution of indoor sources to the total exposure (Dockery and Spengler, 1981; Melia et al. , 1978 ; Ott, 1985; Spengler et al., 1983). The assessment of potential indoor exposures has also been recognized by industry as a key component in the overall risk evaluation for consumer products (Hakkinen et al., 1991; Hakkinen, 1993). Several studies of potential indoor air exposures to chemicals have been reported to confirm the safety of various consumer products (Gibson et al., 1991; Hendricks, 1970; Wooley et aI., 1990).
16.2 OVERVIEW OF GENERAL ISSUES The residential environment should be considered in very dynamic terms. Chemicals that are released into or otherwise enter the residential environment tend to partition into various compartments, either through direct dispersion in indoor air or through adsorption onto surfaces that serve as "sinks" from which material can subsequently be released into the air (Ross et aI., 1990, 1991). The amount of a given chemical in each compartment is depleted over time by various mechanisms, including air exchange with the outdoors or with other rooms of the house, and by chemical or physical transformation and/or degradation. Simulation of the behavior of pesticides (and other chemicals) in residential environments can be modeled using the principles of fugacity (Matoba, 1996). Fugacity employs the unit of pressure (Pa), which refers to the external force of a chemical escaping from one compartment or medium to another. For example, there is evidence that particulate contaminants, whether generated inside the residence or brought in from outdoors are adsorbed to surfaces and are later resuspended and recycled within the house after a disturbance (e.g., walking on floors and rugs, sweeping and dusting, and vacuuming ; Roberts et al., 1992). A simplistic depiction of the relationship between potential sources and exposure pathways in the context of a residential exposure assessment is illustrated in Fig. 16.2. Thus, the residence can be considered an exposure unit containing multiple compartments with which the human
Figure 16.2 1998).
Shematic diagram for nondietary exposure (adapted from ILSr,
receptor can contact. Figure 16.3 illustrates the components of a residential exposure model and Fig. 16.4 illustrates the decision logic associated with construction of a residential exposure assessment. Although inhalation exposure and indoor air quality have received the most attention to date, there are a number of noninhalation exposure pathways that are likely to be of equal or greater importance for human residential exposures to pesticides and other chemicals. These include potential dermal exposure to dislodgeable chemical residues from surfaces such as floors and carpets or from hard surfaces resulting from the use of formulations for cleaning and disinfection and potential ingestion of surface contaminants resulting from hand/objectto-mouth activity, particularly in infants and toddlers. Several studies and/or reviews provide examples of noninhalation residential exposures and the complexities involved (Calvin, 1992; CTFA, 1983; Driver et aI., 1989; Eberhart, 1994; ECETOC, SOURCE CHARACTERISTICS ~ PRODUCT USE PATTERNS & PLAUSIBLE SCENARIOS
~
HUMAN ACTIVITY PATTERNS POPULATION DEMOGRAPHIC:-:::-& STRATIFICATION
NON-DIETARY EXPOSURE MODEL HUMAN EXPOSURE FACTORS _ _ PHYSICO-CHEMICALPROPERTIES RESIDENTIAL BUILDING
~ ~
FACTOR~
TEMPORAL & SPATIAL DOMAINS UNCERTAINTY ANALYSIS MODEL VALIDATION
Figure 16.3 Residential exposure assessment: key components (adapted from ILSI, 1998).
16.3 Lessons Learned from Key Studies Hazard Identification and Toxicity Endpoint Selection (e.g., toxicological benchmarks; time-to-effect; exposure/dose metric selection; relevant subpopulations, etc.)
1
Evaluation of Consumer and Commercial Product Use Information (e.g., application methods, frequencies, locations, rates, formulation types, active ingredient concentrations, post-application activities, consumerlprofessional survey data, market share data, etc.)
1
Evaluation of Exposure Monitoring Data (for actual pesticide and/or relevant surrogate chemicals; data limitations; uncertainty & variability; data quality objectives)
1
Determination of Relevant (Plausible) Aggregate Exposure Scenarios and Related Routes and Pathways of Exposure
~ Development and Implementation of Screening-Level Deterministic and/or Stochastic
Exposure~ Model Validation & Refinement
Figure 16.4 Decision logic and model development/validation process (adapted from ILSI, 1998).
1994; Harris and Solomon, 1992; Harris et aI., 1992; Turnbull and Rodricks, 1989; Vermiere et aI., 1993).
16.3 LESSONS LEARNED FROM KEY STUDIES Pesticides applied in and around homes by both professional applicators and consumers are used in different ways for different purposes: (1) indoor uses (e.g., floor sprays or foggers for fleas) and outdoor uses (e.g., treatment of wasp nests and ant mounds; use of antimicrobial products in swimming pools); (2) turf uses (e.g., granular applications for control of soildwelling insect pests, preemergent and postemergent herbicide sprays) and ornamental uses (e.g., foliar sprays for shrubs); (3) home garden uses (e.g., fungicide dusts for tomatoes); and (4) structural pest control uses (e.g., structural treatment or insecticidal soil barriers to protect against termite invasion). The vast majority of V.S. households use pesticides (Whitmore et aI., 1992) and these uses undoubtedly present many opportunities for exposure during intended, label-directed use, misuse, and accidents (Whitmyre et aI., 1996). Other sources of indoor exposure to pesticides for the general population may be from ambient air, food, water, ambient particles and indoor house dust (Jenkins et al., 1992; Pellizzari et aI., 1993; Wallace, 1991,1993; Whitmore et aI., 1994). Residential pesticide monitoring studies have included general surveys of many different pesticides and measurements of air and surface concentrations of specific pesticides after applications of termiticides, crack and crevice or baseboard treatments, total release aerosols or foggers, broadcast applications, and hand-held sprays (Fenske et aI., 1990; Racke and Leslie, 1993; Whitmore et aI., 1994). These studies typically
437
demonstrate that, after pesticide use, measurable though relatively low-level residues exist in homes and that indoor exposures are often higher than outdoor exposures. Although, in most cases, such exposures are associated with negligible health risks (Whitmore et aI., 1994), potential residential exposures to infants and children continue to be the subject of debate and scientific investigation (Berteau et aI., 1989; Byrne et aI., 1998; Gibson et aI., 1998; NRC, 1993; Ross et aI., 1990, 1991; Vaccaro et aI., 1996; Zweiner and Ginsburg, 1988). However, pesticide biomonitoring of both adults and children demonstrate that absorbed doses from all sources range from fractions of micrograms to single digit micro grams per kilogram of body weight (Adgate et aI., 1998; Hill et aI., 1995; Krieger et aI., 2000, 2001; Vaccaro et aI., 1996). The Nonoccupational Pesticide Exposure Study involving about 250 residents of Jacksonville, Florida and Springfield, Massachusetts, clearly demonstrated measurable levels of indoor exposure. Participants in this study carried personal monitors for 24 hours that provided indoor and outdoor measurements of 32 common household pesticides and structural termiticides (Immerman and Schaum, 1990). The indoor air concentrations of these materials exceeded the outdoor air levels by factors even larger than those measured in the earlier TEAM studies on volatile organic compounds (VOCs). It was hypothesized that, with termiticides and other chemicals used outside the home, some material entered the house on soil particles in addition to infiltration in air from treated areas beneath and around the house. It was also noted that the use of "walkoff" rugs in hallways and the practice of removing shoes on entering the house would help to reduce indoor exposure levels (Roberts et aI., 1992). In the TEAM studies mentioned earlier, median personal concentrations of VOCs in V.S. residences were found to be 2-5 times outdoor levels and maximum personal concentrations were 5-70 times the highest outdoor levels (Wallace, 1993). The variability in indoor personal exposures probably reflects differences in human activity patterns that bring individuals into contact with chemicals indoors and suggests the importance of specific sources of residential exposures that may not be available to all individuals. Smokers, for example, had benzene exposures 6-10 times higher than those of nonsmokers, individuals wearing freshly dry cleaned clothes had significantly higher exposures to tetrachloroethylene (Wallace et aI., 1991 c), and persons using mothballs and solid deodorizers in the residence had greatly elevated exposures to p-dichlorobenzene relative to those of nonusers (Wallace, 1993). The most recent TEAM study, known as PTEAM, focused on measuring personal exposures to inhalable particles (PMIO) of approximately 200 Riverside, California, residents using specially designed indoor sampling devices. A major finding from this study was that personal exposures to particles in the daytime were 50% greater than either indoor or outdoor concentrations (Wallace, 1993). It has been hypothesized that these data suggest that individuals are exposed to a "personal cloud" of particles as they go about their daily activities (Wallace, 1993). Resuspension of household dust through walking in the resi-
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dence or by vacuuming, cooking, or sharing a home with a smoker lead to significant particle exposures. The recent Valdez Air Health Study conducted in Valdez, Alaska (Wallace, 1993) generally supports the findings of the TEAM studies in terms of the greater importance of individual indoor exposure sources than outdoor sources. In the Valdez study, mean personal concentrations of benzene were roughly 3-4 times higher than outdoor levels, despite the presence of a significant outdoor source of benzene in the community (i.e., a petroleum storage and loading terminal).
16.4 GUIDANCE FOR RESIDENTIAL POSTAPPLICATION EXPOSURE ASSESSMENT METHODS AND DATA SOURCES FOR EXPOSURE FACTORS The most recent effort to develop guidance for residential exposure assessment methods was initiated by the EPA's Office of Pesticide Programs in the Standard Operating Procedures for Residential Exposure Assessment (U.S. EPA, 1999a). The passage of the FQPA mandated the EPA to immediately begin routinely addressing nondietary and non-occupational pesticide exposures for the general population. These are exposures that can occur in a residential setting (or other areas frequented by the general population) and that do not occur as part of the diet or as a result of participation in occupational practices. These exposures may include breathing vapors while inside a treated home, exposures to children playing on a treated lawn, or exposures attributable to the mouthing behaviors of infants and children. Before passage of the FQPA, the Agency addressed these kinds of exposures on a case-by-case basis, typically in the "special review" process. The intent of the EPA SOPs is to provide a means for consistently calculating single pathway, screening level exposures and not to provide guidance on other related topics such as aggregate (multisource to a given pesticide) or cumulative (multi source to two or more pesticides with a presumed common mode of action) exposure assessment. These SOPs are the backbone of the Agency's current approach for completing initial tier (screening-level) residential exposure assessments. However, the state-of-the-science continues to evolve since the release of the original document in 1997 and the emphasis of industry, as well as of academia and others, has clearly focused on the scientific and policy issues raised by the implementation of the FQPA and the use of the first-generation SOPs. Thus, revisions to the SOPs are ongoing to reflect the development of scientific information and the development of refined methods for estimation of potential residential exposures to adults and children. Additional guidance for dermal exposure assessment methods and dermal permeability coefficients for some organic chemicals are contained in the EPA's dermal exposure assessment guidance document (U.S. EPA, 1992). Given that skin surface area and body weight are closely correlated, total skin surface area to body weight ratios for use in residential exposure assessments have been recommended (Phillips et aI.,
1993). Another excellent source for methodology and data relevant to consumer product exposure assessments is ECETOC (1994). A number of relevant data sources exist for key variables or factors used in performing residential exposure assessment. Data useful in estimating human exposures (e.g., distributions of body weights and skin surface areas, inhalation rates, and residential occupancy periods) can be obtained from the American Industrial Health Council's Exposure Factors Sourcebook l (AIHC, 1995) and the EPA's Exposure Factors Handbook (U.S. EPA, 1999b), which has recently been updated. Residential "environmental factors" such as air exchange rates have been summarized by Pandian et al. (1993). Human time-activity data in the United States were summarized by the EPA (U.S. EPA, 1991) and compiled in the THERdbASE software (Pandian and Furtaw, 1995), which is available on the Internet at http://www.therd.com. Multiple data sources for time-activity data have been included in EPA's Consolidated Human Activity Database, which is planned for future release via the Internet.
16.5 RESEARCH NEEDS Given that the potential for postapplication exposures largely exists because of product use in and around the home, the need to develop and validate models for prediction of multipathway, multiroute exposures and absorbed dose is evident. Historically, efforts have focused on indoor air and associated inhalation exposures. Jayjock and Hawkins (1993), for example, have explored the complementary roles of indoor air mode ling and data development in improving the level of confidence in estimates of indoor inhalation exposures. More recently, dermal and incidental ingestion exposures have been the focus of monitoring and modeling efforts (e.g., the Outdoor Residential Exposure Task Force, the Non-Dietary Exposure Task Force, the OP Case Study Group, and Residential Exposure Joint Venture (Zartarian and Leckie, 1998; Zartarian et aI., 2000). Multipathway, multiroute modeling efforts for pesticides include the Residential Exposure Assessment Model (REAM), the Stochastic Human Exposure and Dose Simulation model (SHEDS); the Cumulative and Aggregate Risk Evaluation System (CARES), and LifeLine (U.S. EPA, 1999a). The use of real-world data to validate residential exposure models is critical to developing estimates that are more representative than worst-case estimates typically obtained from unvalidated modeling approaches (Whitmyre et aI., 1992a, b). Other research activities related to residential exposure assessment currently being sponsored by the EPA include the National Human Exposure Assessment Survey. In addition, the EPA has recently concluded a cooperative agreement, referred to as the Residential Exposure Assessment Project (REAP) with the Society for Risk Analysis and the International Society of Exposure Analysis, to develop a reference textbook describing 1For the latest version of the Exposure Factors Sourcebook, contact the Update Coordinator, American Industrial Health Council, Suite 760, 2001 Pennsylvania Avenue, N.W., Washington, DC 20006-1807; phone: 202-833-2131.
References
relevant methodologies, data sources, and research needs for residential exposure assessment. The REAP will complement other EPA initiatives, such as the development of the series 875 guidelines and will facilitate a sharing of information and other resources between the EPA, other federal and state agencies, industry, academia, and other interested parties. Residential exposures to pesticides and other chemicals are estimated by means of either monitoring and/or predictive modeling but, unfortunately, little or no guidance is available for those attempting such estimates. Key areas requiring attention include: • Characterization of temporal product use patterns (particularly the likelihood of co-occurrence of more than one product use event) and associated demographic and postapplication activity information relevant to occupants of homes using products. • Source characterization, including emission rates, surface deposition, tranferability to human clothing and skin, and physicochemical factors driving fate and transport processes. • The complex interaction over time of environmental media residue concentrations with humans resulting from variable time-activity patterns that determine subsequent residential exposures (inhalation, dermal, and incidental ingestion). • Identification of the fundamental principles, concepts, and methods for conducting multipathway/multiroute residential exposure assessments, including unique pathways such as incidental dermal exposure to dislodgeable pesticide residues from treated lawns, incidental ingestion of contaminated soil particles during gardening, hand-to-mouth transfer by infants and children, and dermal exposure to dislodge able pesticide residues from carpets and other treated surfaces, and incidental ingestion of postapplication residues in food. • Characterization of key human exposure factors (ranges and distributions of factors such as age-specific inhalation rates, product use patterns, and human time-activity data) and residential building factors (distributional data on housing stock type, number and size of rooms, air exchange rates, source emission rates, and sink effects, i.e., adsorption! desorption from various surfaces in the home) that influence residential exposure and dosimetry. • Continued development and validation of methods for measuring and mode ling indoor chemical fate processes (e.g., volatilization from surfaces and dislodgeable residue kinetics), chemical concentrations in complex matricies (such as house dust) and human intake (e.g., incidental ingestion, inhalation, and dermal exposure). • Development and validation of methods for extrapolating from short-term monitoring data to long-term exposure scenarios and for extrapolation of adult monitoring data to children. • Continued development and application of methods for quantifying uncertainty and variability (e.g., Monte Carlo methods) in residential exposure (and risk) estimates.
439
• The development and use of effective methods for comparing and communicating residential exposure and risk estimates to risk managers and the general public.
REFERENCES Adgate, J., Quackenboss, J., Needham, L., Pellizari, P., Lioy, P., Shubat, P., and Sexton, K. (1998). Comparison of urban versus rural pesticide exposure in Minnesota children. In "Annual Conference of International Society for Environmental Epidemiology (ISEE) and International Conference for Society of Exposure Analysis (ISEA)," July 1998, Vol. 9, No. 4, Suppl., Abstract 920. American Industrial Health Council (AIHC). (1995). "Exposure Factors Sourcebook." AIHC, Washington, DC. Berteau, P. E., Knaak, J. B., Mengle, D. C., and Schreider, J. B. (1989). Insecticide absorption from indoor surfaces. In "Biological Monitoring for Pesticide Exposure" (R. G. Wang, C. A. Frankiin, R. C. Honeycutt, and J. C. Reinert, eds.), ACS Symposium Series, Vol. 382, pp. 315-326. Am. Chem. Soc., Washington, DC. Byrne, S. L., Shurdut, B. A., and Saunders, D. G. (1998). Potential chlorpyrifos exposure to residents following standard crack and crevice treatment. Env. Health Perspect. 106,725-731. California Environmental Protection Agency (Cal-EPA). (1994). CalTOX™, a Multimedia Total Exposure Model for Hazardous-Waste Sites. Spreadsheet user's guide, version 1.5. NTIS Publication No. PB95-100467. Office of Scientific Affairs, Dep. of Toxic Substances Control, Cal-EPA, Sacramento. Calvin, G. (1992). Risk management case history-detergents. In "Risk Management of Chemicals" (M. L. Richards, ed.). Royal Society of Chemistry, United Kingdom. Cosmetic, Toiletry and Fragrance Association, Inc. (CTFA). (1983). "Summary of the Results of Surveys of the Amount and Frequency of Use of Cosmetic Products by Women." Report prepared by ENVIRON Corp. CTFA, Washington, DC. Curry, K. K., Brookman, D. J., Whitmyre, G. K., Driver, J. H., Hackman, R. J., Hakkinen, P. J., and Ginevan, M. E. (1994). Personal exposures to toluene during use of nail lacquers in residences: description of the results of a preliminary study. J. Expos. Anal. Environ. Epidemiol. 4, 443-456. Dockery, D. W., and Spengler, J. D. (1981). Indoor-outdoor relationships of respirable sulfates and particles. Atmos. Environ. 15,335-343. Driver, J. H., Konz, J. J., and Whitmyre, G. K.. (1989). Soil adherence to human skin. Bull. Environ. Contam. Toxicol. 43, 814-820. Driver, J. H., and Wilkinson, C. F. (1996). Pesticides and human health: science, regulation and public perception. In "Risk Assessment and Management Handbook for Environmental, Health and Safety Professionals" (R. V. Kalluro, S. M. Bartell, R. M. Pitblado, and R. S. Stricoff, eds.). McGraw-Hill, New York. Eberhart, D. C. (1994). Current activities in assessing human exposures to lawn chemicals. In "Workshop on Residential Exposure Assessment, Annual Meeting of the International Society for Exposure Analysis and the International Society for Environmental Epidemiology," Research Triangle Park, NC. European Centre for Ecotoxicology and Toxicology of Chemicals (ECETOC). (1994). Tech. Rep. 58, Assessment of Non-Occupational Exposure Chemicals, Brussels. Fenske, R. A., Black, K. G., Elkner, K. P., Lee, c., Methner, M. M., and Soto, R. (1990). Potential exposure and health risks of infants following indoor residential pesticide applications. Am. J. Publ. Health 80, 689-693. Furtaw, E. J., Pandian, M. D., and Behar, J. V. (1993). Human exposure in residences to benzene vapors from attached garages. In "Proceedings of International Conference: Indoor Air '93," Helsinki, Finland. Gibson, J. E., Peterson, R. K. D., and Shurdut, B. A. (1998). Human exposure and risk from indoor use of chlorpyrifos. Environmental Health Perspectives 106,303-306. Gibson, W. S., Keller, F. R., Foltz, D. J., and Harvey, G. J. (1991). Diethylene glycol monobutyl ether concentrations in room air from application of
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CHAPTER 16
Residential Exposure Assessment: An Overview
cleaner formulations to hard surfaces. 1. Expos. Anal. Environ. Epidemiol. 1,369-383. Hakkinen, P. J. (1993). Cleaning and laundry products: human exposure assessments. In "Handbook of Hazardous Materials," pp. 145-151. Hakkinen, P. J., Kelling, e. K, and Callender, J. e. (1991). Exposure assessment of consumer products: human body weights and total body surface areas to use, and sources of data for specific products. Vet. Hum. Toxicol. 33,61-65. Harris, S. A., and Solomon, K R (1992). Human exposure to 2,4-D following controlled activities on recently-sprayed turf. 1. Environ. Sci. Health B27(1),9-22. Harris, S. A., Solomon, K R., and Stephenson, G. R. (1992). Exposure of homeowners and bystanders to 2,4-dichlorophenoxyacetic acid (2,4-D). 1. Environ. Sci. Health B27(1), 23-38. Hendricks, M. G. (1970). Measurement of enzyme laundry product dust levels and characteristics in consumer use. 1. Am. Oil. Chem. Soc. 47,207-211. HilI, R H., Jr., Head, S. L., Baker, S., Gregg, M., Shealy, D. B., Bailey, S.L., WilIiams, e. e., Sampson, E. J., and Needham, L. L. (1995). Pesticide residues in urine of adults living in the United States: reference range concentrations. Environ. Res. 71, 99-108. Immerman, W. W., and Schaum, J. L. (1990). "Nonoccupational Pesticide Exposure Study (NOPES)." NTIS Publication No. PB90-152224, 256 p. Research Triangle Park, NC: Atmospheric Research and Exposure Assessment Laboratory, U.S. Environmental Protection Agency. Report prepared by Research Triangle Institute, Research Triangle Park, NC. International Life Sciences Institute (ILSI) (1998). (S. S. Olin, ed.). "Aggregate Exposure Assessment." ILSI Risk Sciences Institute Workshop Report, Washington, De. Jayjock, M. A and Hawkins, N. e. (1993). A proposal for improving the role of exposure modeling in risk assessment. Am. Ind. Hyg. Assoc. 1. 54, 733-741. Jenkins, P. L., PhilIips, T. H., Mulberg, E. J., and Hui, S. P. (1992). Activity patterns of Californians: use of and proximity to indoor pollutant sources. Atmos. Environ. 26, 2141-2148. Krieger, R I., Bernard, C. E., Dinoff, T. M., Fell, L., Osimitz, T. G., Ross, J. H., and Thongsinthusak, T. (2000). Biomonitoring and whole body cotton dosimetry to estimate potential human dermal exposure to semivolatile chemicals. 1. Expos. Anal. Environ. Epidemiol. 10,50-57. Krieger, R. I., Bernard, e. E., Dinoff, T. M., Ross, J. H., and WiIIiams, R L. (2001). Biomonitoring of persons exposed to insecticides used in residences. Ann. Occup. Hyg. 45(Suppl 1),5143-5153. Matoba, Y. (1996). Simulation of indoor behavior of insecticides applied by various methods. In "SP World," No. 24. Sumitomo Chemical Co., Osaka, Japan. McKone, T. E. (1991). Human exposure to chemicals from multiple media and through multiple pathways: research overview and comments. Risk Anal. 11,5-10. McKone, T. E. (1993). Understanding and modeling multipathway exposures in the home. In "Reference House Workshop II: Residential Exposure Assessment for the '90s." Society for Risk Analysis, 1993 Annual Conference, Savannah, GA. Melia, R J. W, F1orey, e. duV., Darby, S. e., Palmes, E. D., and Goldstein, B. D. (1978). Differences in N02 levels in kitchens with gas or electric cookers. Atmos. Environ. 12, 1379-1381. National Research Council (NRC). (1980). Committee on Urban Pest Management. Nat. Acad. Press, Washington, De. National Research Council (NRC). (1993). "Pesticides in the Diets of Infants and Children. Committee on Pesticides in the Diets of Infants and Children," Board on Agriculture and Board on Environmental Studies and Toxicology, Commission on Life Sciences, Nat. Acad. Press, Washington, DC. Ott, W. R. (1985). Total human exposure: an emerging science focuses on humans as receptors of environmental pollution. Environ. Sci. Technol. 19, 880. Pandian, M. D., Ott, W R, and Behar, J. V. (1993). Residential air exchange rates for use in indoor air and exposure mode ling studies. 1. Expos. Anal. Environ. Epidemiol. 3,407-416. Pandian, M. D. and Furtaw, E. J. (1995). ''THERdbASE: Total Human Exposure Relational Database and Advanced Simulation Environment." Harry Reid
Center for Environmental Studies, University of Nevada at Las Vegas, Las Vegas. Developed under contract to the U.S. EPA, Office of Research and Development, Environmental Monitoring Systems Laboratory, Las Vegas. Pellizzari, E. D., Hartwell, T. D., Perritt, R. L., Sparacino, e. M., Sheldon, L. S., Whitmore, R W., and Wallace, L. A. (1987). Comparison of indoor and outdoor residential levels of volatile organic chemicals in five U.S. geographic areas. Environ. Int. 12,619-623. PeIIizzari, E. D., Thomas, K W, Clayton, C. A, Whitmore, R W, Shores, R e., Zelon, H. S., and Peritt, R L. (1993). "Particle Total Exposure Assessment Methodology (PTEAM): Riverside, California Pilot Study," Vo!. I, EPAJ600/SR-93/050. U.S. EPA, Research Triangle Park, Ne. PhilIips, L. J., Fares, R. J., and Schweer, L. G. (1993). Distributions of total skin surface area to body weight ratios for use in dermal exposure assessments. 1. Expos. Anal. Environ. Epidemiol. 3, 331-338. Racke, K D., and Leslie, A. R (eds.). (1993). "ACS Symposium Series 522. Pesticides in Urban Environments: Fate and Significance." 203rd National Meeting of the American Chemical Society, San Francisco, California, April 5-10, 1992. Published by the American Chemical Society, Washington, DC, ISBN 0-8412-2627-X, 378 p. Roberts, J. W., Budd, W. T., Ruby, M. G., Camann, D. E., Fortmann, R e., Lewis, R. G., Wallace, L. A, and Spittler, T.M. (1992). Human exposure to pollutants in the floor dust of homes and offices. 1. Expos. Anal. Environ. Epidemiol. 1 (Supp!. 1),127-146. Ross, J. H., Fong, H. R., Thongsinthusak, T., Margetich, S., and Krieger, R (1991). Measuring potential dermal transfer of surface pesticide residue generated from indoor fogger use: using the CDFA roller method. Interim report II. Chemosphere, 22, 975-984. Ross, J., Thongsinthusak, T., Fong, H. R, Margetich, S., and Krieger, R (1990). Measuring potential dermal transfer of surface pesticide residue generated from indoor fogger use: an interim report. Chemosphere, 20, 349-360. Spengler, J. D., Duffy, C. P., Letz, R., Tibbets, T. W, and Ferris, B. G. Jr. (1983). Nitrogen dioxide inside and outside 137 homes and implications for ambient air quality standards and health effects research. Environ. Sci. Technol. 17(3), 164--168. Thomas, K W, PeIIizzari, E. D., Clayton, C. A., Perritt, R. L., Dietz, RN., Goodrich, R. W, Nelson, W e., and Wallace, L. A (1993). Temporal variability of benzene exposures for residents in several New Jersey homes with attached garages or tobacco smoke. 1. Expos. Anal. Environ. Epidemiol. 3, 49-73. Thompson, D. G., Stephenson, G.R, and Sears, M. K (1984). Persistence, distribution, and dislodgeable residues of 2,4-D following its application to turfgrass. Pestic. Sci. 15,353-360. Turnbull, D., and Rodricks, J. V. (1989). A comprehensive risk assessment of DEHP as a component of baby pacifiers, teethers and toys. In "The Risk Assessment of Environmental and Human Health Hazards: A Textbook of Case Studies" (D. J. Paustenbach, ed.). WiIey, New York. U.S. Environmental Protection Agency (U.S. EPA). (1991). "Time Spent in Activities, Locations, and Microenvironments: A California-National Comparison," USEPA Pub!. No. 600/4-91/006. Office of Research and Development, Environmental Monitoring Systems Laboratory, Las Vegas. U.S. Environmental Protection Agency (U.S. EPA). (1992). "Dermal Exposure Assessment: Principles and Applications," Washington, DC: USEPA Pub!. No. 600/8-91-011. Exposure Assessment Group, Office of Health and Environmental Assessment, Office of Research and Development, Washington, De. U.S. Environmental Protection Agency (U.S. EPA). (1999a). "Overview of Issues Related to the Standard Operating Procedures for Residential Exposure Assessment." Presented to the EPA Science Advisory Panel U.S. EPA, OPP, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA). (1999b). "Exposure Factors Handbook;' USEPA Pub!. No. EPAJ600/C-99/001. National Center for Environmental Assessment, Cincinnatti. Vaccaro, J. R., Nolan, R J., Murphy, P. G., and Berbrich, D. B. (1996). "The Use of Unique Study Design to Estimate Exposure of Adults and Children to Surface and Airborne Chemicals," STP 1287, pp. 166-183. Am. Soc. for Testing and Materials, West Conshohocken, PA.
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Vermiere, T.G., van der Poel, P., van de Laar, R. T. H., and Roelfzema, H. (1993). Estimation of consumer exposures to chemicals: applications of simple models. Sei. Total Environ. 136, 155-176. WaIIace, L. A. (1987). "The TEAM Study: Summary and Analysis, Vo!. I," EPA 600/6-87/002a. V.S. Environmental Protection Agency, Office of Research and Development, Nat. Tech. Information Service, Springfield, VA. WaIIace, L. A. (1989). The exposure of the general population to benzene. Cell Bio!. Toxieol. 5, 297-314. WaIIace, L. A. (1990). Major sources of exposure to benzene and other volatile organic compounds. Risk Ana!' 10, 59-64. WaIIace, L. A. (1991). Comparison of risks from outdoor and indoor exposure to toxic chemicals. Environ. Health Perspeet. 95, 7-13. WaIIace, L. (1993). A decade of studies of human exposure: what have we learned? Risk Anal. 13, 135-139. Wallace, L. A., Zweidinger, R., Erickson, M., Cooper, S., Whitaker, D., and Pellizzari, E. (1982). Monitoring individual exposure: measurement of volatile organic compounds in breathing-zone air, drinking water, and exhaled breath. Environ Internat. 8, 269-282. WaIIace, L. A., PeIlizzari, E., HartweII, T., Rosenzweig, R., Erickson, M., Sparacino, c., and ZeIon, H. (1984). Personal exposure to volatile organic compounds: I. Direct measurement in breathing-zone air, drinking water, food, and exhaled breath. Environ. Res. 35, 293-319. WaIIace, L. A., PeIlizzari, E., HartweII, T., Sparacino, c., Sheldon, L., and Zelon, H. (1985). Personal exposures, indoor-outdoor relationships and breath levels of toxic air agents measured for 355 persons in New Jersey. Atmos. Environ. 19, 1651-1661. WaIIace, L. A., PeIIizzari, E., HartweII, T., Whitmore, R., Sparacino, c., and Zelon, H. (1986). Total exposure assessment methodology (TEAM) study: personal exposures, indoor-outdoor relationships, and breath levels of volatile organic compounds in New Jersey. Environ. Int. 12,369-387. WaIIace, L. A., Pellizari, E. D., HartweII, T. D., Sparacino, c., Whitmore, R., Sheldon, L., Zelon, H., and Perrit, R. (l987a). The TEAM study: personal exposures to toxic substances in air, drinking water, and breath of 400 residents of New Jersey, North Carolina, and North Dakota. Environ. Res. 43, 290-307. WaIIace, L. A., Pellizari, E., Hartwell, T., Perritt, K., and Ziegenfus, R. (1987b). Exposures to benzene and other volatile organic compounds from active and passive smoking. Areh. Environ. Health 42, 272-279. WaIIace, L. A., PeIlizari, E., Leaderer, B., Hartwell, T., Perritt, R., Zelon, H., and Sheldon, L. (I 987c). Emissions of volatile organic compounds from building materials and consumer products. Atmos. Environ. 21,385-393. WaIIace, L. A., Pellizzari, E. D., HartweII, T. D., Whitmore, R., Perritt, R., and Sheldon, L. (1988). The California TEAM study: breath concentrations and personal exposures to 26 volatile compounds in air and drinking water of 188 residents of Los Angeles, Antioch, and Pittsburgh, CA. Atmos. Environ. 22,2141-2163.
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WaIIace, L. A., PeIIizzari, E. D., Hartwell, T. D., Davis, v., Michael, L. C., and Whitmore, R. W. (1989). The influence of personal activities on exposure to volatile organic compounds. Environ. Res. 50, 37-55. WaIIace, L. A., Nelson, W. c., Ziegenfus, R., and PeIlizzari, E. (l991a). The Los Angeles TEAM study: personal exposures, indoor-outdoor air concentrations, and breath concentrations of 25 volatile organic compounds. J. Expos. Ana!. Environ. Epidemiol. 1(2),37-72. WaIIace, L. A., PeIIizzari, E., and Wendel, C. (l99Ib). Total volatile organic concentrations in 2700 personal, indoor, and outdoor air samples collected in the VSEPA TEAM studies. Indoor Air 4, 465-477. WaIIace, L. A., Pellizzari, E., Sheldon, L., Hartwell, T., Perritt, R., and Zelon, H. (1991c). Exposures of dry cleaning workers to tetrachloroethylene and other volatile organic compounds: Measurements in air, water, breath, blood, and urine. In "Annual Meeting of the International Society for Exposure Analysis and Environmental Epidemiology," Atlanta. Whitford, E, Kronenberg, J., Lunchick, c., Driver, J., TomerIin, R., Wolt, J., Spencer, H., Winter, c., and Whitmyre, G. (1999). "Pesticides and Human Health Risk Assessment: Policies, Processes and Procedures," Purdue Pesticide Programs, Publication PPP-48. Purdue Vniv. Cooperative Extension Service, West Lafayette, IN. Whitmore, R. W, KeIIy, J. E., and Reading, P. L. (1992). "National Home and Garden Pesticide Vse Survey: Final Report," NTIS PB92-174739. V.S. EPA, Office of Pesticide Programs and Toxic Substances, Washington, DC. Whitmore, R. W., Immerman, E W., Camann, D. E., Bond, A. E., Lewis, R. G., and Schaum, J. L. (1994). Non-occupational exposures to pesticides forresidents of two V.S. cities. Areh. Environ. Contam. Toxieol. 26,47-59. Whitmyre, G. K., Driver, J. H., Ginevan, M. E., Tardiff, R. G., and Baker, S. R. (1992a). Human exposure assessment. I: understanding the uncertainties. Toxieo!. Ind. Health 8, 297-320. Whitmyre, G. K., Driver, J. H., Ginevan, M. E., Tardiff, R. G., and Baker, S. R. (l992b). Human exposure assessment. 11: quantifying and reducing the uncertainties. Toxieo!. Ind. Health 8, 321-342. Whitmyre, G. K., Driver, J. H., and Hakkinen, P. J. (1996). Assessment of residential exposures to chemicals. In "Fundamentals of Risk Analysis and Risk Management" (V. Molak, ed.), pp. 125-141. CRC Lewis Publishers, Boca Raton, Florida. Wooley, J., Nazaroff, W W, and Hodgson, A. T. (1990). Release of ethanol to the atmosphere during use of consumer cleaning products. 1. Air Waste Management Assoe. 40, 1114-1120. Zartarian, V. G., and Leckie, J. O. (1998). Dermal exposure: the missing link. Environ. Sei. Teehnol. March I, 134-137. Zartarian, V. G., Ozkaynak, H., Burke, J. M., ZufaII, M. J., Rigas, M. L., and Furtaw, E. J. (2000). Amodeling framework for estimating residential exposure to and dose of chlorpyrifos via dermal residue contact and non-dietary ingestion. Environ. Health Perspeet. 108,505-514. Zweiner, R. J., and Ginsburg, C. M. (1988). Organophosphate and carbamate poisoning in infants and children. Pediatries 81,121-126.
CHAPTER
17 Modeling Dietary Exposure with Special Sections on Modeling Aggregate and Cumulative Exposure Barbara J. Peters en , Susan H. Youngren, and Cassi L. Walls Novigen Sciences
Exposure assessment methodology is rapidly changing to meet the demands of consumers, regulators, and researchers. Passage of the Food Quality Protection Act (FQPA) in 1996 challenged the discipline of risk assessment to develop methods that are scientifically defensible and that provide estimates to meet the legal requirements. New methods that allow quantitative estimation of exposure from different sources and from mUltiple chemicals are now needed. Additionally, richer data sets also need to be incorporated into the new methodologies to reflect realistic exposure patterns. The current methods that use default assumptions or simplistic methodologies generate "worst case" exposure and risk estimates that are extreme. These simplified approaches do not permit any discrimination between factors that are likely to reflect risks that really exist and those that are truly hypothetical. Researchers in academic and industrial positions are generating extensive exposure data sets. However, these data need to be understood. For example, the results of biomonitoring studies need to be interpreted in relation to potential exposures and risks. Specifically, environmental residues are being measured with increasingly sensitive analytical methods that can now detect contaminants at much lower levels. These analyses produce results that require statistical treatment of variations in levels of detection and sampling variability in new ways. Fortunately, computer technology also has improved so that it is possible to conduct analyses relatively rapidly that would have been almost impossible even 10 years ago. This chapter focuses on methodology to estimate dietary exposures and provides options that range from "worst case" screening assessments to much more refined and accurate assessments. Separate sections introduce the relatively new and much more complicated techniques that are under consideration for conducting aggregate and cumulative assessments. Handbook of Pesticide Toxicology Volume 1. Principles
An aggregate assessment estimates exposure to one chemical via multiple pathways including inhalation, dermal, and oral routes. A cumulative assessment estimates exposure to multiple chemicals via multiple pathways. The methods for estimating cumulative exposure are presented from the point that it has been determined that the chemicals cause toxicity by a common mechanism. Figure 17.1 provides a schematic overview of an aggregate and cumulative assessment process.
17.1 OVERVIEW There are several features that are required in any exposure assessment. Examples of common features include the ability to provide estimates of exposure that are statistically representative of the population to be evaluated, the ability to consider the impact of various scenarios and data, and the ability to provide documentation that is suitable for regulatory decisions and/or to support peer review. Typically the analyst also requires estimates for a range of user-specified subpopulations as well as for the overall population. In the case of aggregate or cumulative exposure, it is also necessary to be able to estimate exposure to single or multiple compounds, respectively, for relevant time periods. The goal of aggregate and cumulative exposure assessments is to characterize the exposure of the population of concern (e.g., adults, toddlers), and to identify the variability and uncertainty associated with that exposure. The exposures are characterized by estimating the level of chemical uptake via ingestion, inhalation, and/or dermal absorption of the substance over various time periods. The typical time periods over which exposures may be evaluated include daily/acute, short-term (1-7 days), intermediate-
443
Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
444
CHAPTER 17
Modeling Dietary Exposure Non-Dietary Assessment
Dietary Assessment USDA CS F[]
Pesticide Use Patterns
Other Residue Adjustments
Figure 17.1
Components of an aggregate and cumulative risk assessment.
term, and chronic (up to 1 year) time periods. These time periods are based on the toxicity profile of the chemical. In addition, the exposure assessment also can be useful to identify the potential importance of a specific route relative to other pathways of exposure. In these types of assessments, exposure can result from residues in food, residues in or around the residence, from residues in parks/schools/towns, and/or residues from occupational uses of the pesticide. Therefore, it is desirable to be able to estimate the contribution of each route to exposure. That is, the method should identify the proportion of exposure that can result from oral, dermal, or inhalation, or a combination of these routes. In many cases, exposures from more than one source need to be considered. For example, oral exposure can arise as a result of residues in the diet or from other pathways such as toddler hand-to-mouth activity. The methods should allow the user to aggregate exposures as appropriate for the scenarios under consideration. Aggregation may be relevant to one chemical contained in one product that has multiple routes of exposure. For example, a compound might be used exclusively in a lawn product that is applied
by the homeowner. In this case, the person applying the compound may be exposed by dermal and inhalation routes while applying the compound and using the lawn after the treatment. Aggregation might also be relevant to one chemical contained in multiple products and with multiple exposure routes. For example, if a chemical is used on crops and also as a termiticide, possible routes of exposure are oral (from residues on food), dermal (from contact with treated surfaces in the home), and inhalation (from residues in the air around the home). For cumulative exposure assessments, the methods need to simulate situations that encompass exposure to more than one chemical, with multiple uses or sources and multiple exposure routes (Fig. 17.2). Cumulative assessments should be limited to chemicals that have a similar mode of action toxicologically. This requires that the relative toxicity of the various chemicals included in the analysis be quantitatively specified. It is also necessary that the exposure methods account for temporal and spatial considerations. The temporal aspect should consider when exposures occur simultaneously. For example, the method should account for overlap of exposures based on product usage information and chemical degradation. The spatial or regional aspect should be included in the exposure methodology. For example, the types of contaminants encountered in a home in Florida may be very different from those found in a home in northern Maine. In summary, exposure estimates must be able to assess exposures that are specific to both time and location.
17.2 EXPOSURE MODELS 17.2.1 GENERAL EXPOSURE MODEL Exposure assessment models can be designed for a specific route or for a specific source of exposure. Routes of exposure - Dietary
80
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Figure 17.2
Hypothetical multiuse exposures.
211
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17.2 Exposure Models
to pesticides can include oral, inhalation, and dermal pathways. Exposures can result from the treatment of foods, homes, schools, workplaces, yards, gardens, parks, golf courses, etc. All exposure models incorporate the following variables in one form or another:
445
occur simultaneously (if they do or could) and should not assume that exposure occur simultaneously (if they do not or are unlikely to occur at the same time).
17.2.2 GENERAL DIETARY MODEL contact x residue = exposure The "contact" function refers to the amounts of foods eaten for dietary exposure and breathing rates/activity patterns, etc. for nondietary exposures. The "residue" function refers to the concentration of the pesticide or pesticides on the foods that were consumed for dietary exposure and to levels in other media (air, water, on surfaces) for other routes of exposure. Each function may have multiple parameters and can be represented by worst case estimates or by refined statistically representative data sets. To assess the dietary aggregate or cumulative exposure, three types of data for each product or use are required:
1. Information about how the chemical enters the media. For example, for pesticides used in the home, information that is needed includes use pattern information of products of interest, frequency of application, and amount of product applied. 2. Environmental concentration data on relevant days. Again, for a pesticide used in the home, environmental data are needed before, during, and after treatment (residue factors). 3. Exposure factors such as food consumption information, body weight, breathing rate, and activity patterns (contact factors). Selection of the most appropriate methodology for estimating exposure depends on the intended application or purpose for the exposure assessment, the available data and related resources, and the relative importance of each route of exposure. The exposure route or pathway plays a critical role in choosing the most appropriate exposure algorithms for use in the assessment. The appropriate algorithms need to reflect the manner in which the chemical is contacted. The contribution of the diet to exposure must include estimates of consumption of each food as well as the residues including estimates of the proportion of food that contain residues, the impact of cooking and processing, and the likelihood that the water supply also contains residues. Likewise, factors must be selected for aggregate exposure assessments. For example, if exposure to residues from an indoor fogger is evaluated, then the exposure algorithms must account for residues in the air (for the inhalation route) and residues deposited on the carpet (for incidental ingestion via hand-to-mouth and dermal routes). Parameters that are not independent need to be treated differently. For example, high consumers of one food may be low consumers of another food and such relationships must be included in the assessment. In each exposure analysis, routes should be "linked" if exposures are dependent. Linking should not aggregate scenarios; rather, it should assume that exposures
The equation to calculate dietary exposure, is deceptively simple, exposure = C x R
(17.1)
where C is the amount of the food consumed and R is the concentration of chemical in the food. The estimate of consumption can include both quantity and frequency of consumption. This equation can be modified to incorporate other parameters as necessary. In a detailed exposure assessment, the basic algorithm [Eq. (17.1)] is modified to allow the analyst to better match the available data to the toxicity profile of the chemical. Specifically, the analyst will utilize much more extensive data that permits refining who (i.e., toddlers, all adults, adult females) is exposed and under what conditions (i.e., exposure time). For example, exposure can be estimated for an average or median individual or exposure can be described as the distribution of potential exposures for a specific population. Additionally, the analyst must decide the appropriate period of exposure time that is relevant to the toxicity profile to include in the exposure estimate.
17.2.2.1 Populations Exposures will be needed for a representative sample of the V.S. population as well as for a wide range of user-specified subpopulations (e.g., toddlers, the elderly, adolescents, females of childbearing age). The V.S. population can be represented by a subsample drawn from the U.S. Census. The V.S. Department of Agriculture (VSDA) has developed sample frames for the national food consumption surveys that are representative of the V.S. population and more than 25 subgroups of the population. Demographic data are also available for the VSDA reference population so that it can be used for nondietary assessments as well. The VSDA created the sample frames as a part of the two most recent national food consumption surveys. Each sample frame contains data on more than 10,000 individuals who were surveyed in the VSDA's Continuing Survey of Food Intakes by Individuals (CSFII; VSDA, 1992-1996). Each individual provided demographic information as well as detailed food consumption data for 2 or 3 days. The VSDA developed a statistical weight for each individual that can be applied to develop estimates of exposure for the V.S. popUlation. The demographic variables for each individual can be used in part to support the selection of parameters for aggregate and cumulative exposure assessments as well as to estimate exposure due to residues in the food supply.
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CHAPTER 17 Modeling Dietary Exposure
17.2.2.2 Exposure Period
Generally most models use the calendar day as the basic unit of time for calculating human exposure to one or more chemicals. All reporting periods longer than a day can then be built up from sequential daily exposures to an individual, summed, and averaged over the number of days included in the reporting period to provide an average daily exposure for that individual over the time duration specified in the analysis. The World Health Organization (WHO) through its expert committees, Joint FAOIWHO Expert Committee on Food Additives and Joint FoodlWHO Meeting on Pesticide Residues, often estimates weekly exposures for contaminants. Those weekly estimates are then compared to pesticides and weekly estimate of acceptable intake levels.
--1
CSFII
DemOgrap~
Match Demographic Parameters?
Iteration (j '" 110 N,)
17.2.2.3 Screening Models
Exposures can result from the treatment of foods, homes, schools, workplaces, yards, gardens, parks, golf courses, etc. Although it is theoretically possible that all of these locations could be treated with the same chemical, it is extremely unlikely that such a scenario would in fact occur. Nonetheless, in many instances a worst case estimate of exposure is computed by assuming that simultaneous exposures from all of the treatments actually occur. These are usually called screening estimates and have the advantage of being simple to conduct. More realistic estimates, incorporating the probability of exposure, are derived using more complicated, resource intensive approaches. A point estimate of exposure to a specific chemical in a particular population is a broad estimate generated using one number to represent the residue (i.e., concentration of the chemical) and one number to represent contact with the chemical by that population. In estimating exposure using point estimates, the arithmetic mean is most commonly used; however, if the distribution of parameter of interest is known to be skewed, use of the median (or 50th percentile) concentration is more appropriate (Mosteller and Tukey, 1977). Typically, the most basic models combine data on average contact and average concentration levels of the substance to estimate average exposure. 17.2.3 PROBABILISTIC OR MONTE CARLO MODELS
Probabilistic or Monte Carlo assessments utilize both contact level distributions (amount of food consumed, time of residence in the home, etc.) and residue distributions. Contact levels vary among individuals. Children typically consume more food per unit body weight than adults. Also, the total dermal contact for children may be larger than the dermal contact level for adults because children may have more active contact with the treated surface. Similarly, residue levels present in the residential environment also vary. Variations in the contact and chemical concentrations produce potential variations in the resulting exposure distributions (Fig. 17.3).
~------<
J= N,?
Figure 17.3 General sequence ofprobabilistic exposure analyses.
17.2.4 COMBINATION OF SCREENING AND PROBABILISTIC MODELS
Most analysts use a combination of worst case estimates for some parameters and more realistic estimates for other parameters. Typically more resources are devoted to refining the parameters that have the biggest impact on the exposure assessment. 17.2.5 CONTACT FACTORS FOR DIETARY EXPOSURE MODEL: FOOD CONSUMPTION
There have been numerous food consumption surveys conducted to estimate the intake of nutrients. Such surveys have been conducted at periodic intervals in many countries. The methods used to conduct dietary surveys are in a continuing state of development and refinement. The survey instruments and procedures to collect and analyze samples are continually being improved, and new compounds are being added to reflect current priorities. The optimal approach for assessing dietary exposures has been debated for years in the United States as well as in other countries (FASEB, 1988). The most appropriate survey for use in an exposure assessment must be determined. Considerations can include: • When were the data collected and are current dietary practices similar enough to be relevant?
17.2 Exposure Models
447
• Were appropriate populations and subgroups of the popUlation surveyed? • Were data collected during all seasons? • Were the foods of interest included in the survey? • Was the quantity of each food estimated?
availability is usually reported in terms of raw agricultural commodities. Processed forms of foods are usually not considered and there is no way to distinguish the use of foods as ingredients. Nonetheless, these data allow assessments to be conducted at the international level using comparable information.
The available information about the survey methodology and the resulting data available from a particular survey is an important determinant of the usefulness of the survey for estimating exposure using the data.
17.2.6.2 Household Inventories
17.2.6 COMMON TYPES OF FOOD CONSUMPTION SURVEYS There are four broad categories of food consumption data: (1) food supply surveys (market disappearance), (2) household or community inventories, (3) household food use data, and (4) individual food consumption surveys. 17.2.6.1 Food Supply Surveys Food supply surveys, also called food balance sheets (FBSs), are conducted on a countrywide basis each year by almost every country in the world. These surveys provide data on food availability or disappearance rather than actual food consumption, but may be used to estimate indirectly the amounts of foods consumed by the country's popUlation. Food supply data may be useful for setting priorities, analyzing trends, developing policy, and formulating food programs. For some countries, food supply data are the only accessible data that represent the country's food consumption. Although other types of data are generally better for risk assessment, food supply surveys can be particularly useful when there is a need to compare potential risks in different populations. The FBSs are useful for this because similar methods are used around the world. Food supply surveys also have been useful in some epidemiological studies (Sasaki and Kestelhoot, 1992). FBSs describe a country's food supply during a specified time period. Mean per capita availability of a food or commodity is calculated by dividing total availability of the food by the total population of the country. FBSs, published by the FAO, describe the food supply in countries on all continents. European FBSs are also prepared by the Organization for Economic Cooperation and Development and the Statistical Office of the European Communities. Food supply data in the United States are developed by the USDA's Economic Research Service. There are some limitations in the use of FBSs to estimate exposures. First, waste at the household and individual levels usually is not considered. Therefore, exposure estimates based on food supply data are higher than estimates based on actual food consumption survey data: the magnitude of the error depends on the quantity of waste produced. Perhaps more importantly for exposure assessment, users of foods cannot be distinguished from nonusers. Therefore, individual variations in exposure cannot be assessed, nor can exposure of potentially sensitive subpopulations be estimated. Finally, food
Household surveys generally can be categorized as (1) household or community inventories or (2) household or individual food use. Inventories are accounts of what foods are available in the household. What foods enter the household? Were they purchased, grown, or obtained some other way? What foods are used up by the household? Were they used by household members, guests, and/or tenants? Were they fed to animals? Inventories vary in the precision with which data are collected. Questionnaires mayor may not ask about forms of the food (i.e., canned, frozen, and fresh), source (i.e., grown, purchased, or provided through a food program), cost, or preparation. Quantities of foods may be inventoried as purchased, as grown, with inedible parts included or removed, as cooked, or as raw. Such data are available from many countries including Germany, the United Kingdom, Hungary, Poland, Greece, Belgium, Ireland, Luxembourg, Norway, and Spain (Trichopoulou and Pagona, 1997 and DAFNE 11,1995). 17.2.6.3 Household Food Use Data Food use studies, usually conducted at the household or family level, often are used to provide economic data for policy development and planning for feeding programs. Survey methods used include food accounts, inventories, records, and list recalls (Pao et ai., 1989). These methods account for all foods used in the home during the survey period. This includes foods used from what was on hand in the household at the beginning of the survey period and foods brought into the home during the survey period. Although household food use data have been used for a variety of purposes, including exposure assessment, serious limitations associated with data from these surveys should be noted. Food waste often is not accounted for. Food purchased and consumed outside the household mayor may not be considered. The household members who did and did not consume a particular food cannot be distinguished, and individual variation cannot be determined. Exposures by subpopulations based on age, gender, health status, and other variables for individuals can only be estimated based on standard proportions or equivalents for age/gender categories. 17.2.6.4 Individual Consumption Studies Individual exposure studies provide data on food consumption by specific individuals. Methods for assessing food exposures of individuals may be retrospective (e.g., 24-h or other shortterm recalls, food frequencies, and diet histories), prospective (e.g., food diaries, food records, or duplicate portions), or a combination thereof. The most commonly used studies are those that use the recall or record method and the food frequency method. For example, national dietary surveys were
448
CHAPTER 17
Modeling Dietary Exposure
conducted in Australia in 1983 for adults, in 1985 for school children, and in 1995 for the entire population. Similar studies are available for D.S. populations as well as for many European countries. Food consumption data are available from food frequency surveys for some populations (ANZFA, 1997). Food recall studies are used to collect information on foods consumed in the past. The unit of observation is the individual or the household. The subject is asked to recall what foods and beverages he or she or the household consumed during a specific period, usually the preceding 24 h. Because this method depends on memory, foods are quantified retrospectively, often with the aid of pictures, household measures, or two- or three-dimensional food models. Recalls have been used successfully with individuals as young as 6 years of age, and interviewer-administered recalls are usually the method employed for populations with limited literacy or for individuals whose native language is not English. When individuals are not available for an interview or are unable to be interviewed due to age, infirmity, or temporary absence from the household, surrogate respondents often are used (Samet, 1989). The main disadvantage of the recall method is the potential for error due to faulty memory of respondents. Items that were consumed may be forgotten or the respondent may recall items consumed that actually were not consumed during the time investigated. To aid recall memories, the interviewer may probe for certain foods or beverages that are frequently forgotten, but this probing also has been shown to introduce potential bias by encouraging reporting of items not actually consumed. Food record/diary surveys collect information about current food exposure by having the subject keep a record of foods and beverages as they are consumed during a specific period. Quantities of foods and beverages consumed are entered in the record usually after weighing, measuring, or recording package sizes. Occasionally, photographs or other recording devices are employed. The results of surveys conducted using short-term recalls or food diaries are generally the most useful source of data to estimate exposure to pesticides. Data from these surveys can be used to estimate either acute or chronic exposure. Averages and distributions can be calculated, and exposure estimates can be calculated for subpopulations based on age, gender, ethnic background, socioeconomic status, and other demographic variables, provided that such information is collected for each individual. Food frequency questionnaire (FFQ) surveys typically allow qualitative estimates of exposure. A FFQ or checklist is used to determine the frequency of consumption of the foods of interest. Subjects indicate how many times each day, week, or month they usually consume each food. Occasionally a semiquantitative FFQ survey is utilized that estimates the amounts consumed by having respondents indicate whether their usual portion size is small, medium, or large. For the D.S. population, the data collected in one of two large, national, food-consumption surveys conducted by the D.S. government, (1) the Nationwide Food Consumption Survey conducted by the DSDA, beginning in 1935, or (2) the National Health and Nutrition Examination Survey undertaken
by the D.S. Department of Health and Human Services, beginning in 1971, is appropriate for most exposure assessments. Both surveys employ multi stage area probability sampling procedures to obtain a sample that is representative of the population. The surveys are repeated at periodic intervals and are publicly available. 17.2.7 RESIDDE FACTORS FOR DIETARY EXPOSURE MODEL
Estimates of the residues in foods and other media can be made using a single fixed (deterministic) value by using empirical distributions of residues or using common statistical distributions as a representation of the residues. Often an exposure assessment combines single fixed values for some foods and distributions for others. Where a distribution is created from other distributions, Monte Carlo sampling is used to choose a value from each distribution independently from one another and then these values are combined (e.g., multiplied or added together, depending on the nature of the built-up composite function as specified by the user). In other situations, the residues may need to be handled as dependent variables. Residue data can come from a variety of sources, including controlled research studies, and government and industry monitoring programs. In many cases, additional information about the proportion of the crop that could be treated, pest pressures, etc. may improve the exposure assessment. There are three common approaches that are appropriate for collecting data on concentrations of chemicals in food: (1) duplicate diets, (2) market basket or representative sampling, and (3) controlled experimentation. The DSDA monitors residue levels in selected fruits and vegetables, and meat and pOUltry products. The Food and Drug Administration monitors residue levels in all other foods. California, Florida, and a number of other states have monitoring programs. Depending on the specific D.S. monitoring program, foods or commodities may be sampled at the point of entry to the country, at the farm gate, at the food processing plant, or at the retail level. Controlled experiments are sometimes used to determine the likely levels of pesticides in foods produced under specified conditions and of relevant metabolites. Field trials are conducted for virtually all pesticides. Data from these trials are considered to be representative of the extreme upper limit of potential residues. Analyses of foods collected in supermarkets are more likely to be representative of the residues encountered by consumers. 17.2.8 CONSIDERATIONS FOR CUMULATIVE EXPOSURE ASSESSMENTS
The methods used to estimate exposure to multiple chemicals need to adjust the detected residue levels of each of the chemicals considered by "relative toxicity factors" that reflect the toxicity levels of these chemicals relative to a "standard" chemical. A total adjusted residue then may be derived for each
17.2 Exposure Models
sample by summing the adjusted residue values that correspond to that sample. An exposure assessment is then conducted using these total adjusted residues. This approach is based on concepts proposed by the National Academy of Sciences for the assessment of joint exposure to organophosphate pesticides and is similar to that followed by the Environmental Protection Agency (EPA) in the case of dioxin-like compounds. This adjustment can be done outside of the exposure assessment model and a single residue can be entered into the analysis. However, generally some technique is needed to allow a determination of the relative contribution of each chemical to the exposure assessment.
• Regions of treatment (e.g., treatment in Florida occurs all year round, foods from Latin America are untreated). • Type of applicator (e.g., the product is sold only to homeowners). • Frequency of application, typical preharvest intervals (e.g., the application occurs once a month). • Exposureibiomonitoring studies: In some cases, the exposure to a specific pesticide is already known and the data can be used to directly estimate exposure. Such data may be point estimates, or empirical or parametric distributions of exposure. 17.2.10 ESTIMATING AGGREGATE AND CUMULATIVE EXPOSURES USING A CALENDAR MODEL
17.2.9 CONSIDERATION FOR AGGREGATE EXPOSURE ASSESSMENTS
Typically more realistic estimates are required for aggregate exposure assessments. Worst case approaches compile too many conservative values to provide any meaningful evaluation. Types of information that need to be considered include applicator and postapplication exposures that incorporate information about the likelihood that each treatment and contact occur. Probabilistic methods are used to determine whether a given pesticide is applied in a specific household, the application dates of pesticide treatments in the home, the amount of residue uptake per unit of contact, the level of contact by each individual that results in the uptake of physical chemical residues, and other relevant parameters. Chemical exposure by each individual in the sample population realistically can be estimated repeatedly using Monte Carlo analysis that specifies a new set of daily food consumption data, treatment schedules, contact schedules, and residue concentrations with each iteration. There is a variety of data that can be used to do a better job of estimating aggregate exposure that includes:
A calendar model can be used to account for the variation in application; not all products are applied on the same day or used with the same frequency. Information about the frequency of use or application is very important in accounting for this type of information in the model. This information may be obtained from market use data product labels that specify frequency of application or from reasonable seasonal assumptions that are made based on professional experience. It is important to recognize that residues from various treatments might overlap. For example, a professional applicator may treat a home for cockroaches on March 1. On June 1, the homeowner may spray the lawn with the same active ingredient, and foods may be treated at different times during the growing season. The family dog may be treated for fleas on August 1 using the same active ingredient. The question might then arise, "What is the exposure to a child on September 7?" A realistic model will not assume the worst case for each of these scenarios and simply add the residues together. Instead, a good model will account for such overlap based on the probability of occurrence of such overlap and chemical degradation. Figure 17.4 presents a graphical representation of how a calendar model determines the avail-
• Percentage of households treated (e.g., 75% of the households in the United States use flea and tick products).
z(.)
c'"o (.)
Treatrrent 2
"0
Treatrrent 3
90
100
180
250
270
Days Day 250 • Residue - R2
Figure 17.4
449
+ R3
Available residue for one application scenario of three treatment types with chemical X in I year.
365
450
CHAPTER 17
Mode1ing Dietary Exposure
able residue per unit of contact on September 7 (day 2S0 of the year) from the three treatments previously described. In summary, a calendar model: • Uses the probability that individual exposures occurs around specific dates. • Calculates exposure for individual chemical uses and exposure routes. • Combines the exposure-probability distributions for individual uses using Monte Carlo sampling techniques. The calendar model is able to estimate the available residue value on the day of exposure by first computing the number of elapsed days between the application day and the day of exposure for each use of each chemical. A degradation function can then be used to adjust the levels of chemical residues present. This approach is often not feasible to estimate food residues, and assumptions must be made about which residues are present at a given time. Estimates of residues by season are available for many chemical and crop combinations. For each day of the year, the calendar model can combine exposure (contact) distributions with the probability that an exposure to a given compound occurs as a result of a previous or concurrent application of a product containing that chemical. The model also can take into account the probability that exposures to more than one product may occur on a single day, which provides a more realistic exposure estimate than if exposures from single uses are summed. Monte Carlo techniques are used to estimate the distribution of potential exposures (contacts) and to combine these distributions with information about product use. These conditional distributions are combined with the usage probabilities associated with each product to generate exposure distributions for specific calendar days specified by the user and are repeated many thousands of times in the analysis.
17.3 SAMPLE MODEL CALCULATION METHODOLOGIES To provide better guidance, this section presents a representative set of the methodologies and algorithms. The analyst will want to select the exposure methodology that is most appropriate for the desired analysis. 17.3.1 TYPICAL EXPOSURE METHODOLOGY SEQUENCE Prior to the analysis: 1. Define product and route-specific parameters for each source of the pesticide. Each exposure route (dermal, oral, or inhalation) of concern for each treatment type (e.g., treatment of agricultural crops, pet treatment, turf treatment, crack and crevice treatment) must be identified. If the scenario represents different exposure routes for the same treatment type, the model should have the same set of application scenarios, but will likely utilize different data. 2. Determine all treatments of interest (e.g., crops that are treated for dietary assessment and all nondietary uses of the chemical for aggregate assessment). 3. Select the most appropriate software to use to determine exposure. 4. Identify available contact and residue data to use in the assessment (see Fig. 17.S).
CSFfl Demographics
CSFII Food Consumption
I DEEM Re,idue File
17.2.11 TOXICITY DATA USED TO PLACE THE EXPOSURE ASSESSMENT IN CONTEXT
Day (2),
Day (NO)
Dietary
~ Treatment (1)
C~ntact ~GX(1) lIbrary-----
-------+
Inhalation
AGX (2) ----------.
-+
Residue
Ideally exposure estimates are compared to compound-specific toxicity measures to derive risk estimates. The most appropriate toxicity measure depends on the type of assessment being conducted. Examples of toxicity values are dermal, oral, and inhalation no observed effect levels (NOELs), effective doses (EDSO), or reference doses (RIDs). Different risk estimates may be needed to place the compounds in proper perspective for either aggregate or cumulative assessments than when a compound is evaluated individually. Estimates of chronic exposure are typically compared to a chronic reference dose and may be expressed as a percentage of that reference dose. Estimates of acute exposures are usually compared to the acute NOEL. Acute risk estimates are expressed as margin of exposures (MOEs). Estimates of short- and intermediate-term exposures are usually compared to the short- and intermediateterm NOELs, respectively.
Day(1),
Dermal
AGXx (3)------.
Oral
+'____________~
Cb,,,,>
Treatment (2)
C,~ntact ( G X (4) Library--+-
-------+
AGX (5) -----to-
Residue Library
Co, OI,aCI
-+
Llbrary----t-
AGX (6)
--------+
<
AGX (N-2)
Inhalation Dermal
Treatment (NT)
---+
AGX (N-1) ------to>
Residue Library
-+
AGX (N)
-
Oral
-----+
Inhalation Dermal Oral
Summarize and Bin Results
Figure 17.5 Compute exposure.
17.3 Sample Model Calculation Methodologies
Conduct the analysis: 5. Define analysis parameters. Typical parameters to be defined are the type of analysis of interest (given in the following sublist), the number of Monte Carlo iterations and random number seed, the appropriate NOEL or RID values by exposure route, and the subpopulations of interest (e.g., toddlers 1-3, southern United States, etc.). Typical analysis types include: • Single Day. In this analysis, exposure is calculated for a randomly chosen day where all days of the year are given equal probability Exposure also may be calculated for specific sequential days if that is appropriate. Typically the output provides a distribution of the daily exposures for each of the specified days. • Multiple Weeks. Exposure is calculated for a specified week or combined set of weeks (up to 52 combined weeks) that are chosen by the user. The output can provide a distribution of the daily exposures averaged over the combined number of days. Exposure also can be calculated for a series of weeks or a series of a combined set of weeks. 6. Dietary exposure is usually an important source of exposure and the analyst may conduct the dietary exposure separately prior to conducting aggregate exposure assessments. 7. Usually exposure calculations are repeated for each treatment or exposure, and the total exposure is calculated and stored for later use to estimate a distribution and to estimate exposure by route (dermal, inhalation, oral). For all analyses: 8. The average daily exposure (by route) needs to be determined. One approach is to calculate the average daily exposure by dividing the sums of the multiple-day exposures by the number of days (n) specified in step 3. 9. Express exposure on a unit body weight (BW) basis. 10. For cumulative exposures, determine how relative potency will be expressed and adjust residues to reflect the differences. 17.3.2 MONTE CARLO SIMULATION METHODOLOGY Monte Carlo analysis methods are used to bring together the wide range of probability distributions needed to calculate each individual's exposure. Specifically, for each treatment type (e.g., agricultural crop treatments, residential turf treatments, etc.) included in the analysis, Monte Carlo analysis can be used to define the use scenarios. For example, Monte Carlo sampling could be used to decide (1) whether the pesticide is applied to each crop, (2) whether it is applied professionally or by someone in the home, (3) whether that treatment type is made at the
451
individual's residence, (4) how many time a year the treatment is made, (5) the dates that each treatment is made during the year, (6) what residue amount results from each treatment on the day(s) of treatment, and (7) how much contact with potentially residue-bearing surfaces (or other residue transfer media) is made by the individual on the days of interest. Ideally the analyst determines whether each of these decision variables is to be based on a deterministic value or on a distribution from which a value is drawn using a new random number with each draw. Often food consumption amounts used in dietary analyses are treated as deterministic values, whereas the residues assigned to those food consumption amounts can be either deterministic or drawn from a distribution. In addition, most, if not all, demographic parameters associated with each respondent are considered to be deterministic. That is, the person's reported weight and height are used in the analysis, not simulated using a probabilistic distribution. 17.3.3 DEGRADATION CALCULATION METHODOLOGY The degradation calculation model incorporates residue degradation by using either actual residue values on specific days after treatment or using degradation equations. If market basket data are available, degradation equations are not required to estimate dietary exposure. Degradation equations are important for most nondietary evaluations. The degradation equations are as follows: Half-life method, Rx
=
x h1 Ro(0.5 / )
(17.2)
where hI denotes half-life in days. Straight-line method, Rx
=
Ro(l - X/z)
(17.3)
where z represents days to zero residue level. Rx is determined directly from a residue distribution function other than Ro: Given that ldi is the last day that residue function i is valid, find Ri such that ld(i - 1) < X <= ld(i). Then Rx = Ri (selected from probability distribution) where X is the integer number of days elapsed between application and contact, Ro is the residue concentration factor on day of application (mg/unit of contact), and Rx is the residue concentration factor on day X (mg/unit of contact). Note: When X = 0, Rx = Ro with no adjustment. That is, on the day of application, the treatment was assumed to be applied before contact and the residue is not degraded. 17.3.4 RISK (MOE) CALCULATION METHODOLOGY When a single exposure route is evaluated for a chemical or multiple routes are evaluated in which the NOEL for each route is the same, the margin of exposure (MOE) can be calculated
452
CHAPTER 17
Modeling Dietary Exposure
for any exposure value in the resulting distribution. For example, if the 90th percentile exposure amount is EX90, then the MOE is simply NOELlEx90. However, when multiple exposure routes that each have a different NOEL are evaluated, the total MOE cannot be calculated directly for any given exposure in the exposure distribution. Instead, the MOE must be calculated for each individual from the individual components of his or her total exposure, that is, MOEl = ExpdNOELl, MOEl
= EXP3/NOEL3
where subscript 1 refers to the inhalation exposure route, 2 to the dermal exposure route, and 3 to the oral exposure route (which includes both dietary and incidental ingestion). Then the total MOE (MOET) can be calculated based on the equation in the Office of Pesticide Programs draft Guidance for Performing Aggregate Exposure and Risk Assessments (EPA, 1999b), which is
MOET = 1/
(M~El + M~E2 + M~En)
(17.4)
where, for example, MOEl denotes a given margin of exposure (e.g., adult, inhalation route) and MOE2 denotes another specified margin of exposure (e.g., adult, dermal route). This total MOE (MOET) concept was presented to and endorsed by FIFRA's SAP in 1997 (EPA, 1999b).
17.4 UNCERTAINTY The EPA (1992) has classified uncertainty in exposure assessments in three categories scenario uncertainty, parameter uncertainty, and model uncertainty. Examples of how these uncertainties may arise in an exposure assessment follow.
exposures often include very few measurements, and typically are conducted for a limited number of scenarios. 17.4.3 MODEL UNCERTAINTY A comparison of the results of the various analyses provides the assessor with a measure of the impact of the uncertainty in the exposure model used. Another example of model uncertainty may result from using the wrong model to represent the degradation, over time, in air (or soil) concentrations.
17.5 EXAMPLE OF AGGREGATE EXPOSURE 17.5.1 DESCRIPTION This example illustrates an approach that is used to aggregate exposures for adults (18 + years old) and toddlers (1-3 years old) exposed to residues of chemical X from a hypothetical turf product that is also applied to fruit crops. The product is a granular formulation used to control weeds and is applied using a push type spreader. It was assumed that the homeowner makes five applications of this product at 4-week intervals, with the first application in the first week of May. Therefore, the
Table 17.1 Results of the Turf and Dietary Aggregate Example: Single Day (randomly selected) Exposures (mg/kg BW/day; per capita)
Adults Percentile
Exposure
Dermal (applicator
17.4.1 SCENARIO UNCERTAINTY Scenario uncertainties include descriptive errors, aggregation errors, and incomplete analysis. For instance, for residues on imported crops, scenario uncertainty may result from incorrect information regarding the regions in which the product is used and how it is used.
99.9 99 95 90
+ postapplication)
0.459265 0.316427 0.199563 0.140733
Percentile
0.090714 0.000000 0.000000 0.000000
99.9 99 95 90
220 N/A N/A N/A
99.9 99 95 90
0.004861 0.001313 0.000182 0.000029
Aggregate (dermal
99.9 99 95 90
MOE
4,114 15,238 109,690 681,860
99.9 99 95 90
0.795279 0.539392 0.312677 0.205951
107 158 272 413
Oral (nondietary)
0.029207 0.020774 0.011949 0.008066
685 963 1,674 2,479
Oral (dietary)
Oral (dietary)
99.9 99 95 90
Exposure
Dermal (postapplication)
185 269 426 604
Inhalation (applicator)
99.9 99 95 90
17.4.2 PARAMETER UNCERTAINTY Parameter uncertainty includes measurement errors, sampling errors, variability, and use of surrogate data. Two examples of measurement uncertainty in the data may be the presumed tendency of some survey respondents to underestimate their body weights or to underreport food consumption. In the first example, parameter uncertainty may result in potential overestimation of the exposures, whereas in the second example, it may result in potential underestimation of exposures. Sampling errors may result from sampling too few observations or nonrepresentative sampling. Generally, studies of residential
Toddlers MOE
+ inhalation + oral) 190 280 440 620
0.046856 0.018387 0.005180 0.001630
Aggregate (dermal
99.9 99 95 90
427 1,088 3,861 12,266
+ oral) 100 150 250 360
17.5 Example of Aggregate Exposure Table 17.2
Results of the Turf and Dietary Aggregate Example: 4-Week Exposures (mg/kg BW/day; per capita)a
Percentile
Adults Exposure
MOE
Dermal (applicator + postapplication) 0.247649 343 99.9 0.206313 412 99 0.170271 499 95 99.9 99 95
Inhalation (applicator) 0.005990 3,339 0.004656 4,296 0.003526 5,673
99.9 99 95
Oral (dietary) 0.001410 0.000649 0.000273
14,185 30,801 73,202
Aggregate (dermal + inhalation + oral) 99.9 310 99 375 95 457
Toddlers Exposure Percentile
PDR = potential dose rate (mg/kg per day)
= unit exposure (dermal or inhalation) (mg/lb ai) Point estimate: Pesticide Handlers Exposure
MOE
Database (Versar, 1995) UEdermal
= 3 mg/lb ai; UEinhalation = 6.3 mg/lb ai
AR = application rate (lb ai/acre) Hypothetical uniform distribution: 0.5-1.5 lb ai/acre A = Area treated (acre/day)
Oral (nondietary) 99.9 0.015054 0.012200 99 95 0.009980
1329 1639 2004
Oral (dietary) 0.012927 0.007424 0.003905
1547 2694 5121
99.9 99 95
where
UE
Dermal (postapplication) 0.386618 99.9 220 0.310551 99 274 95 0.256527 331
453
Aggregate (dermal + oral) 99.9 180 226 99 95 276
aOnl y the
results from the 4-week time period (weeks 22-25) with the highest exposure values are presented here.
Hypothetical uniform distribution: 0.25-1.0 acre/day BW
= Body weight (kg) Empirical distribution: CSFII data
17.5.3 ADULT AND TODDLER POSTAPPLICATION DERMAL EXPOSURES
PDR
DFR x TC x ET x CF1 =------BW
(17.6a)
where PDR = potential dose rate (mg/kg per day) DFR = dislodgeable foliar residue (J1.g/cm 2)
potential exposures that result from the use of this product include dermal and inhalation adult applicator exposures, dermal postapplication exposure for adults and toddlers, and toddler incidental ingestion via hand-to-mouth behavior. For simplicity, we did not include the other oral exposures as specified in the EPA residential (SOPs; EPA, 1999a) for the toddlers. Chemical X was assumed to be completely degraded after 35 days. It was also assumed that the toxic effects from the dermal, oral, and inhalation routes are the same; therefore, the exposures from these routes were aggregated together. The dermal NOEL was assumed to be 85 mg/kg BW/day, whereas the inhalation and oral NOELs were 20 mg/kg BW/day. For simplicity it was assumed that the short and intermediateterm NOELs are the same. Novigen Sciences, Inc., Calendex™ software was used to calculate exposure. The exposures were estimated for a single randomly chosen day (Table 17.1) and for five 4-week periods (weeks 18-21; weeks 22-25; weeks 26-29; weeks 30-33; and weeks 34-37; Table 17.2). The EPA SOP (EPA, 1997a) equations and parameters, which are presented subsequently, were used as the basis for the nondietary exposure algorithms. The distribution types and sources of data are listed below each parameter.
See Eq. (17.6b) TC = transfer coefficient (cm 2jh) Point estimates: EPA SOPs Adult short term
Toddler short term = 5200 Toddler intermediate term
= 2600
ET = exposure time (h/day) Cumulative distribution of amount of time spent playing on grass: EPA Exposure Factors Handbook (EPA, 1997b) Adults Percentile 0
CFI
17.5.2 ADULT APPLICATOR DERMAL AND INHALATION EXPOSURES
= 14,500
Adult intermediate term = 7300
Toddlers ET 0
Percentile 0
ET 0
25
0.5
25
0.25
50
2.0
50
1.0
100
2.0
75
2.0
100
2.0
= conversion factor (0.001 mg/J1.g)
BW = body weight (kg) Empirical distribution: CSFII data
PDR = UE(dermal or inhalation) x AR BW
(17.5)
DFR = AR x F x CF2 x CF3
(17.6b)
454
CHAPTER 17
Modeling Dietary Exposure
where DFR AR
= dislodgeable foliar residue (!l-g/cm2)
= application rate (lb ai/acre)
Hypothetical uniform distribution: 0.5-1.5 lb ai/acre F = Fraction of ai transferred from foliage (%) Point estimate: EPA SOPs, 5% CF2 = Conversion factor (4.54E8 !l-g/lb) CF3 = Conversion factor (24.7E-9 acre/cm2) 17.5.4 TODDLER POSTAPPLICATION INCIDENTAL INGESTION EXPOSURES
PDR
= _D_F_R_x_SA_x_FQ-=-x_E_T_x_SE_x_C_F_1 BW
The exposure and risk estimates should be presented for the full range of the population. For an individual treatment (e.g., treatment of apples or lawn treatment or only pet treatment), exposure and risk estimates are presented for each individual route (e.g., dermal, inhalation, incidental ingestion) as well as any aggregated routes (e.g., inhalation and incidental ingestion combined). For an aggregate assessment (e.g., lawn treatment + pet treatment + dietary), exposure and risk estimates should be presented for the individual routes, but the estimates are aggregated by uses (e.g., the dermal estimate would be an aggregation of the dermal from lawn care plus the dermal from pet treatment). Additionally, if there were similar toxicological endpoints from both dermal and oral exposure, then these estimates could be combined as well. In that case, there could be a total aggregation.
(17.7)
where
17.6 QUALITY AUDIT AND VALIDATION
PDR = Potential dose rate (mglkg per day) DFR = Dislodgeable foliar residue (fLg/cm 2 ) See Eq. (17.6b) SA = Surface area of two fingers (cm2/event) Point estimate: EPA SOPs, 20 cm 2 FQ
= Frequency of hand-to-mouth activity (eventslh)
Point estimate: EPA SOPs, 20 eventslh ET = Exposure time (h/day) Cumulative distribution of amount of time spent playing on grass: Exposure Factors Handbook Toddlers Percentile
ET
0 25
0 0.25
50 75 100
1.0 2.0 2.0
SE = Saliva extraction (%) Point estimate: EPA SOPs, 50% CF1 = Conversion factor (0.001 mg/!l-g) BW = Body weight (kg) Empirical distribution: CSFII data 17.5.5 REPORTING RESULTS
A comprehensive report of each exposure analysis should always be prepared to fully document the assumptions and data that were used and to link that information to the actual analysis. The report should include estimates of exposure and risk (presented as either an MOE, %NOEL, %RfD, or other toxicity parameter) for each subpopulation identified by the analyst.
Audits of the computational algorithms used in an assessment must be derived using an independent spreadsheet calculation or other software. These include the algorithms for deriving the interval limits, allocation of the observations to the intervals, and calculation of the various statistics, including the means, standard deviations, and percentile estimates.
REFERENCES Australia New Zealand Food Authority (ANZFA) (1997). "Dietary Modelling: Principles and Procedures." Australia New Zealand Food Authority, Canberra, Australia. Data Food Networking II (DAFNE) (1995). "Newtwork for the Pan-European Food Data Bank Based on Household Budget Surveys." Federation of American Societies of Experimental Biology (FASEB) (1988). "Estimation of Exposure to Substances in the Food Supply." Life Sciences Research Office, Bethesda, MD. Hill, R. H., Head, S. L., Baker, S., Gregg, M., Shealy, D. B., Bailey, S. L., Williams, C C, Sampson, E. J., and Needham, L. L. (1995). Pesticide residues in urine of adults living in the United States: Reference range concentrations. Environ. Res. 71,99-108. Mosteller, E, and Tukey, J. W. (1977). "Data Analysis and Regression." Addison-Wesley, Reading, MA. Pao, E. M., Sykes, K. E., and Cypel, Y. S. (1989). "USDA Methodological Research for Large-Scale Dietary Intake Surveys. 1975-88." Home Economics Research Report No. 49. U.S. Department of Agriculture, Human Nutritition Information Service, Washington, DC Samet, S. M. (1989). Surrogate measures of dietary intake. Am. 1. Clin. Nutr. 50, 1139. Sasaki, S., and Kestelhoot, H. (1992). Value of Food and Agriculture Organization data on food-balance sheets as a data source for dietary fat intake in epidemiologic studies. Am. 1. Clin. Nutr. 56,716. Shurdut, B. A., Barraj, L., and Francis, M. (1998). Aggregate exposures under the Food Quality Protection Act: An approach using chlorpyrifos. Regul. Toxico!. Pharmacol. 28(2),165-177. Trichopoulou, A., and Pagona, L., eds. (1997). "Methodology for the Exploitation of HBS Food Data and Results on Food Availability in 5 European Countries." DAFNE (Data Food Networking) European Communities. United States Department of Agriculture (USDA) (1992). Nationwide Food Consumption Survey: Continuing Survey of Food Intakes by Individuals 1989-90. Dataset, Human Nutrition Information Service.
References
USDA (1993). Nationwide Food Consumption Survey: Continuing Survey of Food Intakes by Individuals 1990--91. Dataset, Human Nutrition Information Service. USDA (1994). Nationwide Food Consumption Survey: Continuing Survey of Food Intakes by Individuals 1991-92. Dataset, Human Nutrition Information Service. USDA (1995). Nationwide Food Consumption Survey: Continuing Survey of Food Intakes by Individuals 1993-94. Dataset, Food Survey Research Group, Agricultural Research Service. USDA (1996). Nationwide Food Consumption Survey: Continuing Survey of Food Intakes by Individuals 1994-95. Dataset, Food Survey Research Group, Agricultural Research Service. United States Environmental Protection Agency (EPA) (1997a). "Standard Operating Procedures for Residential Exposure Assessments." Residential
455
Exposure Assessment Work Group, Office of Pesticide Programs, Health Effects Division, EPA, Washington, DC. (Draft version dated December 18, 1997.) EPA (1997b). "EPA Exposure Factors Handbook," Vol. I-Ill. Office of Research and Development, EPA, Washington, DC. EPA(1992). EPA (1999a). "Standard Operating Procedures for Residential Exposure Assessments-Proposed Changes Presented to Science Advisory Panel." Office of Pesticide Programs, Health Effects Division, EPA, Washington, DC. EPA (1999b). "Guidance for Performing Aggregate Exposure and Risk Assessment." Office of Pesticide Programs, EPA, Washington, DC. Versar (1995). "PHED: The Pesticide Handlers Exposure Database," Reference Manual, Version 1.1. Prepared for the PHED Task Force that represente Health Canada, USEPA, and American Crop Protection Association.
CHAPTER
18 Greenhouse and Mushroom House Exposure Joop J. van Hernrnen, Derk H. Brouwer, and Johan S. de Cock Department of Chemical Exposure Assessment, TNO Chemistry
18.1 INTRODUCTION The enclosed environment of greenhouses and mushroom houses is unique with respect to the behavior of pesticide aerosols in air. Mixing and loading of pesticide formulations, especially powders, is normally not done in an enclosed environment, apart from closed systems and water-soluble bags. Small droplets in the outdoor environment will easily drift away from the worker. In enclosed environments the operator (applying the pesticide to the crop) cannot use wind direction to avoid the spray mist. However, Methner and Fenske (1994) showed that forced ventilation during application enabled experienced operators to reduce exposure, compared to nonexperienced operators in situations without ventilation. On the other hand, indoor sprays may have finer droplets than outdoor sprays in view of drift properties. This may have considerable effects on the level of exposure for the mixerlloader and applicator. The crops in indoor environments are not directly in contact with sunshine and rain. This may lead to differences in dissipation due to differences in photodegradation, hydrolysis, uptake of the pesticides by the crop, etc. Dissipation of the pesticide as a function of time is an important factor in determining the level of exposure, depending on the interval between application and reentry activities, such as harvesting. In this chapter the literature on occupational exposure to pesticides will be reviewed with respect to relevant exposure data for enclosed environments. Such data are important for the evaluation of the labor conditions in specific greenhouses for specific crops, as well as for registration of pesticides. The workers who will be considered are the operators who handle and apply pesticides and the workers who have contact with the treated crop or enter the enclosed environment after the application. The resulting potential exposure is important for generic use in registration procedures. It is also a good estimate for the potential risks in the work environment. Potential exposure is determined by the chemical and physical properties of the pesticide, the physical parameters related to the type of crop, and the nature of the operations of the worker, which for generic use Handbook of Pesticide Toxicology Volume I. Principles
can be expressed in an exposure scenario. The concept of potential exposure is highlighted by the fact that the actual clothing of the worker is not taken into account. This is also a major disadvantage since for assessment of systemic uptake of the pesticide through the skin, by inhalation, or by swallowing, the actual exposure and degree of absorption must be known. For dermal exposure the actual exposure to the skin is determined by the degree of penetration of the clothing. For inhalation the actual exposure is determined by the degree of inspiration of the particles involved and by interception of respiratory protective equipment as a function of the potential exposure. For oral exposure no measures are yet available. The most elaborate method for the determination of the absorbed dose is biological monitoring. The method requires a basic knowledge of the pharmacokinetics of the active ingredient (preferably for humans) and is therefore compound-specific. For risk assessment, an estimate of absorbed dose is the most relevant parameter for exposure. However, few biological exposure indices or other comparable measures are available at present. This may be one of several reasons that biological monitoring is, as yet, infrequently used in field studies for the estimation of risk. The characteristics of greenhouses vary appreciably throughout the world. In several countries in (sub)tropical climates, many greenhouses will be relatively small with plastic ceilings. Some will be open at the end. In colder climates the majority of the greenhouses will be all glass, and some will have mechanical ventilation. Mechanical ventilation will occur mainly in (sub)tropical climates or larger greenhouses. Most larger greenhouses are all glass. Their heights however may vary appreciably. The size generally is several hectares although compartmentilization may occur. These variations and especially the size of the greenhouse will affect the choice of an application technique. Generally hand-held equipment is used for the smaller surface areas, whereas in larger greenhouses more automated application techniques will be used.
457
Copyright © 200 1 by Academic Press. All rights of reproduction in any form reserved.
458
CHAPTER 18
Greenhouse
Variation in exposure of crop workers will to a lesser extent depend on the type of greenhouse, although the duration of the work may be a more prominent factor. Furthermore, inhalation exposure may vary considerably when open and closed greenhouses are compared. However, exposure studies are mainly done in larger, closed greenhouses; therefore no emphasis can be given to these variables. With respect to enclosed mushroom houses there is a tendency toward increased size in recent years. In the Netherlands there are presently about 700 mushroom houses with an average cultivation area of about 0.15 ha. Mushroom houses generally contain cells of 200-400 m2 with two or three scaffolds. Each scaffold contains several beds with mushrooms. The growing of mushrooms occurs year-round. Harvesting can be done manually or mechanically. Techniques for applying pesticides are similar to the techniques used in greenhouses, although the confined spaces preclude the use of some techniques. Special techniques involve spray lorries and drencher systems. With the low-volume techniques spraying time per cell requires minutes and may be somewhat longer when high-volume techniques are used.
18.2 EXPOSURE ASSESSMENT The exposure variables will not be discussed in any detail here, emphasis will only be put on the implications for sampling methods and strategies.
18.2.1 RESPIRATORY EXPOSURE MONITORING
Because enclosed environmental conditions in greenhouses and mushroom houses reduce drift, specific application techniques are used, some of which may generate relatively small droplets (Table 18.1). The lifetime of the droplets is heavily influenced by the greenhouse environmental conditions, such as high humidity and temperature. In greenhouses, relatively small droplets will be available for respiratory exposure. In addition, together with environmental conditions, the physicochemical properties of the pesticide formulation, for example, solubility and saturated vapor pressure, will determine the partition of the mass concentration of an aerosol between the liquid or solid phase and the vapor phase (Perez and Soderholm, 1991). Therefore, respiratory exposure monitoring methods should enable sampling of mixed-phase aerosols and the sampling head must match the specifications of the American Conference of Governmental Industrial Hygienists (ACGIH, 1985). Traditionally, respiratory exposure sampling of pesticides in greenhouses has focused on vapor trapping and impingers (Kangas et aI., 1993), solid adsorbent tubes (Stamper et aI., 1988; Waldron, 1985), or filter impinger combinations (Nilsson, 1995). Recently, a sampling method has been described that is considered to meet the criteria for mixed-phase sampling (Brouwer et al., 1994a). In an OECD protocol an IOM sampling head with a 25-mm filter, connected to a personal sampler pump with a 2 Vmin flow rate is suggested (OECD, 1997). Respiratory exposure monitoring methods during mixing and loading should be consistent with the formulation used.
Table 18.1 Overview of Application Techniquesa
Volume rate Method
(iiters/ha)
High-volume
>500
Technique Spray pistol
Hand-held
Spray lance
hydraulic
Droplet
Application
diameter
speed
(VMD; flm)
(halh)
Reference
85-100
0.2-D.6
Lindquist et aI., 1993
n.a.
n.a.
Knapsack Medium-volume
200-500
Low-volume
50-200
Spray tree
Semiautomatic
Mist blower
Automatic
15-25
0.4
Brouwer et aI., 1992c
Thermal pulse
(Semi)
<15
2
Lindquist et al., 1993
fogger
automatic
Air-assisted rotary
Semiautomatic
10-100
n.a.
Lindquist et aI., 1993
(Semi)
30
0.4-1
Lindquist et aI., 1993
n.a.
n.a.
25-30
0.2-D.6
disc mister Cold fogger
automatic Ultra-Iow-volume
<50
Spray cans
Manual automatic
Dusting Electrostatic an.a., no data available; VMD, volume median diameter.
Brouwer et aI., 1992c Lindquist et aI., 1993
18.3 Exposure of Operators
Solid formulations, that is, wettable powders, dry ftowables, granules, and field strength dusts, are considered to generate airborne particles, which allows trapping of the solid-phase aerosols by sampling heads. Because a considerable part of all formulations used for greenhouse applications are liquids, for example, emulsifiable concentrates and suspension concentrates, and solvent evaporation may enhance release of the active substance as well, vapor sampling will be necessary in some cases. 18.2.2 DERMAL EXPOSURE MONITORING
The pattern of dermal exposure may determine which dermal exposure monitoring method is preferred. During mixing and loading of liquid formulations dermal exposure, that is, predominantly exposure of the hands, may result from transfer from contaminated surfaces from previous loadings, liquid splashes during pouring and direct liquid contacts during (manual) mixing. This may limit the use of surrogate skin samplers, such as patches and cotton gloves, since absorption of liquids by the fabrics may overestimate exposure. Similar limitations of the surrogate skin techniques are expected for manual application where the applicator's hands are close to the spray nozzles (spray pistol) and leakages are likely to occur or for application to tall crops with high plant densities in combination with narrow walks between the beds. Generally, whole body methods to assess dermal exposure are preferable to patches. For operators in greenhouses and mushroom houses this may be even more important, since exposure for these workers is likely to result in a highly variable and nonuniform distribution.
18.3 EXPOSURE OF OPERATORS 18.3.1 EXPOSURE PROCESS
In greenhouses and mushroom houses the operator (= applicator) will often prepare the spray liquid (mixing and loading). Therefore, the operator will be exposed during both mixing/loading and application. Compared to most outdoor cultivations, both the total amount of pesticides applied and the duration of the mixing/loading are limited. One operation, for example, the preparation of a 600-liter tank, takes 7-10 min. Therefore, the calculation of an exposure rate (mg/h) is generally not appropriate. Liquid formulations are often used; therefore, dermal exposure during pouring of the concentrated liquid from the bottle into the spray tank is considered to be limited to the hands and incidentally by splashes to the other parts of the body, which may result in large variations in exposure. Respiratory exposure may be relevant for loading powders but is considered to be low for liquids. However, the presence ofrelatively highly volatile solvents in formulations [especially emulsiable concentrates (ECs)] may enhance the occurrence of nonvolatile active ingredients in air.
459
During application, three major processes of dermal exposure can be distinguished: (1) deposition of aerosols, (2) impaction of aerosols by back bouncing, especially when pot benches (AI-Jaghbir et aI., 1992; Fenske et aI., 1987) or mushroom scaffolds (Schipper et aI., 1996) are sprayed, and (3) transfer mainly from contaminated hoses. In a study of within-worker variation for high-volume hand-held spray operators (N = 4, three applications) in greenhouses for the cultivation of carnations, Brouwer et al. (2000a) observed large variations in the exposure of the hands [coefficient of variation (CV) ranging from 62 to 119%] and moderate within-worker variation of exposure of the body (CV from 11 to 55%). Because between-worker variation of exposure differed significantly, and no significant differences were observed between workers for exposure of the hands, it was concluded that exposure of the body will be determined by (1) deposition of generated aerosols and individual working habits and (2) transfer from the crop (determined by crop height and density). Exposure of the hands will be determined by (1) transfer of the spray liquid from equipment and foliage, (2) leakage and splashes, and (3) individual working habits. Respiratory exposure will be highly determined by (1) the size distribution of the generated aerosols, (2) the vapor pressure of the active ingredient, and (3) the direction of the movement of the operator related to the movement of the aerosols. For dermal exposure, Methner and Fenske (1994) clearly demonstrated the effect of both the operators' experience to avoid spray mists and the effect of forced air movement on the degree of the resulting exposure. 18.3.2 OPERATOR EXPOSURE DATA 18.3.2.1 Mixing and Loading
Very few data on mixing and loading in greenhouse pesticide operations have been published and are mostly limited to exposure of the hands. Fenske et al. (1987) reported a mean exposure rate of 35.6 I-lg/h for actual exposure of the hands to fosetylAl (below protective gloves). These authors showed that this was only about 6% of the total actual exposure. In this case mixers loaded 12 times, handling a total of 649 g [547 g of active substance (a.s.)] of the formulated material (wettable powder). De Vreede et al. (1996) reported actual exposure (no protective gloves were worn) to methomyl (EC formulation) ranging from 0.11 to 174 mg during preparation of a spray solution (N = 19; arithmetic mean (AM) = 15.9; SD = 42.5 while handling 152 ± 100 g a.s.). Brouwer et al. (2000a) observed a range of actual exposure (no protective gloves were worn) from 0.09 to 5.8 mg (N = 9) during mixing and loading of 130 ±11O g a.s. propoxur (EC formulation). 18.3.2.2 Application
Compared to outdoor applications, data on pesticide operator exposure in greenhouses and mushroom houses published in the literature are relatively scarce. Most data are from exposure
Table 18.2 Summary of Exposure Studies for Inhalation and Dermal Exposure in Greenhousesa No. of Application Technique
Crop
High-volume spraying
Vegetables
Active substance Pirimiphos-methyI
Exposure assessment methodology Respiratory exposure
Hand exposure
Body exposure
Reference
3
Impinger
Gloves
Pads
Adamis et al., 1985
applicators
dimethoate; permethrin Ornamentals
Fosethyl-Al
4
Filter
Hand rinse
Pads
Fenske et at., 1987
Ornamentals
Fluvalinate; chlorpyrifos; ethazol;
5
Foam plugs
Hand wash (not specified)
Pads
Nigg et al., 1988, 1993
6
FilterIXAD4
Gloveslhand wash
Pads
Rech et al., 1988
Impinger
Pads
Pads
AI-Jaghbir et al., 1992
FilterIXAD2
Gloves
n.d.
Brouwer et al., 1992a
dicofol; chlorothalonil Ornamentals
Abamectin
Vegetables
Dimethoate
Ornamentals
Chlorothalonil;
6
13
thiophanate-methyl ..a;. ~
0
Ornamentals
Abamectin; dodemorph
n.d.
Gloves
n.d .
Brouwer et al., 1992b
Ornamentals
Chlorothalonil
19 4
Filter
Gloves
Coverall
Groenewegen, 1992
Ornamentals
Mevinphos
4
XAD4
Ethanol rinse
Pads
Kangas et al., 1993
Ornamentals
Methomyl
19
XAD2
Gloves
Coverall
Brouwer et at., 1994b
Empty benches
Tracer
9
n.d.
VITAE
VITAE
Methner and Fenske, 1994
Mushrooms
Benomyl; carbendazim
3
Filter
Gloves
n.d.
Schipper et aL, 1996
Ornamentals
Propoxur; methiocarb
5
XAD2
Gloves
Coverall
De Vreede et al., 1996
Stationary sampling
Gloves
Pads
Mestres et al., 1985
3
Foam plugs
Hand rinse
Pads
Nigg et al., 1988
4
XAD4
Ethanol rinse
Pads
Kangas et at., 1993
10
Filter
Gloves
n.d.
Brouwer et al., 1992a
Filter
Hand wipe
Cotton underwear
Giles et al., 1994
Low-volume spraying!
Vegetables
Deltamethrin
fogging
Ornamentals
Fluvalinate; chlorpyrifos
Stamper et al., 1988
ethazol Ornamentals
Mevinphos
Dusting
Ornamentals
Chlorothalonil; zineb; maneb
Air-assisted reduced volume
Omamentals
Permethrin
electrostatic spraying an.d., not determined; VITAE, video imaging technique to assess dermal exposure.
18.3 Exposure of Operators
461
Table 18.3 Summary of Exposure Data for High-Volume Applicationa Potential exposure Hands
Actual exposure Body
(~g/h)
Hands
Body
Range
AM
AM
Range
AM
AM
Reference
(mg/h)
(mg/h)
(mg/kg a.s.)
(mg/h)
(mg/h)
(mglkg a.s.)
Range
AM
Range
AM
Adamis et aI., 1985
n.d.
n.d.
n.d.
0.57-3.36
n.g.
n.g.
n.d.
n.d.
n.d.
n.d.
(incl. hands) Fenske et aI., 1987
n.d.
n.d.
n.d.
n.g.
15
27
2-72
23.6
Nigg et aI., 1988, 1993
n.d.
n.d.
n.d.
1.9-35.9
11.7
674
0.04-9
n.g.
n.g.
n.g.
Rech et aI., 1988
n.d.
n.d.
n.d.
n.g.
0.8
175
18-322
85
n.a.
n.a.
250
AI-Jaghbir et aI., 1992
n.d.
n.d.
n.d.
208-252
228
1842
n.d.
n.d.
n.d.
n.d.
Brouwer et aI., 1992a
n.g.
17.8 (GM)
45
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
Groenewegen, 1992
4.1-60.7
28
62
5-30
18
40
n.d.
n.d.
n.d.
n.d.
Brouwer et aI., 1993b
n.g.
0.7 (GM)
23
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
19.2 (GM)
32
Kangas et aI., 1993
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
2.5-511
n.d.
n.d.
n.d.
Methner and Fenske, 1994
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
158-1996
985
De Vreede et aI., 1996;
0.03-15.2
1.9
n.d.
0.2-12.8
3.2
162
n.d.
n.d.
n.g.
330
Schipper et aI., 1996
0.8-7.1
n.g.
n.g.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
Brouwer et aI., 1997b, 2000a
0.01-10.9
1.04
9.9
0.2-15.2
4.3
21
n.d.
n.d.
n.d.
n.d.
0.04-18.9
2.1
10.6
0.3-93.2
13.9
68
Brouwer et aI., 1994b
aIf not given in the study, the arithmetic mean (AM) is calculated from the presented data. n.d., not determined; n.g., not given; GM, geometric mean.
during use of hand-held high-volume spraying equipment in ornamental plants (Adamis et aI., 1985; Al-Jaghbir et aI., 1992; Brouwer et aI., 1992a, c, 1993b, 1994b, 1997b; De Vreede et al., 1996; Fenske et al., 1987; Groenewegen, 1992; Kangas et al., 1993; Methner and Fenske, 1994; Nigg et al., 1988, 1993; Rech et al., 1988). In reports and papers by Mestres et al. (1985), Nigg et al. (1988), Stamper et al. (1988), Groenewegen (1992), Brouwer et al. (1992c), and Giles et al. (1994), data were given on exposure during low-volume application (cold fogging, thermal fogging, dusting, and electrostatic application), whereas only Schipper et al. (1996) provided some data on the application to mushroom crops (Table 18.2).
exposure. Only few data are available for actual exposure of the body. Table 18.4 illustrates the distribution of the dermal exposure of the body from six studies. A consistent, relatively high, exposure of the legs can be observed. Depending on the height of the crop or benches, the upper parts of the body will be exposed more heavily. Available datasets do not enable a comparison between low- and high-volume techniques (Table 18.5).
Dermal Exposure Very few studies contain complete data sets on potential exposure of the hands and the body, mainly because studies in the United States that have been reported are conducted in compliance with label instructions that prescribe the use of protective gloves. Potential exposure of the hands (Table 18.3) shows large ranges (0.01-60.7 mg/h). Based on five studies in the Netherlands an average potential exposure of about 30 mg/kg a.s. can be calculated for high-volume application. Potential exposure of the body shows a larger variation (0.2-252 mg/h). A mean potential body exposure of about 375 mg/kg a.s. could be calculated based on all available studies (N = 8). However, excluding the very high values in the study by Al-Jaghbir et al. (1992) revealed an average of 167 mg/kg a.s. The range of the actual exposure of the hands (0.04511 J.lg/h) is about 1/100 of the range reported for potential
Table 18.4 Distribution of the Potential Dermal Exposure (Excluding Hands) as a Percentage of Total Body Exposure during Hand-Held High-Volume Application
Respiratory Exposure Table 18.6 summarizes the concentrations in the operators' breathing zone. Again large variations
No. of Torso
Head
observations
Reference
Legs
Arms
Fenske et aI., 1987
51
43
5
4
Rech et aI., 1988
75
11
13
6
AI-Jaghbir et aI., 1992
44
40
12
Groenewegen, 1992
72
5
13
6 n.d.
4
Brouwer et al., 1994b
50
17
33
19
Brouwer et aI., 1997b, 2000a
72
11
16
6
Median
62
14
13
6
Arithmetic mean
61
21
15
6
n.d., not determined.
462
CHAPTER 18
Greenhouse
Table 18.5 Summary of Exposure Data During Low·Volume Application Respiratory exposure
Potential exposure (mg/h)
Concentration (flg/m 3)
Body
Hands
Exposure (flg/h)
AM
Range
AM
Range
AM
Range
0.07 (actual)
n.a.
3 (actual)
n.a.
5.2 (stationary sampling)
n.a.
n.a.
n.a.
0.004-1.5 (actual)
0.4 (actual)
0.5-2.2
1.1
21-96
52
17
42
Brouwer et aI., 1992a
n.g.
39.4 (GM)
n.d.
n.d.
n.g.
670 (GM)
n.g.
n.g.
Groenewegen, 1992
13-34
24
8-100
54
330-410
370
410-540
475
Giles et aI., 1994
n.a.
0.4
n.a.
188
n.a.
n.a.
n.a.
Reference
Range
Mestres et aI., 1985 Nigg et aI., 1988;
AM
Stamper et aI., 1988
n.d., not determined; n.g., not given; n.a., not applicable (1 data point); AM, arithmetic mean; GM, geometric mean.
Table 18.6 Respiratory Exposure Data during High. Volume Application, Adjusted to an Inhalation Rate of 1.25 m 3/h Exposure (flg!h)
Concentration (flg/m 3)
Reference
AM
Range
Range
Adamis et aI., 1985
0.02
n.d.
0.007-1.8
Fenske et aI., 1987
27.8
7.7-50.6
6.5-42
Rech et aI., 1988
0.3
0.13-0.63
0.2-0.5
AI·]aghbir et aI., 1992
3053
2748-3268
n.d.
Brouwer et aI., 1992a
n.d.
n.d.
40 (GM)
Groenewegen, 1992
135
9-315
7.4-252
Kangas et aI., 1993
n.d.
n.d.
0.04-11.4
Brouwer et aI., 1994b
11.6
n.d.
0.36-34.4
Brouwer et aI., 2000a
18.3
0.8-36.8
0.6-29.4
n.d., not determined; AM, arithmetic mean; GM, geometric mean.
are shown. Mean respiratory exposure ranged from 0.02 to 135 Ilgih, excepting the mean reported by Al-Jaghbir et al. (1992). Climatological conditions, vapor pressure of the active ingredient, application rate, and droplet size distribution are likely responsible for the large within- and between-studies variation. Reported concentrations during the application by lowvolume techniques were somewhat higher than those reported during high-volume application. The highest concentrations were observed during dusting (Brouwer et aI., 1992a).
18.4 REENTRY WORKER EXPOSURE Reentry exposure is determined by the pesticide concentration available in the surroundings of the worker. This includes the amount still present in air as vapor and/or aerosols or available by evaporation from crop, soil, objects, or materials present or the amount brought into the air by the workers' activities (e.g., resuspension of deposited residues) and the degree of contact of the worker with contaminated surfaces (mainly the crop itself).
In the latter case, exposure is determined by transferability of the contamination from the crop or materials to the clothing or skin of the worker. 18.4.1 EXPOSURE DATA
The literature contains relatively few reports on the levels of exposure in greenhouses during crop activities and these are limited almost exclusively to harvesting. The reason for this lack of information may be that the agricultural size is relatively small and the use of organophosphorous compounds is less important than in outdoor citrus culture, for instance. The major crops cultivated in greenhouses are ornamental flowers and edible commodities. The pesticide treatment for the latter group is regulated by preharvest intervals and residue limits to protect the consumer. This is not the case for ornamental crops. In addition, the export of flowers may require a virtually complete absence of crop diseases, which may lead to relatively high frequencies and rates of application. The majority of relevant European data have been collected in Germany, Denmark, Finland, and The Netherlands, with special emphasis on the level of exposure, its relationship with dislodgeable foliar residue, and the dissipation of pesticides on foliage. Similar data have been collected in the United States. The results of these exposure studies are summarized in Table 18.7. 18.4.1.1 Ambient Air Concentrations
The exposure to fumigants, such as methyl bromide, is not considered here, because it requires a complete compound-specific analysis. The use oflow-volume misters and fogging equipment will give rise to relatively high levels of liquid aerosols and vapor in the greenhouse over quite long time periods (Brouwer et aI., 1992d; Kangas et aI., 1993; Liesivuori et aI., 1988; Lindquist et aI., 1987; Manninen et aI., 1996; Williams, 1978; Williams et aI., 1980). In most of these studies workers were not monitored. The concentration levels in air were monitored by stationary techniques. The most relevant exposure data are summarized in Table 18.8. The results are quite variable. Ventilation usually leads to a quick drop of the ambient concentrations.
Table 18.7 Summary of Exposure Studies in Greenhouses Relevant for Worker Reentrya No. of Reference
workers
Not given
Sulfotep; nicotine
None
Impinger train
n.d.
n.d.
No crop
Sulfotep
None
Gas analysis
n.d.
n.d.
Chrysanthemums
Methyl parathion
5 workers
Stationary: XAD4
Hand wash
n.d.
Crop
Williams, 1978
Smoke fumigation
Williams et aI., 1980
Fumigation cans
Maddy et al., 1981
Not given
Bud removal Lobel and Schunk, 1982
Hand spreading
Exposure assessment methodology
Active substance
Application technique
Chrysanthemums
Respiratory exposure
Hand exposure
Body exposure
twice Aldicarb
None
Stationary: active coal
n.d.
n.d.
Dicofol
None
Stationary:
n.d.
n.d.
n.d.
n.d.
No data
No data
Gerbera Mestres et al., 1985
Pneumatic disc harrow
Lemon trees
(0.55 Vmin) Mestres et aI., 1985
Vaporizer
ftorisil French beans
Deltamethrin
None
(1.5 Vmin)
.a;;..
=" ~
Waldron,1985,
High-volume and
describing unpublished
low-volume
ftorisil Various
Various
Not given
Stationary: amberlite; charcoal;
reports Wagner and Hermes, 1987
Stationary:
PUF and glass fibers
Spraying and spreading
Not given
Dichlorvos; metharnidophos
Not given
FPP filter or glass filter
Hand wash or pads
aldicarb Liesivuori et al., 1988
Fumigation (no details)
Roses
Pirimicarb; nicotine; suIfur
Pads on inner and outer clothing
None
Stationary: impinger or
n.d.
n.d.
Hand rinse with water
n.d.
Long-sleeved cotton gloves
n.d.
n.d.
Long-sleeved cotton gloves
n.d.
Stationary: MCE filter (T)
n.d.
n.d.
Long-sleeved cotton gloves
n.d.
XAD2 Boleij et aI., 1991 Brouwer et aI., 1992a, b
Dusting and spraying
Tomato
(no details)
Cucumber
High-volume
Carnations
High-volume
Roses
low-volume misting Brouwer et aI., 1992d
Low-volume mister
Not given
PAS-6 with glass fiber filter
Chlorothalonil; thiophanate-
94
methyl; zineb; thiram
knapsack dust blower Brouwer et aI., 1992c
Methomyl
Abamectin; dodemorph;
MCE filter/IOM head XAD2IIOM head
75
bupirimate Freesias
Thiophanate-methyl;
None
dichlorvos
glass fiber filter and impinger (D)
Peelen et aI., 1992
High-volume
Chrysanthemum:
Mancozeb
cutling
Chlorothalonil
12
MCE filterlIOM head
harvesting (continues)
Table 18.7 (continued) No. of
Exposure assessment methodology
Reference
Application technique
Crop
Active substance
workers
Respiratory exposure
Hand exposure
Body exposure
Kangas et al., 1993;
Mist blower
Various
Mevinphos
17
Stationary and PAS:
Hand wash with ethanol
Pads
J auhiainen et al., 1992
Manual sprayer
Ornamentals
Fogger
Undescribed activities
High-volume spray gun
Carnations
n.d.
n.d.
Wipe and hand wash under
Undergarments under
Brouwer et aI., 1993b
XAD4 Propoxur
16
knapsack dust blower Giles et aI., 1994
High-volume and
Double glass fiIterlIOM head
Roses
Permethrin
Not given
Stationary: glass fiber
leather gloves
normal work clothing
Quartz filter and XAD2
n.d.
n.d.
reduced volume Lenhart and Kawamoto, 1994
High-volume and
Various
Diazinon; chlorpyrifos
None
cold fog
.J;o.
0'.
.J;o.
Veerman et al., 1994
High-volume
Chrysanthemums
Chlorothalonil
36
MCE filterlIOM head
Long-sleeved cotton gloves
n.d.
Hoekstra et al., 1996
Surveillance study
Not given:
Benomyl and degradation
Not given
Stationary and PAS:
Cotton gloves
Patches
nurseries
products
Manninen et al., 1996
Cold fog and
Roses and
Deltamethrin; diehlorvos
10
Stationary and PAS:
high-volume
mixed flowers
High-volume and
Cucumbers
Nilsson et al., 1996
glass fiber filters
Vinclozolin; triadimefon
None
Not given
Stationary: Teflon filter
Pads outside and inside clothing
n.d.
n.d.
Cotton gloves:
Cotton overalls or
and impinger
cold fog Nilsson and Papantoni, 1996
No data
XAD2
Cucumbers
Vinclozolin
3
None
one or two per hand
cotton sleeves (forearms)
Schipper et aI., 1996
High-volume sprinkler
Mushrooms
Benomyl
6
n.d.
Cotton gloves
Carnations
Methiocarb; propoxur
49
n.d.
Long-sleeved cotton gloves
n.d.
and drencher Brouwer et aI., 1997a
High-volume
Cotton whole-body overalls
Kirknel et al., 1997
Hydraulic spray boom
Various
Pirimicarb; paclobutrazol;
Hand rifle
ornamentals
endosulfan; methomyl;
Cold fog Martinez Vidal et aI., 1997
High-volume
SKCXAD2
Cotton gloves
(two sections)
Cotton whole-body overalls
mercaptodimethur Peppers
(l'-
and fJ-endosulfan;
lindane a n.d., not determined.
36
None
Stationary: PUF
n.d.
n.d.
18.4 Reentry Worker Exposure
465
Table 18.8 Ambient Air Concentrations after Application of Pesticides in Greenhouse
Reference
Application technique and rate (glha)
Ambient concentration (J.!g/m 3 )
WilIiams, 1978 (sulfotep) WilIiams, 1978 (nicotine)
11 g/280 m3 (smoke fumigation) 14 g/280 m3 (smoke fumigation)
6 x 10 + 5 after 3 ha; 500 after 3.5 h
WilIiams et aI., 1980 (sulfotep)
Fumigation cans; 22 g/450 m 3
3-50 ppb after initial ventilation
Maddy et aI., 1981 (methyl parathion)
2250; technique not given
3.2-8.9 after 24 h; 0.25-4.8 after 2 days
Lobel and Schunk, 1982
Not given; hand spreading
<20-60 after 1 and 2 days; <20 after 4 days
Mestres et aI., 1985 (dicofol)
960 (pneumatic disc harrow; low-volume)
2.8 after 1-2 days; 0.55 after 7 days; <0.1 after 9 days
3 x 10 + 6 after 3 ha; 200 after 24 h
Mestres et aI., 1985 (deItamethrin)
25 (vaporizer; low-volume)
0.008 after 30 min; <0.001 after 2 days
Waldron, 1985 (dichlorvos)
Not given (fog)
From 200-400 down to 1-3 in 10 h; then constant for about 10 h
Waldron, 1985 (permethrin)
Not given (high-volume and low-volume)
Unclear picture
Waldron, 1985
Not given (various techniques)
Aldicarb (granules): no airborne residues; captan and carbaryl:
(aldicarb, carbaryl, and captan)
undetectable after 4 h
Wagner and Hermes, 1987
Not given
<6 after 3-4 days (aldicarb and methamidophos)
Liesivuori et aI., 1988 (nicotine)
Fumigation (no details)
28 after I h; <0.5 after 8 h
Liesivuori et aI., 1988 (pirimicarb)
Fumigation (no details)
1.2 after 0.5 h;
Liesivuori et aI., 1988 (sulfur)
Fumigation (no details)
0.7 after 4 h
Boleij et aI., 1991 (methomyl)
700-5600 (dusting) and
LOD 4.7 at day 1; up to 1.7 at day 2; around 0.5 at following days
3800-3900 (spraying); no details Brouwer et aI., 1992d (thiophanate-methyl)
1000 (iow-volume mister)
1000-2000 after 1 h; 100-200 after 3 h
Brouwer et aI., 1992d (dichlorvos)
2000-4000 after 1 h; 800-1500 after 3 h; 600-1000 after 5 h
Kangas et aI., 1993 (mevinphos)
825 (Iow-volume mister) 0.3-5.7 J.!g/hab (various techniques)
31 after 10 h (fogger); stilI detectable after 40 h (all techniques)
Lenhart and Kawamoto, 1994 (diazinon)
Not given (cold fog)
70-250 after 1 day; 27--67 after 2 days; 20-59 after 3 days;
Hoekstra et aI., 1996 (benomyl)
Not given
Below LOD (0.2 J.!g/sample)
Manninen et aI., 1996 (dichlorvos)
33 (mg/m 3 ) cold fog
80 after 2 h; 35 after 6 h; 10-12 after 18-21 h; 4-7 after 24-42 h 754 after 2 h; 518 after 6 h; 110-140 after 18-24 h; 7 after 42 h
Manninen et aI., 1996 (deItamethrin)
50 (mg/m3) cold fog 0.3 to 2.9 (mg/m 3 ) cold fog and
Nilsson et aI., 1996 (triadimefon)
0.016-0.024
<3 after 0-2 h 4-74 after 8.5-11 h; 1.3-37 after 15.5-18.5 h
19-40 after 4 days
5.7-218 after 2 h;
high-volume Nilsson et aI., 1996 (vinclozolin)
1.0-2.08 (cold fog)
Nilsson et aI., 1996 (vincIozolin)
1.0-1.82 (high volume)
3.7-25 after 13.5-16.5 h
Kirknel et aI., 1997 (pirimicarb)
400-523 (boom)
25-50 after 7 h; 7 after 1 day; 2 after 2 days
742 (boom)
45 after 2 h; 6 after 1 day; 1 after 2 days
2230 (rifle)
23 after 3 h; 6 after 1 day; < 1 after 2 days
745 (boom); also 298 (methomyl)
24 after 3 h; 2.6 after 1 day; < 1 after 2 days; methomyl: 7 after 3 h
1630 (boom)
4 after 1 day
882 (rifle)
1.2 after 1 day
Kirknel et aI., 1997 (pacIobutrazol)
10-16 (boom); 14 (boom)
Kirknel et aI., 1997 (endosulfan)
710 (cold fog); 710 (cold fog)
44 after I day; 60 after 2-3 days
Kirknel et aI., 1997 (methomyl)
750 (cold fog); 834 (cold fog)
25 after 6 h; 10 after 1 day
Kirknel et aI., 1997 (mercaptodimethur)
814 (cold fog); 3200 (rifle);
360 after 2 h; 6 after 6-7 h;
Martfnez Vidal et aI., 1997 (lindane)
0.4 kg/ha (high-volume)
3300 at application; 287 after 24 h
Martfnez Vidal et aI., 1997 (endosulfan)
0.6 kg/ha (high-volume)
4300 at application; 323 after 24 h
LOD, limit of detection; LOQ, limit of quantitation. aThese values probably contain a printing error; a factor 1000 lower is likely. bThese values probably contain a printing error; a factor 1000 higher is likely.
466
CHAPTER 18
Greenhouse
In some cases, the ambient concentrations remain quite high without ventilation of the greenhouse. This is especially true for dichlorvos, a relatively volatile compound (Brouwer et aI., 1992d; Manninen et aI., 1996). The conclusion is that reentry intervals of 8 h, with windows completely open during the last 2 h to ventilate the greenhouse, are sufficient for all pesticides and low-volume application techniques to ensure safe inhalation levels. For volatile compounds (vapor pressure above 10-100 mPa at 20°C), windows should be open to some extent during the next few days to prevent a buildup of possibly toxic concentrations. When highvolume techniques are used for application, a reentry interval of 8 h (for nonvolatile pesticides) or 2 h with the windows wide open is considered sufficient. These guidelines are based mainly on the work of Brouwer et al. (1992d) and have become part of Dutch regulations. During harvesting the worker may be exposed to the volatilized pesticide vapor, although this exposure is expected to be very low, when reentry intervals as described above are used. Another source of pesticide may be the reintranement (resuspension in air) of (dust) particles, especially when the crop is treated with a dustable powder. Brouwer et al. (1990, 1993b) have shown that during harvesting of carnations the inhalation exposure may be detectable but relatively low and is probably related to the dislodgeable foliar residue. As long ago as 1980, Williams et al. observed that the concentration of sulfotep in air increased due to worker activities in the crop. 18.4.1.2 Inhalation Exposure In Table 18.9 the results of worker exposure studies in greenhouses are summarized. As can be seen from the table the inhalation exposure varies considerably. Likely reasons are the variations in sizes of greenhouses and the variations in application rates. These variables are not always presented by the authors. Overall, it should be stated that not only is the number of studies small, but also the presentation of the results is generally incomplete with respect to exposure-influencing variables and there are frequently reasons to believe that printing errors are made in the formats of the exposure. Because the inhalation exposure will be related to the application rate, it seems appropriate to "normalize" the level of exposure by dividing it by the application rate. This leads to the format (J.Lg/h)/(kg/ha). Van Golstein Brouwers et al. (1996) have followed this approach for analyzing available data from Dutch greenhouses. They considered data related to harvesting of flowers (cutting and sortingibundling). The average value for the exposure levels during cutting was about 0.1 (mg/h)/(kg/ha), with a maximum about 1 order of magnitude higher. For sortingibundling, the average exposure was about 1 order of magnitude lower. 18.4.1.3 Dermal Exposure Investigations by Morse et al. (1979) showed that one should not underestimate the possible risks due to the presence of pesticides relatively long after treatment, because on chrysanthemums and carnations imported into the United States from
South and Central America, appreciable amounts of various pesticides could be found. These findings were confirmed by Saiz et al. (1992) who studied exposure to pesticides by handwashes of florists in Sacramento County in California. Lobel and Schunk showed in 1982 that 10-20 days after granule application of aldicarb to the soil of growing chrysanthemums, the picking of cuttings caused an appreciable cholinesterase inhibition in workers, which was completely absent in a control group using rubber gloves. Although the dislodgeable foliar residue was not measured; (metabolites of) aldicarb, possibly stemming from crop residues and or juices, apparently got to the hands during picking. Wagner and Hermes (1987), lauhiainen et al. (1992), and Lander et al. (1992) also showed appreciable cholinesterase inhibition in workers due to crop-related activities. In Table 18.9 the dermal exposure data have been summarized for activities (mainly harvesting) with greenhouse crops. The data are largely fragmentary, with possible exceptions for the harvesting of ornamental crops. Van Golstein Brouwers et al. (1996) summarized the exposure data after "standardization" by application rates. The average exposure level is close to 10 (mg/h)/(kg/ha) for the cutting of the flowers and about 5 (mg/h)/(kg/ha) for sorting and bundling of flowers. It should be noted that exposures were only measured for the normally bare hands and forearms. Possible exposure through clothing or on other uncovered body parts was not considered. However, the exposure data were observed in high crops and may give an overestimation for low crops. Dermal Exposure ModeIing As indicated before, there appears to be a direct relationship between exposure on one hand and application rate and degree of contact between crop and worker on the other hand. Therefore, many authors (Krieger et aI., 1992; Popendorf and Leffingwell, 1982; Ross et aI., 1993, 1994; Van Hemmen, 1993b; Van Hemmen et aI., 1995; Zweig et aI., 1985) described exposure in a standardized way (Brouwer et aI., 1997b, 2000a, b). The basic assumption of the developed models is that exposure results from the transfer of pesticide residue present on the crop during worker activities. More specifically, it was stated that dermal exposure [DE (mg/day)] is determined by the amount of transferable residue [expressed as (one-sided) dislodgeable foliar residue (DFR) (mg/m2), a crop- and taskspecific transfer factor (TF) (m 2/h) and duration of reentry (T), (h/day)], which can be expressed in the most general form as
DEi = DFRi,t x TFm x Tm
(18.1)
where i is the ith pesticide, m is the mth task, and t is the tth day after application. The TF is an empirical factor that is assumed to be cropand task-specific and relatively pesticide-independent. TFs have been derived from exposure data and data on dislodgeable foliar residue established for a variety of crop activities and are accepted as relevant for risk assessment. However, variances between TFs may be substantial (Krieger et aI., 1990, 1992).
Table 18.9 Summary of Exposure Data for Reentry Workers in Greenhouses a Application rate
Inhalation exposure
Dermal transfer factor
Reference
(glha)
(~glh)b
(cm 21h)
Maddy et aI., 1981
2,250
n.d.
(methyl-parathion) Wagner and Hermes, 1987
Potential exposure (mglh)
< 6 after 3-4 days
(~glh)
Body
Hands
Body
Range: 0.009-0.053 after
n.d.
9-53 after 5-6 days;
n.d.
5-6 days; AM: 0.02 Not given
Actual exposure
Hands
Included in body data
(aldicarb and
AM: 20 Highest 0.74
Included in body data
(methamidophos)
Lowest 0.008 (methamidophos)
methamidophos) Boleij et aI., 1991
700-5,600
(methomyl)
0.9-18
Up to 0.362;
(GM per experiment)
GM: 0.01-0.32
n.d.
rc (SD)
GM (GSD)
cutting: 2,900 (500)
cutting: 14.4 (2.3);
Brouwer et aI., 1992a, b
Spraying:
(chlorothalonil)
2,610
Brouwer et al., 1992a, b
Dusting:
GM (GSD)
rc (SD)
GM (GSD)
(chlorothalonil)
3,030
cutting:
cutting: 6,500 (600)
cutting: 4.4 (3.3);
Brouwer et aI., 1992a, b
Spraying:
n.d.
(thiophanate-methyl)
3,750
Brouwer et al., 1992a, b
Dusting:
(zineb)
n.d.
Up to 362;
n.d.
GM: 10-322 n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
GM (GSD)
n.d.
sortinglhundling: 27 (1.3)
112.5 (6.8) ~
=" -...I
sortinglhundling: 3.5 (5.7) rc (SD)
GM (GSD)
cutting: 3,700 (1,400)
cutting: 16.1 (2.0);
GM (GSD)
rc (SD)
GM (GSD)
1,000
dusting/cutting:
cutting: 2,800 (300)
cutting: 8.7 (5.0);
Brouwer et aI., 1992a, b
Dusting:
GM (GSD)
rc (SD)
GM (GSD)
(thiram)
740
dusting/cutting:
cutting: 10,000 (1,100)
cutting: 10.4 (2.5);
Brouwer et al., 1992c
7-15
sortingfbundling: 11.5 (2.3)
125 (3.4)
sortingfbundling: 3.9 (7.7)
50 (2.5) n.d.
(abamectin)
Brouwer et aI., 1992c (dodemorph)
1,130-3,000
n.d.
sortingfbundling: 4.6 (1.9) rc (SD)
GM (GSD)
cutting: 1,200 (850);
cutting: 0.013 (2.0);
cutting: 13 (2.0); sortingfbundling:
sorting: 2,400 (500);
sorting bundling:
bundling: 2,250 (1,200)
0.018 (3.8)
rc (SD)
GM (GSD)
cutting: 4,550 (1,000);
cutting: 1.8 (1.7);
cutting: 1,8000.7);
sorting: 2,400 (2,650);
sortingfbundling: 1.9 (2.2)
sortingfbundling:
bundling: 6,250 (2,350)
18 (3.8) n.d.
GM (GSD)
n.d.
1,900 (2.2) (continues)
Table 18.9 (continued)
Application rate (g/ha)
Inhalation exposure (J.lg/h)b
Brouwer et a!., 1992c (bupirimate)
520-540
n.d.
Peelen et al., 1992
Surveillance study
Reference
(chlorothalonil) Pee1en et al., 1992
Surveillance study
(mancozeb) Kangas et aI., 1993;
0.3-5.7 J.lglha
Dennal transfer factor (cm 2 /h)
Potential exposure (mg/h) Hands
Body
Hands
re (SD)
GM (GSD)
n.d.
GM (GSD)
cutting: 2,400 (600)
cutting: 2.2 (2.0)
1-20;
Range: 1-5;
AM: 5
AM: 3
12-119
Range: 4-20
AM: 35
AM: 9
Fog application:
Jauhiainen et al., 1992
4.5-12 after 9-12 h
(mevinphos)
manual spraying: 2.5
1.1
AM: 133
Actual exposure (J.lglh) Body n.d.
cutting: 2.2 (2.0) n.d.
Range: 1-5;
n.d.
AM: 3 n.d.
Range: 4,000-20,000;
n.d.
AM: 9,000 0.03-15.2
1.9
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
<12
< 100 (upper body)
± 0.6 (Ist day), ± 0.5 (2nd day)
Brouwer et al., 1993b
810 (spray gun)
GM (GSD)
(propoxur)
250 (dust blower)
cutting: 15 (2.5), range 2.4-39;
""-
=" QC
bundling: 3.8 (2.1), range 0.9-7.5 Giles et aI., 1994
450
(low-volume;
<0.15 after 1 h
157 (lower body)
pennethrin) Giles et aI., 1994
900
(high-volume;
<0.15
n.d.
n.d.
<12
after 1 h
< 100 (upper body) < 100 (lower body)
pennethrin) Veennan et al., 1994
Surveillance study;
GM (GSD)
GM (GSD)
GM (GSD)
(harvesting)
not given
30 (1.6) after
pulling: 1,165 (2.2);
manual: 7.2 (lA) after
(chlorothalonil)
n.d.
GM (GSD)
11-17 days,
bundling: 792 (2.1);
11-17 days,
11-17 days,
4 (2.9) after
other: 2,008 (2.6)
0.9 (1.6) after
900 (1.6) after
32-56 days;
32-56 days;
automatic: 3.6 (1.7) after
automatic: 3,600 (1.7) after
32-56 days
n.d.
manual: 7,200 (lA) after
11-17 days,
11-17 days,
0.59 (4.9) after
590 (4.9) after
32-56 days
32-56 days (continues)
Table 18.9 (continued)
Potential exposure (mg/h)
Application rate (g/ha)
Inhalation exposure (J.!g/h)b
(cm2 /h)
Hands
Body
Hands
Body
Veerman et al., 1994
Surveillance study;
GM (GSD)
GM (GSD)
GM (GSD)
n.d.
GM (GSD)
n.d.
(wrapping)
not given
10 (2.1) after
holding: 1,174 (2.4);
wrapping: 3.3 (2.1) after
11-17 days,
wrapping: 496 (4.0)
(chlorothalonil)
Hoekstra et aI., 1996 (benomy1)
Manninen et al., 1996
Not given
Not given
(dichlorvos)
Dermal transfer factor
Actual exposure (J.!g/h)
Reference
wrapping: 3,300 (2.1) after
11-17 days,
11-17 days,
3 (2.4) after
0.35 (5.5) after
350 (5.5) after
32-56 days
32-56 days
Below LOD
Range: left: ND 118.4; right: 8.7-75; AM: left:
12.5-175 after 1 day,
32-56 days Range: thigh: 0.06-
Range: left: ND 118.4;
0.22 (ll-g/cm 2 Jh)
right: 8.7-75; AM: left:
34.9; right: 26 (formats
AM: 0.14
34.9; right: 26 (formats
not given)
(J.!g/cm2/h)
not given)
No data
Range: 0.66-0.69
n.d.
next day; AM: 0.3
62.5 after 40 h
~
=" \C
Not given
No data
<0.125
(no data given) no data
AM: 0.145
(deltamethrin) Nilsson and Papantoni,
Range: 0.004--0.409;
l.4c
AM: 2.25 (calculated)
n.d.
AM: 0.50
19-33% of potential exposure
after 42 h Manninen et al., 1996
n.d.
Range: 0.1-19.7; AM: 7.1
n.d.
n.d.
(calculated)
1996 (vinclozolin) 2,500-2,600
n.d.
Brouwer et al., 1997b (methiocarb)
Spraying:
n.d.
Brouwer et at., 1997b
Spraying:
0.5--40;
(propoxur)
AM: 10.2 n.d.
Schipper et a/., 1996
<0.06-0.23
n.d.
<60-230
n.d.
AM: 1,500
Range: 1.6-10.7; AM (SD): 2.5 (2.7)
Range: 0.08-12.1; AM (SD): 3.9 (4.7)
Range: 1,600-10,700; AM (SD): 2,500 (2.7)
Range: 20-2,300; AM: 310
AM: 2,000
Range: 0.03-1.07;
Range: 0.1-7.3;
Range: 30--1070;
Range: 20-190;
AM (SD): 0.44 (0.58)
AM (SD): 1.77 (2.5)
AM (SD): 440 (0.58)
AM: 60
Range: 0.007--0.062;
n.d.
Range: 7-62;
n.d.
(benomyl) 148
Kirknel et al., 1997
379 123; pot roses,
(pirimicarb)
cuttings
Kirknel et al., 1997
742: pot roses,
(pirimicarb)
trimming of cuttings
GM: 18
GM: 0.018 n.d.
DFR below LOQ
Range: 0.0535 (after
n.d.
2-3 days); 0.0075--0.013
Range: 7.5-13; GM: 9
n.d.
53.5
(after 9-10 days); GM: 0.009 Kirknel et al., 1997
2230: kalanchoe,
(pirimicarb)
cuttings
2.5-3
No hand exposure
0.0085--0.009
n.d.
(continues)
Table 18.9 (continued)
Actual exposure (I-lg/h)
Application rate (g/ha)
Inhalation exposure (I-lglh)b
Dennal transfer factor (cm2 /h)
Hands
Body
Hands
Body
Kirlmel et al., 1997
1630: hedera helic pots,
540
2.1
0.35
210
n.d.
(pirimicarb)
placing, leaf removal
Kirlmel et al., 1997
882: cut roses
AM: 3,391; GM: 4,553
0.022,
0.414,
22, 117
n.d.
(cutting)
0.117
0.415
0.004-0.029
At about the LOQ
4-29
n.d.
LOQ to 0.007
LOQto7
n.d.
2,638-2,990
90th: 0.245 (cutters);
90th: 0.61 (cutters);
90th: 245 (cutters);
n.d.
(complex: 2 isomers,
90th: 0.658 (packers)
90th: 0.23 (packers)
90th: 658 (packers)
31,
n.d.
n.d.
Reference
n.d.
(pirimicarb)
Potential exposure (mg/h)
AM: 4,838; GM: 4,553
(bud removal) Kirlmel et al., 1997
10-16:
(pacIobutrazol)
pot mini roses,
(manual); 300 (mechanical); 505 (tagging)
packing and tagging
"'....:a"
Kirlmel et aI., 1997
14: pot roses,
(paclobutrazol)
bud removal,
Kirknel et al., 1997
71 0: begonia, cuttings,
(endosulfan)
packing pot plants
Kirlmel et al., 1997
750: various pot plants,
(methomyl)
packing pots
Kirknel et at., 1997 (methomyl)
packing and leaf removal
Q
packing: AM: 1,000
1,295 (bud removal)
cuttings
peculiar DFR)
834: various pot plants,
Kirknel et aI., 1997
3,200: large pot plants,
(mercaptodimethur)
packing
Kirlmel et at., 1997
1,000: hedera pot plants,
(mercaptodimethur)
packing and cutting
o(no hand exposure)
144
8-10
<0.001
0.018-0.036
<1.0
n.d.
<40
1,168, 1,555
3.3,3.9
0.31,
3,290, 3,890
n.d.
5,850,9,950
n.d.
2.4 38,115
4,150, 7,070
5.8,10.0
1.1,1.5
(surrounded by plants)
an.d., not detennined; rc, regression coefficient; AM, arithmetic mean; GM, geometric mean; GSD, geometric standard deviation; LOD, limit of detection; LOQ, limit of quentitation. b Assuming a breathing rate of 1.25 m 3 /h (for fomlat change). CJn Nilsson and Papantoni (1996) the application rate is about 1000 times higher for the same compound.
18.4 Reentry Worker Exposure
The actual dislodgeable foliar residue at the time of reentry is considered to be an important source of strength for dermal exposure. The level of actual pesticide residue depends on the initial amount ofDFR, the dissipation rate, and the elapsed time since application. The initial foliar residue is determined by the amount applied to the crop, the distribution over the crop, that is, the leaf area index, and a factor that accounts for the nonhomogeneity as a result of uneven spraying. This can be depicted as DFRo
= AR x
I jLAI
(18.2)
where AR is application rate (kg/ha), LAI is leaf area index (m 2/m 2), and I is a factor for uneven spraying, i.e., interception of the spray by the crop (dimensionless). The pseudo leaf area index (LAIj l) can be calculated from formula (18.2) when application rate and DFRo are known. Typical values assumed for I are 0.2-1, and typical values for LAI are 1-15. For worker exposure estimations, the DFR must be determined in the contact zone between worker and crop. The dissipation of the foliar pesticide deposit is considered to be a complex process of environmental factors, metabolism, translocation due to foliar penetration and plant growth, and pesticide formulation (Bentson, 1990). Willis and McDowell (1987) assumed that the process of dissipation may in many cases be described by a first-order process. Thus, the relationship between initial and actual DFR can be described as lnDFRt
= lnDFRo + kt
(18.3)
where DFRo is the initial DFR (day 0 after application) and k=-ln2(t1/2)-1
(18.4)
where tl/2 is half-life (days). Dissipation curves of pesticides have been used as a risk assessment tool to establish reentry intervals; that is, the calculated expectation that pesticide residues would be reduced to the safe reentry exposure levels within this interval (Dong et aI., 1992). Pesticide Dissipation and Transfer Factors The rates of foliar residue dissipation have been reported for many pesticides and many crops; however, very limited data on greenhouse crops have been published. As a part of several field studies on reentry exposure, dissipation curves for various pesticides have been established. In Finnish greenhouses, mevinphos in particular was studied (Jauhiainen et aI., 1992; Kangas et aI., 1993) as used for chrysanthemums and roses. The dissipation on foliage correlated well with the decrease in the observed dermal exposure of workers over the same time period (2 days). The half-life on foliage was estimated to be about 9 h and the transfer factor based on a figure from Kangas et al. (1993) was about 133 cm2/h (Table 18.9). This value is extremely low compared to all other available values for similar tasks. On request, the authors indicated (J. Kangas, personal communication) that the main reason for this low value lies in the use of protective gloves, whereas actual exposure was measured on the hand
471
underneath gloves. Liesivuori et al. (1988) estimated the halflives on roses for some pesticides (Table 18.10). In German greenhouses the half-lives of several pesticides on various commodities (Table 18.10) were determined by Goedicke (1987, 1988a, 1988b, 1988c, 1989) and Goedicke et al. (1989). Although initial dislodgeable foliar residues were presented, the information on the application rates was insufficient to estimate a reasonable LA!. From the data a LAI of 130 for tomato and 16 for apple foliage could be estimated. For methamidophos on gherkins, roses, and gerberas, a transfer factor of about 700 cm 2/h was estimated by Goedicke (1989). In Dutch greenhouses the investigations of Brouwer et al. (1992a, b, c, 1993b) and Van Hemmen (1992) aimed to determine transfer factor(s) during the harvesting of roses and carnations for various pesticides to see whether the data had a strong "generic power." They concluded that the harvesting of roses led to somewhat lower transfer factors (with average values for three pesticides during cutting, sorting, and bundling of 1,200-6,250 cm 2/h) than for the harvesting of carnations (with average values for four pesticides during cutting of 2,800-10,000 cm 2/h) (Table 18.9). A recent reevaluation of all available data (published and unpublished), not considering variations due to crop-pesticide peculiarities for the harvesting of sprayed carnations amounted to values of about 1,400-2,200 cm2/h as the most appropriate data (Brouwer et aI., 1997a, 1997b; Emmen et aI., 1996). For the picking of chrysanthemum cuttings the transfer factor was about 400 cm2/h for two pesticides (S. Peelen, D. H. Brouwer and J. J. Van Hemmen, unpublished observations). Veerman et al. (1994) measured the transfer factors for the harvesting of chrysanthemums. The data are included in Table 18.9. The values varied from about 500 to 2,000 cm 2/h for the tasks involved. In this work only the transfer to hands and forearms of the workers was estimated. The application rate was a major determining factor for the level of dislodgeable residue. On the basis of a comparison between the dislodgeable foliar residues and the application rates for the various rose/pesticide combinations in 18 greenhouses an estimate for the pseudo leaf area index was about 10-40. For the various carnation/pesticide combinations in 18 other greenhouses this index was about 5-10. In a study Kirknel et al. (1997) presented a large survey of Danish greenhouses. In 8 different greenhouses the reentry activities for 12 different species were considered using 5 different pesticides. Twenty-one transfer factors were calculated for a range of work activities. These transfer factors had a log normal distribution. The geometric mean was about 1,500 cm2/h, with a 90th percentile of about 5,200 cm 2/h for N = 16, leaving out some unrealistic low values « 10 cm 2/h), which could not be explained. The transfer factors were estimated for hand exposure, since the bare hands were the main area of exposure. In some experiments the level of potential total body exposure was similar to the actual exposure of the hands. In the majority of the cases actual body exposure rarely exceeded the observed hand exposure. In the Dutch studies, the dissipation of several pesticides was observed on ornamental crops. These data are presented in
472
CHAPTER 18
Greenhouse
Table 18.10 Estimated Initial Half-Lives Observed for Various Pesticides on Various Crops in Greenhouses Pesticide
Crop
Half-lives (days)
Abamectin
Chrysanthemum
1.1-2.4
6
Netherlands
Benomyl
Rose
1.8
7
Finland
Reference/country
Benomyl
Lettuce
14
8
Germany
Bitertanol
Chrysanthemum
2.5
6
Netherlands
Carbendazim
Gherkin
7.2
2
Germany
Carbendazim
Gherkin
7.9
3
Germany
Carbendazim
Lettuce
10-12
8
Poland
Carbendazim
Tomato
4--6
8
Poland
Carbendazim
Rose
2.2
7
Finland
Chlorothalonil
Carnation
>14
4
Netherlands
Chlorothalonil
Chrysanthemum
3-5
6
Netherlands
12
France
Deltamethrin
French bean
>10
Diazinon
Gherkin
6
8
USA
Diazinon
Tomato
7
8
USA
Dichlorvos
Chrysanthemum
6
Netherlands
Dicofol
Gherkin
0.2 1,2-5
8
Germany
Dicofol
Tomato
14
Dicofol
Lemon
12
8
Italy France
Dimethoate
Tomato
>10 3.25
Dimethoate
Gherkin
1.6
3
Germany
Dimethoate
Gherkin
2.7
8
Germany
Dimethoate
Rose
0.9
7
Finland
Ethephon
Tomato
7-10
8
Germany
3
Germany
Fenazox
Gherkin
1.4--2.4
8
Germany
Fenazox
Tomato
1.3
3
Germany Germany
Fenazox
Tomato
2.3
3
Iprodione
Tomato
10.2
8
Italy
Lindane
Cucumber
2.5
2
Germany
Lindane
Gherkin
2-3
8
Germany
Malathion
Gherkin
4.35
3
Germany
Mancozeb
Chrysanthemum
3
6
Netherlands
Metalaxyl
Tomato
8-10
8
Germany
Methamidophos
Tomato
4.3
2
Germany
Methamidophos
Gerbera
1.4-3.9
3
Germany
Methamidophos Methamidophos
Rose Gherkin
5.7-6.2 2.4
1 Germany 3 Germany
Methiocarb
Carnation
8-15
6
Methomyl
Chrysanthemum
4.7
6
Netherlands
Methomyl
Cucumber
2.5-4
9
California, USA
Methomyl
Tomato
2.1-3.7
13
Methomyl
Cucumber
1.7-2.3
13
Mevinphos
Rose/chrysanthemum
0.4
5
Netherlands
Netherlands Netherlands Finland
Permethrin
Pot chrysanthemum
10-26
Pirimiphos-methyl
Tomato
0.7
Propoxur
Carnation
0.5-0.8
6
Netherlands
Thiophanate-methyl
Carnation
>14
4
Netherlands
10 3
California, USA Germany
Thiophanate-methyl
Carnation
17-41
6
Netherlands
Vinclozolin
Tomato
9.8
8
Italy
Vinclozolin
Cucumber
7
11
Sweden
a 1: Goedicke, 1989; 2: Goedicke, 1988a; 3: Goedicke et ai., 1989; 4: Brouwer et aI., 1992b; 5: Kangas et ai., 1993; 6: Brouwer et ai., 1997b; 7: Liesivuori et ai., 1988; 8: Goedicke, 1988c, see references therein; 9: Edmiston et ai., 1991; 10: Giles et aI., 1992; 11: Nilsson et aI., 1996; 12: Mestres et ai., 1985; 13: Boleij et ai., 1991. Additional fragmentary data obtained in Californian greenhouses for several pesticides on various crops can be obtained from O'ConneII et ai. (1987) and Rech and Edmiston (1988).
18.5 Risk Management Table 18.10, together with compiled data from other sources on greenhouse crops. The present greenhouse data may not give a complete overview of all the available data, but are indicative of the possible variations, not only between crops, but also within crops. Information on pesticide dissipation for nongreenhouse crops can be obtained from Popendorf (1992) and Monte and Van Hemmen (1997). The reader should realize that the dissipation may depend on many factors, which were not completely described. The half-lives presented in Table 18.10 must be considered approximate values, because a first-order dissipation is assumed in several cases. An example of the influence of the season on the half-life of dissipation is given by Ross et al. (1994). These authors showed that methomyl dissipates more than two times faster in the autumn than in early summer. Boleij et al. (1991) observed a half-life of methomy Ion tomato foliage of2.l-3.7 days and on cucumber foliage of 1.7-2.3 days. They observed a good correlation between dislodgeable foliar residue and hand exposure, but did not present transfer factors. 18.4.1.4 Exposure Data Bases and Use for Authorization Procedures The inhalation and dermal exposures observed for workers in greenhouses are summarized in Table 18.9. There are two approaches for estimating surrogate exposure values. Both approaches have been mentioned in earlier paragraphs. One approach is to use formulae (18.1)-(18.4) and the other is normalization for application rate. At present, the first approach can only be used for dermal exposures, because available data for inhalation exposure are not sufficiently detailed for that purpose. The present data base for exposures reflecting the various greenhouse and mushroom reentry activities is too small to give a precise evaluation of the data, but several approaches are being developed (Krebs et al., 1996). Therefore, the following approaches can only be indicative and need further field studies for validation: 1. On the basis of the theoretical approach given in formulae (18.1)-(18.4), it is possible to estimate a worst-case level of dislodgeable foliar residue, using the application rate and a low value for the leaf area index. At 1 kg/ha and a leaf area index of 1, the initial dislodgeable residue estimate is 10 I-lg/cm 2 . Assuming no dissipation of the pesticide in the period between application and reentry activities, this will be the worst-case dislodgeable foliar residue. The second step in this approach is the choice of a relevant transfer factor. On the basis of the data presented in Table 18.9, the value of 10,000 cm2/h (one-sided surface area) is a worst-case choice. For activities not involving bare hands and forearms or not involving high crops, this value may be a gross overestimation. At present it is not possible to estimate a value for inhalation exposure using this approach, although there is some evidence (Brouwer et al., 1992b, 2000b) that a similar approach should be feasible when more data are available.
473
Overall, for harvesting a transfer factor of 2000-3000 cm 2/h seems reasonable as a central tendency value. However, these values are largely derived from exposure data that only reflect hand and forearms. Brouwer et al. (1997b) measured whole-body exposure during harvesting of high carnations. The potential exposure appeared to be about a factor of 5 higher than the exposure to hands and forearms alone. The actual exposure on the rest of the body (measured on undergarments) was so low, that hand and forearm (generally bare during the harvesting) exposure accounted for virtually the whole exposure. In addition, Kirknel et al. (1997) indicated a low contribution of actual body exposure to the actual hand exposure. The value of about 2000-3000 cm2 /h can be compared with estimations of Krieger et al. (1990, 1992) as presented in Van Hemmen et al. (1995). Krieger et al. observed that the transfer factor was strongly related to the degree of worker surface area contacting crop during the activities. A value of 5000 cm2/h for risk assessment of activities in a high crop seems reasonable. 2. The other approach uses experimental data obtained during harvesting of ornamental crops. This is based on high crops and assumes no significant exposure on skin other than hands and forearms. A reasonable value for dermal exposure (Van Golstein Brouwers et al., 1996) that might be used is 20 (mg/h)/(kg/ha). In this approach no further assumptions have to be made on transfer factors and leaf area index, because they are incorporated in the data. With this approach a reasonable worst case exposure for inhalation is 0.2 (mg/h)/(kg/ha).
18.5 RISK MANAGEMENT 18.5.1 RISK MANAGEMENT STRATEGY Reducing risk, by either hazard or exposure, needs a systematic approach or strategy in which four levels can be distinguished with descending preference of action (Table 18.11). Table 18.11 Risk Management Strategya Replacement
Source
(lower toxicity) (lower application rate) (other formulation)
Equipment/ methods/process
Optimalization application Automatization (application, harvesting bundling, internal transportation)
Management
Duration and rotation of task Frequency of application Reentry interval
Protection a For
explanation, see text.
Personal protective measures
474
CHAPTER 18
Greenhouse
The first level is reduction or elimination of the source. This can achieved by substitution of the pesticide by a less hazardous one. Substitution by a pesticide with a lower application rate may reduce the source, or substitution by another formulation form, for example, a wettable powder instead of a suspension concentrate or a granular form, may also result in a reduction of the source. The second level is reduction of exposure by replacement or modification of the process or equipment. Moreover, a critical review of work methods may result in improvement with respect to exposure reduction. The exposure data available (Tables 18.3 and 18.5), however, do not clearly support this method. Examples are other application techniques, for example, more mechanized low-volume spraying techniques that do not require the presence of the applicator during application and automation of the transport of potting plants. Brouwer et at. (1994b) reported a significant reduction with a more mechanized system of clod-breaking during harvesting chrysanthemums. The third level is related to the organization of work and work practice. Reducing the frequency of application will lower total exposure of the applicator as well as the total amount of dislodgeable foliar residue and thus reentry worker exposure. Because there is dissipation of the dislodgeable foliar residue with time, the reentry interval; that is, the time between application and reentry, also can be considered a management tool, although it will not be useful when crops are harvested on a daily basis or even twice a day. The reentry interval can be calculated by comparison of a safe level of the dislodgeable foliar residue as determined by toxicological parameters with the dissipation curve and the initial foliar residue, using formulae (18.1)-(18.4). This approach has been described in detail by Ross and Dong (1996, 1997), based on absorbed daily dose, incorporating probalistic modeling for the various involved parameters. The (fourth) level with the lowest preference is personal protection. Unfortunately, it is quite often the only possibility. However, personal protective equipment has to be fitted to the hazard. Therefore, the nature and level of exposure should be known to select appropriate and comfortable equipment. Brouwer and Van Hemmen (1994) presented an approach for selection of the proper personal protective equipment for both pesticide operators in greenhouses and reentry workers.
18.5.2 PENETRATION THROUGH CLOTHING (MATERIALS)
Several studies describe the protective capacity of normal clothing and of specially designed protective clothing. In view of the large differences in work practice and climatic conditions and the large variation in clothing, it is difficult to estimate default values that can be used for the degree of protection in a general case. The largest body of work is aimed at the mixer/loader and the applicator. For reentry situations it is evident that the
distribution of the contamination will vary largely with the type of crop that is harvested and the type of work. Zweig et at. (1985) showed that harvesters of strawberries, growing on ground level, were exposed mainly on the hands, whereas weeders in the same crops, who created dusty surroundings, were mainly exposed on the torso. For dermal exposure the distribution of the potential dermal contamination is important to determine the parts of the body that have to be protected. Cotton work clothing will generally reduce potential exposure. De Vreede et at. (1996) reported penetration of methomyl through work clothing ranging from a few percent up to 30% for applicators. Brouwer et at. (1997b) observed similar penetration data for propoxur for applicators (N = 58 locations, AM = 11.4%, median = 9.5%) and reentry workers (N = 60 locations, AM = 8.9% and median = 6.6%). Figure 18.1 illustrates the relationship between the external loading and the percentage of penetration. When protective gloves are used, the effectiveness of field practice in greenhouses is reported to be high. Brouwer et at. (2000a) reported efficiencies of 95 and 87% for operators (nitrile gloves) and reentry workers (cotton gloves), respectively. These observations fit well with other studies. Nigg et al. (1986) and Chester et at. (1990) reported 84 and 87% reduction of actual hand exposure of mixer/loader/applicators. For disposable coveralls similar reduction rates were reported. However, overall efficiency of protection based on urinary excretion data in field studies are lower, ranging from 25% up to 50% (Brouwer et at., 2000a; Davies et at., 1982; Nigg and Stamper, 1983). In Table 18.12, default values for protection are presented that may be used for registration purposes (Brouwer et at., 1993a). The values in Table 18.12 are based mainly on data for mixer/loader/applicators from several national authorities, but are not necessarily representative. Major drawbacks are the small amount of underlying data, the nature of the contamination (solid or liquid), and the increase in penetration of wet textile clothing or gloves by sweating or by crop juices.
ACKNOWLEDGMENTS The authors thank Dr. Chris Maas and Ad Vijlbrief of the Netherlands Ministry of Social Affairs and Employment for their stimulative support throughout the years and the Ministry for the financial support to the various field studies that have been carried out in Dutch greenhouses. The authors thank their colleagues who have contributed to these field studies, especially Lambert Leenheers who was in charge of the analytical chemical part of the field studies. For the present manuscript the authors thank Hans Marquart, and Andrea Schipper who have given valuable comments and contributed extensively. The authors also thank the various colleagues in the area of exposure assessment from academia, governments, and industry, who through their criticism have contributed to a great extent to better quality studies and a more elaborate use of these studies in risk assessment. Furthermore, the various farmers and their personnel are thanked for their participation in the various studies. Without their participation these studies and concomitant evaluations could not have been carried out.
18.5 Risk Management
• 35
I, ;
•
•
•
I
25
i
•• •
15
10
I
I
• • • •
•
•••
•
5
• • •
! •
i i
•
••
...
•
•
•
•
o
•
•
100
300
•
•
400
•
• •
500
•
• 600
•
•• 700
• I
•
• 800
900
palontIaIexposure ("*"ograms propoxur)
Figure 18.1 Relationship between the penetration (%) of propoxur through a cotton coverall and the potential exposure of different locations during reentry activities (N = 53).
Table 18.12 Reduction Coefficients for Exposure to Pesticides Using Protective Clothinga Reduction coefficient
Protective clothing Germanyb Gloves
EnglandC
0.01
During rnII of solid formulations
0.01
During rnII of SCs
0.05
During rnII of ECs
0.1
During spraying
0.1
Califomiad
EUROPOEMe
0.1
0.1
Normal clothing Variable
(shirt with long sleeves and long trousers)
0.1
mixingfloadinglapplication
0.25
harvesting Coverall/overall with boots
0.05
Coverall type 3
0
0.1
0.1
0.05
Rainsuit Broad-rimmed hat
0.5
Face shield
0.05
Half mask with filter
0.8
arnll, mixing and loading; SC, suspension concentrate (aqueous solutions); EC, emulsion concentrate (formulations based on organic solvents); variable (5-20%) depending on the volume of the surface contamination. bLundehn et aI., 1992. CJMP, 1986. dBrodberg and Sanbom, 1992; Thongsinthusak et aI., 1990. eEUROPOEM, 1997.
1000
475
476
CHAPTER 18
Greenhouse
REFERENCES Adamis, Z" Antal, A., Fiizesi, 1., Molmrr, J., Nagy, L., and Susan, M. (1985). Occupational exposure to organophosphorus insecticides and synthetic pyrethroids. Int. Arch. Occup. Environ. Health. 56, 299-305. AI-Jaghbir, M. T., Salhab, A. S., and Hamarsheh, F. A. (1992). Dermal and inhalation exposure to dimethoate. Arch. Environ. Contam. Toxico!. 22, 358-361. American Conference of Governmental Industrial Hygienists (ACGIH). (1985). "Particle Size Selective Sampling in the Workplace." ACGIH, Cincinnat. Bentson, K P. (1990). Fate of xenobiotics in foliar pesticides deposits. Rev. Environ. Contam. Toxico!. 114, 125-161. Boleij, J. S. M., Kromhout, H., F1euren, M., Tieleman, W., and Verstappen, G. (1991). Re-entry after methomyl application in greenhouses. App!. Occup. Environ. Hyg. 6, 672-676. Brodberg, R K, and Sanborn J. R. (1992). "Compilation of Clothing Penetration Values: Harvesters." California Dept. of Pesticide Regulation, HS-1652, Cal-EPA, Sacramento. Brouwer, D. H., Brouwer, R, and Van Hemmen, J. J. (1990). Respiratory exposure to field strength dusts in greenhouses during application and re-entry. In "Book of Abstracts Seventh International Congress of Pesticide Chemistry, Hamburg" (H. Frehse, E. Kessler-Schmitz, and S. Conway, eds.), Vo!. Ill, p. 375. IUPAC, Germany. Brouwer, D. H., Brouwer, R, De Mik, G., Maas, C. L., and Van Hemmen, J. J. (l992a). Pesticides in the cultivation of carnations in greenhouses: Part IExposure and concomitant health risk. Am. Ind. Hyg. Assoc. J. 53,575-581. Brouwer, D. H., De Haan, M., Leenheers, L. H., De Vreede, J. A F., and Van Hemmen, J. J. (l997a). Half-lives of pesticides on greenhouse crops. Bull. Environ. Contam. Toxieo!. 58, 976-984. Brouwer, D. H., De Haan, M., and Van Hemmen, J. J. (2000b). Modelling re-entry exposure estimates. Techniques and application rates. In "Worker Exposure to Agrochemicals." Brouwer, D. H., De Vreede, J. A. F., De Haan, M., Van de Vijver, L., Veerman, M. c., Stevenson, H., and Van Hemmen, J. J. (l994b). Exposure to pesticides during and after application in the cultivation of chrysanthemums in greenhouses. Health risk and risk management. Med. Fae. Landbouww. Univ. Gent, Belgium 59(3b), 1393-1401. Brouwer, D. H., De Vreede, J. A F., Meuling, W. J. A, and Van Hemmen, J. J. (2000a). Determination of the efficiency for exposure reduction of protective clothing by biological monitoring in a field study. In "Advanced Methods to Determine Pesticide Worker and Residential Exposure," ACS Symposium Series. Am. Chem. Soc., Washington, DC. Brouwer, D. H., De Vreede, J. A F., Ravensberg, J. C., Engel, R, and Van Hemmen, J. J. (l992d). Dissipation of aerosols from greenhouse air after application of pesticides using a low-volume technique. Implications for safe re-entry. Chemosphere 24, 1157-1169. Brouwer, D. H., et a!. (l997b). Unpublished observations. Brouwer, D. H., Ravensberg, J. c., De Kort, W. L. A. M., and Van Hemmen, J. J. (I 994a). A personal sampler for inhalable mixed-phase aerosols. Modification to an existing sampler and validation test with three pesticides. Chemosphere 28, 1135-1146. Brouwer, D. H., Van Golstein Brouwers, Y., De Haan, M., Peelen, S. J. M., Stevenson, H., De Vreede, J. A. F., and Van Hemmen, J. J. (I 993a). "De Effectiviteit van Huidbeschermingsmiddelen bij het Werken met Bestrijdingsmiddelen in de Glastuinbouw." ("The Effectiveness of Protective Clothing during Exposure to Pesticides in Greenhouse Crops.") TNO Medical Biological Laboratory, MBL 1993-53, Rijswijk, The Netherlands. Brouwer, D. H., and Van Hemmen, J.J. (1994). Fitting personal protective equipment (PPE) to the hazard. Selection of PPE for various pesticide exposure scenarios in greenhouses. In "Book of Abstracts of the American Industrial Hygiene Conference & Exposition," Anaheim, CA, American Industrial Hygiene Association. Brouwer, R., Brouwer, D. H., Tijssen, S. C. H. A, and Van Hemmen, J. J. (l992b). Pesticides in the cultivation of carnations in greenhouses: Part IIRelationship between foliar residues and exposures. Am. Ind. Hyg. Assoc. 1. 53,582-587.
Brouwer, R., Marquart, J., De Mik, G., and Van Hemmen, J. J. (I 992c). Risk assessment of dermal exposure of greenhouse workers to pesticides after re-entry. Arch. Environ. Contam. Toxieol. 23, 273-280. Brouwer, R., Van Maarleveld, K, Ravensberg, J. c., Meuling, W. J. A, De Kort, W. L. A. M., and Van Hemmen, J. J. (1 993b). Skin contamination, airborne concentrations and urinary metabolite excretion of propoxur during harvesting of flowers in greenhouses. Am. J. Ind. Med. 24, 593-603. Chester, G. (1993). Evaluation of worker exposure to, and absorption of, pesticides during occupational use and crop re-entry. Ann. Oeeup. Hyg. 37, 509-523. Chester, G., Loftus, N. J., Woollen, B. H., and Anema, B. P. (1990). The effectiveness of protective clothing in reducing dermal exposure to, and absorption of, the herbicide f1uazifop-P-butyl by mixer-loader-applicators using tractor sprayers. In Book ofAbstracts, Seventh International Congress of Pesticide Chemistry (H. Freshe, and E. Kesseler-Smith), Vo!. Ill. Conway, Hamburg, Germany. Davies, J. E., Freed, V. H., Enos, H. F., Duncan, RC., Barquet, A, Morgade, C., Peters, L. J., Danauskas, J. X. (1982). Reduction of pesticide exposure with protective clothing for applicators and mixers. J. Oeeup. Med. 24,464-468. De Vreede, J. A. F., De Haan, M., Brouwer, D. H., Van Hemmen, J. J., and De Kort, W. L. A M. (1996). "Exposure to Pesticides. Part Ill. Application to Chrysanthemums in Greenhouses." Report S-131-4, Ministry of Social Affairs and Employment, The Hague, The Netherlands. Dong, M. H., Krieger, R. 1., and Ross, J. H. (1992). Calculated re-entry interval for table grape harvesters working in California vineyards treated with methomy!. Bull. Environ. Contam. Toxieol. 49, 708-714. Edmiston, S., Brodberg, R, and Quan, v. (1991). "Dissipation of Methomyl Residues on the Foliage of Greenhouse-Grown Cucumber." HS-1621, CalEPA, Sacramento. Emmen, H. H., Hoogendijk, E. M. G., Brouwer, D. H., Muijser, H., Van Hemmen, J. J., and Kulig, B. M. (1996). "Cumulative Effect of Pesticide Exposure on Human Nervous System Functioning." TNO Report V96.320, Zeist, The Netherlands. EUROPOEM (1997). "The Development, Maintenance and Dissemination of a European Predictive Operator Exposure Model (EUROPOEM) Database." Project AIR3 CT93-1370, BIBRA International, Carshalton, UK Fenske, R A., Hamburg, S. J., and Guyton, C. L. (1987). Occupational exposure to fosetyl-AI fungicide during spraying of ornamentals in greenhouses. Arch. Environ. Contam. Toxico!. 16,615-621. Giles, D. K., Blewett, T. c., Saiz, S. G., Welsh, A. M., and Krieger, R 1. (1992). Foliar and nontarget deposits from conventional and reduced-volume pesticide application in greenhouses. J. Agric. Food Chem. 40, 2510-2516. Giles, D. K, Blewett, T. c., Saiz, S. G., Welsh, A. M., and Krieger, R. 1. (1994). Greenhouse applicator and harvester exposure to pesticides. Acta Hortieult. 372,151-158. Goedicke, H.-J. (1987). Riickstande von Pflanzenschutzmitteln auf Pflanzenoberflachen als Quelle fUr Intoxikationen und Moglichkeiten der Expositionsnormierung. Z. Ges. Hyg. 33, 339-342. Goedicke, H.-J. (l988a). Zum Riickstandsverhalten von Pflanzenschutzmitteln an Pflanzenoberflachen. Z. Ges. Hyg. 34, 279-282. Goedicke, H.-J. (l988b). Zum Riickstandsverhalten von pirimiphos-methyl auf Tomaten, Blatt- und Bodenoberflachen im Gewachshaus. Nahrung 5, 475480. Goedicke, H.-J. (1988c). Riickstandstoxikologische Bewertung des Einsatzes von Pflanzenschutzmitteln und Mitteln zur Steuerung biologischer Prozesse in Gemiise unter Glas und Plasten. Nahrung 6, 565-570. Goedicke, H.-J. (1989). Exposition durch Riickstande auf Blattoberflachen nach Anwendung von phosphororganischen Insektiziden im intensiven Apfelanbau. Z. Ges. Hyg. 35, 533-535. Goedicke, H.-J., Hermes, H., and Wagner, R (1989). Exposition durch Riickstande auf Pflanzenoberflachen nach Anwendung von Pflanzenschutzmitteln im Gewachshaus. Z. Ges. Hyg. 35, 531-533. Groenewegen, 1. (1992). "Orienterend onderzoek naar de dermale en inhalatoire blootstelling aan chloorthalonil tijdens gewasbehandeling." ("Dermal and Respiratory Exposure During the Application of Chlorthalonil: A Pilot Study.") TNO Medical Biological Laboratory, Rijswijk, The Netherlands.
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Hoekstra, E. J., Kiefer, M., and Tepper, A. (1996). Monitoring of exposure to benomyl in nursery workers. 1. Oeeup. Environ. Med. 38, 775-781. Jauhiainen, A., Kangas, J., Laitinen, S., and Savolainen, K. (1992). Biological monitoring of workers exposed to mevinphos in greenhouses. Bull. Environ. Contam. Toxieol. 49, 37-43. Joint Medical Panel of Scientific Subcommittee on Pesticides (MAFF) and British Agrochemicals Association (JMP). (1986). "Estimation of Exposure and Absorption of Pesticides by Spray Operators." SC 800 I, Ministry of Agriculture, Fisheries and Food, Harpenden, Herts, UK. Kangas, J., Laitinen, S., Jauhiainen, A., and Savolainen, K. (1992). "Exposure of Workers to Mevinphos in Greenhouses." Poster presented at a workshop on Occupational Skin Exposure to Chemical Substances of the First International Scientific Conference on Occupational Hygiene in Brussels. Kangas, J., Laitinen, S., Jauhiainen, A., and Savolainen, K. (1993). Exposure of sprayers and plant handlers to mevinphos in Finnish greenhouses. Am. Ind. Hyg. Assoe. J. 54, 150-157. Kirknel, E., N0hr Rasmussen, A., and Emde, G. (1997). "Bekaempelsesmiddelforskning fra Milj0styrelsen." ("Pesticide Re-Entry Exposure of Workers in Greenhouses.") No. 31. Ministry of Environmental Protection Agency, Denmark. Krebs, B., Maasfeld, w., Schrader, J., and Wolf, R. (1996). "Uniform Principles for Safeguarding the Health of Workers Re-Entering Crop Growing Areas after Application of Plant Protection Products." Ks\950822, IVA document, personal communication. Krieger, R., Blewett, C, Edmiston, C, Fong, H., Gibbons, D., Meinders, D., O'Connell, L., Ross, J., Schneider, E, Spencer, J., and Thongsinthusak, T. (1990). Gauging pesticide exposure of handlers (mixer/loader/applicators) and harvesters in California agriculture. Med. Lav. 81, 474-479. Krieger, R. I., Ross, J. H., and Thongsinthusak, T. (1992). Assessing human exposure to pesticides. Rev. Environ. Contam. Toxieol. 129, I-IS. Lander, E, Pike, E., Hinke, K., Brock, A., and Nielsen, J. B. (1992). Anticholinesterase agents uptake during cultivation of greenhouse flowers. Arch. Environ. Contam. Toxieo!. 22, 162-169. Lenhart, S. W., and Kawamoto, M. M. (1994). Residual air concentrations of pesticides in a commercial greenhouse. App!. Oeeup. Environ. Hyg. 9,9-15. Liesivuori, J., Lihkkonen, S., and Pirhonen, P. (1988). Re-entry intervals after pesticide application in greenhouses. Seand. 1. Environ. Health 14 (Suppl. 1),35-36. Lindquist, R. K., Krueger, H. R., and Powell, C C (1987). Airborne and surface residues of permethrin after high- and low-volume applications in greenhouses. 1. Environ. Sei. Health B22, 15-27. Lindquist, R. K., Powell, C C, and Hall, R. E (1993). Glasshouse treatment. In "Application Technology for Crop Protection" (G. A. Matthews, and E. C Hislop, eds.), pp. 275-290. Cab International, Oxon, UK. Lobel, H., and Schunk, W. (1982). Zur Exposition und GesundheitsgeHihrdung der Werktiitigen in Gewiichshausern durch Ruckstiinde cholinesterasehemmender Pflanzenschutzmittel. Z. Ges. Hyg. 28,697-699. Lundehn, J.-R., Westphal, D., Kieczka, H., Krebs, B., Locher-Bolz, S., Maasfeld, W., and Pick, E.-D. (1992). "Einheitliche Grundsiitze zur Sicherung des Gesundheitsschutzes fUr den Anwender von Pflanzenschutzmitteln, Mitteilungen aus der Biologischen Bundesanstalt fur Landund Forstwirtschaft." Heft 277, Berlin, Germany. Maddy, K. T., Reeves, R., Lowe, J., Richmond, D., and Fredrickson, A. S. (1981). "Exposure of Greenhouse Workers to Encapsulated Methyl Parathion (Penncap-M) Being Used on Chrysanthemums in Santa Clara and San Mateo Counties in 1981." HS-884, Cal-EPA, Sacramento. Manninen, A., Kangas, J., Tuomainen, A., and Tahuonen, R. (1996). Exposure to insecticides in the use of cold fog generators in greenhouses. Toxieol. Environ. Chem. 57, 213-224. Martfnez Vidal, J. L., Egea Gonzales, E J., Glass, C. R., Martfnez Galera, M., and Castro Cano, M. L. (1997). Analysis of lindane, (X- and ti-endosulfan and endosulfan sulfate in greenhouse air by gas chromatography. 1. Chromatogr. A 765, 99-108. Mestres, R., Fran~ois, C, Causse, C, Vian, L., and Winnett, G. (1985). Survey of exposure to pesticides in greenhouses. Bull. Environ. Contam. Toxieol. 35,750-756.
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Methner, M. M., and Fenske, R. A. (1994). Pesticide exposure during greenhouse applications. Part I. Dermal exposure reduction due to directional ventilation and worker training. App!. Oeeup. Environ. Hyg. 9,560-566. Monte, S., and Van Hemmen, J. J. (1997). "Dissipation of Pesticides on and in Crop Foliage." TNO Report V97. 132, Zeist, The Netherlands. Morse, D. L., Baker, E. L., and Landrigan, P. J. (1979). Cut flowers: A potential pesticide hazard. Am. 1. Public Health 69, 53-56. Nigg, H. N., and Stamper, J. H. (1983). Exposure of spray applicated and miyerloaders to chlorobenzilate miticide in Floride citrus grows. Arch. Environ. Toxieol. 12,397-482. Nigg, H. N., Stamper, J. H., Easter, E., and DeJonge, J. O. (1993). Protection afforded greenhouse pesticide applicators by coveralls: A field test. Arch. Environ. Contam. Toxieo!. 25,529-533. Nigg, H. N., Stamper, J. H., and Machon, W. D. (1988). "Pesticide Exposure to Florida Greenhouse Applicators." Report 660/2-88/033, U.S. Environmental Protection Agency, Cincinnati. Nigg, H. N., Stamper, J. H., and Queen, R. M. (1986). Dicofol exposure to Florida citrus applications: effects of protective clothing. Arch. Environ. Contam. Toxieo!. 15, 121-134. Nilsson, U. (1995). "Chemical Health Risks after Pesticide Spraying in Greenhouses." Report 199, Swedisch University of Agricultural Sciences, Alnarp, Sweden. Nilsson, U., Nybrant, T., Papantoni, M., and Mathiasson, L. (1996). Long-term studies of fungicide concentrations in greenhouses. 2. Fungicide concentrations in air and on leaves after different exposure times and under different climate conditions. 1. Agrie. Food Chem. 44, 2878-2884. Nilsson, U., and Papantoni, M. (1996). Long-term studies of fungicide concentrations in greenhouses. 3. Exposure risks after spraying in greenhouses. 1. Agrie. Food Chem. 44, 2885-2888. O'Connell, L., Fong, H. R., Cooper, C, Maykowski, R., and Wroe, M. (1987). "A Study to Establish Degradation Profiles for Six Pesticides (Triforine, Endosulfan, Chlorothalonil, Sulfotep, Dodemorph Acetate, and Daminozide) Used on Ornamental Foliage in San Diego County California during Fall 1986. "Cal-EPA, Sacramento. Organisation for Economic Co-operation and Development (OECD). (1997). "Guidance Document for the Conduct of Studies of Occupational Exposure to Pesticides During Agricultural Application." Environmental Health and Safety Publications, Series on Testing and Assessment, No. 9. Peelen, S., Groenewegen, I., Brouwer, R., and Van Hemmen, J. J. (1992). "Blootstelling aan en Gezondheidsrisico's van Mancozeb en Chloorthalonil tijdens Stekplukken in Chrysantenbedrijven." TNO Report MBL 1992-13, Zeist, The Netherlands. Perez, C, and Soderholm, S. C (1991). Some chemicals requiring special consideration when deciding whether to sample the particle, vapor or both phases of an atmosphere. Appl. Oeeup. Environ. Hyg. 6, 859-864. Popendorf, W. (1992). Re-entry field data and conclusions. Rev. Environ. Contam. Toxieol. 128,71-117. Popendorf, W. J., and Leffingwell, J. T. (1982). Regulating organophosphate residues for worker protection. Residue Rev. 82, 125-201. Rech, C, and Edmiston, S. (1988). "A General Survey of Foliar Pesticide Residues and Air Concentration Levels Following Various Greenhouse Applications." HS-1403, Cal-EPA, Sacramento. Rech, C, Bissel, S., and Vall, M. del. (1988). "Potential Dermal and Respiratory Exposure to Abamectin during Greenhouse Applications." Report HS-1491, CDFA, Sacramento. Ross, J. H., and Dong, M. H. (1996). The use of probabilistic modeling to determine re-entry intervals. Fund. App!. Toxieo!. 30, 254. (Poster 1299 at the 35th Society of Toxicology Meeting, March 1996, Anaheim, CA.) Ross, J. H., and Dong, M. H. (1997). The use of probabilistic modeling to determine re-entry intervals. In Operator Exposure and Agroehemieals, IBC UK Conferences, April 1997, London. Ross, J. H., Thongsinthusak, T., and Dong, M. H. (1993). "Factors Influencing Estimates of Re-entry Worker Exposure to Pesticides." HS-1677, Cal-EPA, Sacramento. Ross, J. H., Thongsinthusak, T., and Dong, M. H. (1994). "Factors Influencing Estimates of Re-entry Worker Exposure to Pesticides." HS-1707, Cal-EPA, Sacramento.
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Saiz, S. G., Blewett,T. c., Haskell, D., and Margetich, S. (1992). "Florist Hand Exposure to Pesticides." HS-1642, Cal-EPA, Sacramento. Schipper, H. J., Brouwer, D. H., and Van Hemmen, J. J. (1996). "Exposure to Pesticides during Application and Harvest Activities. Pilot Study Performed in Mushroom Houses." TNO Report V96. 136, Zeist, The Netherlands. Stamper, J. H., Nigg, H. N., Nielsen, A. P., and Royer, M. D. (1988). Pesticide exposure to greenhouse foggers. Chemosphere 17, 1007-1023. Thongsinthusak, T., Brodberg, R. K., Ross, J. H., Gibbons, D., and Krieger, R. I. (1990). "Reduction of Pesticide Exposure by Using Protective Clothing and Enclosed Cabs." California Dept. of Pesticide Regulation, HS-1616. Van Golstein Brouwers, Y. G. C., Marquart, J., and Van Hemrnen, J. J. (1996). "Assessment of Occupational Exposure to Pesticides in Agriculture. Part IV. Protocol for the Use of Generic Exposure Data." TNO Report V96. 120, Zeist, The Netherlands. Van Hemmen, J. J. (1992). Agricultural pesticide exposure databases for risk assessment. Rev. Environ. Contam. Toxieo!. 126, 1-85. Van Hemmen, J. J. (1993b). Re-entry exposure and product development. Pesticides and greenhouse crops: an example. In "Agrochemical Occupational Risk Assessment. The Future," Brussels, Jelllinek, Schwartz and Connolly, Washington, DC. Van Hemmen, J. J., Van Golstein Brouwers, Y. G. c., and Brouwer, D. H. (1995). Pesticide exposure and re-entry in agriculture. In "Methods of Pesticide Exposure Assessment" (P. B. Curry, S. Iyengar, P. A. Maloney, and M. Maroni, eds.), pp. 9-19. Plenum, New York.
Veerman, M. c., Van de Vijver, L., De Haan, M., Brouwer, D. H., and Van Hemmen, J. J. (1994). "Exposure to Pesticides. Part IV. The Harvesting of Chrysanthemums in Greenhouses." S-131-5, Ministerie van SZW, The Hague, The Netherlands. Wagner, R., and Hermes, H. (1987). Exposition der Giirtner wiihrend und nach der Applikation von Dichlorvos, Methamidophos sowie Aldicarb in Gewlichshausanlagen. Z. Ges. Hyg. 33, 255-257. Waldron, A. C. (1985). The potential for applicator-worker exposure to pesticides in greenhouse operations. In "Dermal Exposure Related to Pesticide Use. Discussion of Risk Assessment" (R. C. Honeycutt, G. Zweig, and N. N. Ragsdale, eds.), ACS Symposium Series 273. Am. Chem. Soc., Washington, DC. Williams, I. H. (1978). Dissipation of sulfotep and nicotine in a greenhouse atmosphere following their use as fumigants. J. Environ. Sci. Health B13, 235-241. Williams, D. T., Denley, H. v., and Lane, D. A. (1980). On site determination of sulfotep air levels in a fumigating greenhouse. Am. Ind. Hyg. Assoe. 1. 41,647-651. Willis, G. H., and McDowell, L. L. (1987). Pesticide persistence on foliage. Rev. Environ. Cantam. Taxieol. 100,23-73. Zweig, G., Leffingwell, J. T., and Popendorf, W. M. (1985). The relationship between dermal exposure by fruit harvesters and dislodgeable foliar residues. J. Environ. Sei. Health B20, 27-59.
CHAPTER
19 Coping With Aggregate Pesticide Exposure Assessment: An Integration Approach Michael H. Dong John H. Ross Worker Health and Safety Branch, CallEPA Department of Pesticide Regulation Using four case studies, this chapter demonstrates the impacts of three basic aggregate exposure assessment approaches on the risk characterization of pesticides. The approaches considered were deterministic accumulation, probabilistic addition, and probabilistic integration. Deterministic accumulation is a conventional approach in which high-end or otherwise conservative point estimates are calculated for various exposure events and these doses are then totaled to yield the aggregate dose in question. Probabilistic addition and probabilistic integration as discussed in this chapter both use the Monte Carlo simulation technique to determine the distribution of each exposure event. Instead of summing the 95th (or any predetermined upper) percentile doses simulated for the various events, the integration approach involves one additional simulation. In this extra simulation, the output probabilistic distributions simulated for the various events are reused to describe the input parameters of the integration exposure model. The four cases include a 6year-old child playing on treated playground structures, an adult swimming in a treated pool, a strawberry harvester reentering a treated field, and a worker mixing, loading, and applying a pesticide. In all four cases, the accumulation approach yielded the highest aggregate daily doses whereas the integration approach yielded the lowest. The swimmer case showed that the accumulation approach could yield an aggregate dose 10 times higher than the dose simulated by the integration approach. The pesticide handler case showed that the aggregate dose determined from available biomonitoring data was 6 times less than the 95th percentile aggregate dose simulated by the integration approach.
19.1 INTRODUCTION Pesticide exposure assessment plays a central role in the risk characterization process. It is considered to be the most comHandbook of Pesticide Toxicology
Volume 1. Principles
plex component compared to others in the process, in that current tools for measuring human exposure to pesticides are quite medium- and route-specific. As pointed out by the National Research Council (NRC, 1983), "In contrast with hazard identification and dose-response assessment, (pesticide) exposure assessment has very few components that could be applicable to all media (or all routes)." To paraphrase the Council's argument, an assessment model for inhalation exposure is necessarily quite different from a model used to describe exposure from dermal contact or dietary intake. Complicating further the task of performing pesticide exposure assessment are the facts that submission of chemical-specific exposure data is not mandatory (see, e.g., USEPA, 1994, 1998a, b, 1999) and that assessors often can or need to make different assumptions about several basic exposure parameters, such as exposure frequency and duration, rates of dermal contact and dietary intake, or dermal and inhalation absorption. Until the recent emergence of probabilistic analysis in exposure assessment, common practice has been to select high-end, or otherwise conservative, point estimates for the exposure scenario in question. This traditional approach often results in an overestimation of the health risk involved, especially when multiple exposure pathways are considered. As an alternative approach to offering a more realistic exposure estimate, the U.S. Environmental Protection Agency (EPA) in 1997 recently released its first set of guiding principles for the use of probabilistic analysis in chemical exposure assessment (USEPA, 1997a). In California, the guidelines for probabilistic exposure assessment were drafted by the CallEPA Office of Environmental Health Hazard Assessment in December 1996 (OEHHA, 1996). Preceding the release of these (draft) guidance documents, the CallEPA Department of Pesticide Regulation (CDPR) had used the probabilistic technique to complete an exposure assessment for diazinon (Dong et aI., 1994a).
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Coping With Aggregate Pesticide Exposure Assessment: An Integration Approach
Although the concept of totaling pesticide exposures from all potential sources is not new, most regulatory agencies did not begin to implement this totaling or aggregation approach until after the U.S. Congress passed the 1996 Food Quality Protection Act (FQPA). Section 408(a)(4) of the FQPA specifies with respect to a tolerance "that there is a reasonable certainty that no harm will result from aggregate exposure to the pesticide chemical residue, including all anticipated dietary exposures and all other exposures for which there is reliable information." Now that consideration of aggregate exposure has become the law rather than an option, regulatory agencies must be more attentive to the mechanism in which exposures from all sources are aggregated. They must realize that simply summing conservative point estimates calculated for several exposure pathways will not give a reliable aggregate exposure estimate except for use as a screening level. This simple accumulation approach may not be appropriate because not all exposure events occur with equal probability. To aggregate exposures from all pathways scientifically and realistically, probabilistic assessment is apparently the only practical alternative. This view is currently shared by many people in the pesticide exposure assessment community, including those who have participated in the recent workshop held by the International Life Sciences Institute (ILSI, 1998) on aggregate exposure assessment. As the movement of aggregate exposure assessment is still in its infancy, much of ILSI's workshop discussion focused on the underlying principles rather than on the processes or methodologies that would lead to the estimation of a comprehensive aggregate exposure. The probabilistic integration approach, to which this chapter directs much of its attention, is one step beyond the current use of the probabilistic addition technique in pesticide exposure assessment. It uses probabilistic analysis to simulate the aggregate dose in question. The four case studies that follow will contrast the aggregate doses estimated from using the probabilistic integration approach with those from using the probabilistic addition and the deterministic accumulation approach. Overall, this chapter is an attempt to demonstrate the impacts of three aggregate exposure assessment approaches on the risk characterization of pesticides.
19.2 APPROACHES FOR AGGREGATE EXPOSURE To date, there are primarily two types of exposure estimation techniques available from which a few basic approaches to totaling human exposures can be made. Point estimation, otherwise known as deterministic analysis, and probabilistic analysis are the two types of techniques currently in use for the estimation of human exposure to pesticides. Point estimation is the conventional method in which conservative or otherwise high-end point estimate values are used for most parameters in an exposure algorithm. These high-end point estimates are typically extremely improbable and yield highly conservative intake or uptake potential that will most likely overestimate
the risk involved. The probabilistic analysis, also known as Monte Carlo simulation or the stochastic analysis, is considered a more realistic alternative wherein probabilistic distributions for the various key exposure factors (e.g., environmental concentrations, exposure duration, body surface, etc.) are used in the algorithm instead of their point estimates. The general distinction between the two exposure estimation techniques, along with a basic description of the Monte Carlo simulation process, has been discussed in the aforementioned federal (USEPA, 1997a) and state (OEHHA, 1996) guidance documents and by numerous investigators, including Thompson et al. (1992), Whitmyre et al. (1992), Copeland et al. (1994), Finley and Paustenbach (1994), and Dong et al. (1994a, b). The three basic approaches to aggregating exposures from all relevant sources are described below, with their overview presented graphically in Fig. 19.1. The four case studies that follow are intended as a point for discussion on the use of the three approaches in aggregate pesticide exposure assessment. Table 19.1 lists the algorithms (later on also referred to as equations) used to calculate the exposures for the various events included in the case studies. It is important to note that, in practice, the use of these algorithms is essentially bounded by the data available. For example, the dermal exposure for strawberry harvesters can be measured directly from patch or whole body dosimetry data, instead of (as indicated in Eq. 12 in Table 19.1) through extrapolation using a dermal transfer factor and foliar dislodgeable residues. In the four case studies, Crystal Ball (1996) was the computer software used to implement the Monte Carlo simulations, although other stochastic software could be used. 19.2.1 THE DETERMINISTIC-BASED ACCUMULATION APPROACH
On a single day, a person may be exposed to the same pesticide outdoors at the treatment site and later indoors at home. The exposure may be from dermal contact, dietary intake, ingestion of drinking water, or inhalation. Some of the source media may involve mUltiple exposure pathways. The daily dose received from each of these potential sources can be calculated using the point estimate values. The daily doses calculated from the various sources or events can then be summed to yield an estimate for the aggregate exposure in question. For lack of a better term and to be consistent with regulatory problems associated with the underlying estimation technique used, this method of totaling the relevant exposures is referred to herein as the deterministic-based or the deterministic accumulation approach. This approach is deterministic in nature because the various event-specific exposures to be added are point estimates with each used typically without a distribution or a statistical error term. Deterministic accumulation as an aggregate exposure approach has been used for years by the CDPR in combining occupational exposure with dietary intake to yield a total acute or chronic dose. Interestingly, the FQPA does not require aggregating exposures for the most highly exposed sUbpopulation (i.e., workers).
19.2 Approaches for Aggregate Exposure
481
DETERMINISTIC ACCUMULATION
I
AIIDay Dose
probabilistic - --------,
PROBABILlSTIC ADDITION
probabilistic PROBABILlSTIC INTEGRATION
Figure 19.1
Three basic approaches to estimating aggregate exposure.
Table 19.1 Basic Exposure Assessment Algorithms per Activity, as Used in the Case Studies
= (daily consumption rate) x (food residue level) x (oral absorption), for all relevant food items
1.
Dietary intake
2.
Inhalation exposure = (air level: indoor ambient, outdoor ambient, or at treatment site) x (inhalation volume, per event) x (inhalation absorption)
3.
Drinking water = (daily consumption rate) x (water concentration) x (oral absorption)
4.
Dermal uptake absorption)
5.
Dermal uptake [treated surface] absorption)
[soil]
(soil
concentration)
x
(skin-soil
loading,
per
event)
x
(soil
residues)
x
(dermal
= (dislodgeables on surface) x (hourly dermal transfer rate for entire body) x (exposure duration) x (dermal
= (soil concentration) x (hourly soil ingestion rate) x (exposure duration) x (oral absorption)
6.
Hand-to-mouth [soil]
7.
Hand-to-mouth [treated surface/foliage] = (dislodgeables on surface/foliage) x (hourly dermal transfer rate for entire body) x (exposure duration) x
8.
Dermal [swimmer]
9.
Inhalation [swimmer] = (air concentration around pool) x (inhalation volume, per event) x (inhalation absorption)
(body portion as hand surface) x (residues portion on hand for oral intake) x (oral absorption)
= (water concentration) x (hourly permeability coefficient) x (surface area of exposed skin) x (exposure duration) = (water concentration) x (hourly ingestion rate) x (exposure duration) x (oral absorption) = (residues on each strawberry) x (number of strawberries picked and eaten per work day) x (oral absorption)
10.
Oral/drinking [swimmer]
11.
Oral intake [in fields]
12.
Dermal [harvesters] = (dislodgeables on foliage) x (hourly dermal transfer rate for entire body) x (exposure duration) x (dermal absorption)
13.
Dermal [handlers]
= (daily dermal as measured by patch dosimetry) x (dermal absorption)
19.2.2 THE PROBABILISTIC-BASED ADDITION APPROACH
As discussed earlier, for each of the route- or medium-specific exposure events considered, the input parameters used for exposure calculation can assume a probabilistic distribution instead of a point value. As a result, a probabilistic distribution would be simulated for each of the route- or medium-specific exposure events involved in the aggregation. An estimate can then be calculated for the aggregate exposure of concern by simply adding the 95th percentiles (or other appropriate upper-bound) of the various simulated output distributions. This method of aggregating exposures from all relevant sources or events is referred to here as the probabilistic-based addition or summation
approach. The 95th percentile, instead of a higher percentile, is referenced here because a value beyond the 95th percentile would be subject to greater uncertainty, as pointed out by Finley and Paustenbach (1994).
19.2.3 THE PROBABILISTIC-BASED INTEGRATION APPROACH
Conceptually, it can be seen that even adding all the simulated 95th percentiles for exposures received from all relevant sources would potentially yield a much higher (or sometimes lower depending on the direction in which the distributions are skewed) 95th percentile aggregate exposure estimate unless all
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event variables were properly linked. At the least, simple addition would assume incorrectly that the various simulated 95th percentiles all bear the same weight (i.e., the simulated distributions all having the same exact shape or skewness). This chapter proposes to go one step further in using the various simulated 95th percentiles to calculate the aggregate exposure estimate. This extra step or process involves one additional final Monte Carlo simulation in which the output probabilistic distributions from previous simulations will be reused to describe the input parameters of the final aggregate exposure model. As depicted in Fig. 19.1, unlike those for the route- or medium-specific exposure events (Table 19.1) involved in the aggregation, the input parameters for simulation in the integration exposure model are additive rather than multiplicative. Because the effort in this extra process is to integrate appropriate weights to the simulated output probabilistic distributions (which are reused as input distributions for the final model), it is referred to here as the probabilistic-based integration approach. One apparent advantage of this integration approach is that it offers a probabilistic distribution, rather than a point estimate, for the aggregate dose in question. As discussed later in the fourth case below, there is another less-attainable version of the integration approach that relies on biomonitoring data, not probabilistic analysis, to account for the doses received from all potential exposure sources.
19.3 CASE STUDIES For proprietary reasons, an effort was made in this chapter not to disclose the identity of the active ingredients used in the four case studies. The various environmental (i.e., ambient air, drinking water, soil, treated structure surface, etc.) concentrations were either exaggerated or reduced by roughly 1.5-2 times of what had been published in the open literature or in registration documents. To work with as real or as credible a case study as possible, the values used for many of the nonchemical specific parameters were either the usual defaults adopted by regulatory agencies including CDPR (Thongsinthusak et aI., 1993) or derived directly from the EPA's Exposure Factors Handbook (USEPA, 1997b). The values used in this chapter for a few nonchemical specific parameters, such as the time that a 6-year-old child would spend at a playground in a given day, were based on professional judgement, as such data were not readily available. There are some empirical data (USEPA, 1996) showing that on average a child would spend somewhere between 20 and 60 minutes outdoors per day, but none specifying the time spent on playground structures. There were also a few parameters whose values and distributions were assumed purely for illustration purposes. For example, there were no data or basis, empirical or otherwise, to support the presumption made in the third case that in a day's work it is as probable for a harvester to eat one strawberry picked off the field as to eat 40 or more. Following each case description below is a summary table listing the exposure-related parameters used, along with their
point values and, where applicable, their probabilistic distributions. As can be seen in these tables, a few parameters (e.g., dermal transfer factor, exposure duration, clothing penetration, etc., as included in the first case) were assigned a fixed value when in fact they could be treated as a random variable. Treating these few parameters as a constant reduced the complexity of the overall exposure assessment model undertaken. More importantly, they were so treated because presently a realistic probabilistic distribution could not be determined or assumed for them. In 1997, a model was proposed for prediction of the percent penetration of pesticides through clothing (Ross et aI., 1997). That model had its focus on handler activities, however, because it was based entirely on the data presented in the Pesticide Handlers Exposure Database (PHED, 1995), the data base that has been described by Hamey et al. in this Handbook in Chapter 22. The children's outdoor activities, such as rolling on the playground, can be different from the activities of those handling pesticides in the field. Another complication of using the clothing penetration model is that its prediction is a function of the outside deposition (i.e., a function of the amount of residues present on the outer clothing layer), which in most cases is yet to be determined. Simply put, this second complication involves a correlation between two variables that cannot be resolved easily. In any case, in this chapter assigning a fixed value to some input parameters did not violate the guiding principles set forth by the EPA (USEPA, 1997a), which assert specifically that "From a computational standpoint, a Monte Carlo analysis can include a mix of point estimates and distributions for the input parameters to the exposure model." 19.3.1 THE 6-YEAR-OLD CHILD
This case study focused primarily on the application of a wood preservative to outdoor playground structures. The active ingredient in this wood preservative is an inorganic metal, and is ubiquitous in the environment. This inorganic metal can be found in drinking water, ambient air, food, and even cigarette smoke (which leads to a difference between the outdoor and indoor ambient air concentrations even in the same residence). In food, much of this active ingredient is in the less toxic organic form and hence was not added to the inorganic portion. Figure 19.2 summarizes the potential sources (i.e., the various pathways and media) through which a 6-year-old child could be exposed to this wood preservative. The parameters and their values used for this case study are listed in Table 19.2. 19.3.2 THE ADULT SWIMMER
In addition to drinking water or other potential sources, people can be exposed to antimicrobial agents applied to swimming pools and spas. To deal with this type of exposure potential, the EPA (Dang, 1996) developed an assessment model specific to swimmer exposure. In their assessment model, known as the Swimmer Exposure Assessment Model, the EPA provided algorithms for calculating the exposures that a swimmer
19.4 Results and Discussion
483
~ .potential but not included in this case study as considered inconsequential due to low vapor [ pressure, low dermal absorption , etc.
Figure 19.2 Multiexposure pathways via multimedia for the 6-year-old child who has spent time at a playground treated with a wood preservative.
can receive from seven potential routes of entry. The seven routes of entry are the greater skin surface proper, inhalation (at the pool), oral/drinking (of pool water), the buccal/sublingual area, the orbital/nasal area, the aural area, and the sexual organ area. In this second case study, only the first three of the seven routes of entry were considered since there are not sufficient data to distinguish the permeability of these other various minor epidermal areas. Also for purposes of this second case study, drinking water and ambient air away from the pool vicinity were considered the two only additional sources from which a swimmer may be exposed to an antimicrobial agent applied to water. Figure 19.3 provides a graphic account ofthe five potential pathways and media through which a swimmer, in this case an adult, can receive exposure to a volatile water antimicrobial agent. Table 19.3 lists the parameters and their values used for this case study. 19.3.3 THE STRAWBERRY HARVESTER For agricultural field workers, the primary exposure pathway is through dermal contact during reentry. In this case study, the potential sources for exposure considered for a harvester reentering a treated strawberry field were dermal contact with treated foliage, inhalation exposure from working in the field, hand-to-mouth exposure at the field, eating treated strawberries picked off the field, exposure to ambient air (away from the work site), and dietary intake during off-work hours. The dose received from drinking water ingestion that was not included in this case study as such was presumed, for simplicity, to be inconsequential. Another justification for the exclusion of this source is that compared to reentry exposure, drinking water is indeed a secondary or tertiary source at best for the pesticide involved. The six potential sources for exposure are
summarized graphically in Fig. 19.4. The parameters and their values used for each of these exposure sources are listed in Table 19.4. 19.3.4 THE AGRICULTURAL HANDLER The fourth case involved the agricultural handler. In addition to the inhalation and dermal exposures received from mixing, loading, and applying the pesticide in the field, dietary intake and ambient air were included as potential sources for exposure to this pesticide. Again, the dose received from drinking water ingestion was not included here as it was presumed to be inconsequential. Contrary to common practice, the dermal dose calculated for this case study was not based upon exposure rate, such as per pound of active ingredient handled, but on passive patch dosimetry data already extrapolated to dermal exposure for the whole body. Figure 19.5 and Table 19.5 summarize the exposure sources and the parameters used in this case study, respectively. This fourth case study was quite different from the first three cases in that biomonitoring data for the pesticide handler were available and hence were used to validate the aggregate dose estimated from the probabilistic-based integration approach. It is intuitive and has been asserted by many investigators (see Wang et aI., 1989), that well-conducted biomonitoring data will reflect all the exposures already aggregated from all sources and pathways.
19.4 RESULTS AND DISCUSSION The four case studies together include a wide range of potential pathways through which people from different sectors may be exposed to pesticides. In the first case presented, the 6-year-old
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Coping With Aggregate Pesticide Exposure Assessment: An Integration Approach
Table 19.2 Parameters and Assumptions Used for the 6-Year-Old Childa Point value
Parameter
5-95th% tile; mean
Distribution
Outdoor air at playground, ng/m3
85.00
10.0-85.0; 34.0
Lognormal
Nonplayground outdoor air, ng/m3
65.00
10.0-65.0; 30.0
Lognormal
daytime, ng/m3 Indoor air/nonsmoking - night, ng/m3
50.00
3.0-50.0; 18.0
Lognormal
15.00
3.0-15.0; 7.6
Lognormal
Inhalation absorption (default)
100%
Indoor air/smoking -
Soil concentration, ppm
120.00
5.0-120.0; 40.0
Lognormal
Soil ingestion rate, mg/hr Skin-soil loading, mg/cm2
100.00
5.0-100.0; 34.0
Lognormal
1.50
0.5-1.5; 1.0
Uniform
2.00
0.01-2.0; 0.5
Lognormal
Surface residues, ~g/cm2 Dermal transfer, ~g/hr per ~g/cm2 Dermal absorption
7,500.00 5%
Body portion as hand surface
10%
Residues portion on hand for oral
25% 100%
Oral absorption Body surface for skin-soil loading
55%
~g/person
60.00
Dietary (precalculated),
~gll
300.00
Water consumption, l/day
1.30
Water concentration,
Body weight, kg (BW)
22.25
Body surface area, cm 2 (SA)
8,552.34
Hours spent at playground
1.00
Hours spent outdoors -
1.00
nonplayground
Hours spent indoors -
at night
10.00
Hours spent indoors -
daytime
12.00
Inhalation rate -
at night, m 3/h
0.29
Inhalation rate -
other times, m 3/h
0.78
~g/person
Inhalation dose,
~g/person
Dermal uptake -
soil,
Dermal uptake -
surface,
Hand-to-mouth -
soil,
Hand-to-mouth -
surface,
Dietary intake,
Lognormal
10.0-300.0; 95.0
Lognormal
0.25-1.30; 0.78
Normal
15.5-29.0; 22.3 SA
=
(4 x BW + 7)/(BW
Normal
+ 90)
Half of which with/at soil
~g/person
~g/person
nighttime
Inhalation dose,
Aggregate dose,
~g/person
0.07
(see Eq. 2, Table 19.1)
42.33
(see Eq. 4, Table 19.1)
375.00
(see Eq. 5, Table 19.1)
6.00
(see Eq. 6, Table 19.1)
187.50
(see Eq. 7, Table 19.1)
0.05 0.05
(see Eq. 2, Table 19.1)
450.47
~g/person
Drinking water, Indoor -
~g/person
daytime
Inhalation dose,
~g/person
~g/person
nonpJayground
Inhalation dose, Indoor -
5.0-60.0; 23.0
610.90
Playground
Outdoor -
Onto the skin layer
~g/person
~g/person
0.47
(see Eq. 2, Table 19.1)
60.00
(see Eq. 1, Table 19.1)
390.00
(see Eq. 3, Table 19.1)
0.04 0.04
(see Eq. 2, Table 19.1)
1,061.46
aYalues in bold are sub- and grand totals, some of which are relisted in Table 19.6 under Point estimation; those in italics are for probabilistic analysis.
child represented nonuser residents whose exposure to pesticides is a critical part of the assessment. The exposure pathways included in the swimmer case were likewise more complex, or less traditional, than those in the last two, at least in pesticide exposure assessment. Both the harvester case and the handler
case involved individuals of the most highly exposed subpopulation which, ironically, was not covered by the original intent of the FQPA. It is fair to say, nonetheless, that a widely used pesticide could yield exposure from even a more complex scenario than one of those presented in this chapter.
19.4 Results and Discussion
~
485
@D 001
i-n-h-a-l-a-ti";;;o~n"'l
1"'"1
Jt -''potential but not included ] in this case study as [ considered inconsequential
dermal uptake Figure 19.3 Multiexposure pathways via multimedia for the adult swimmer who has been in a pool treated with a volatile antimicrobial agent.
Table 19.3 Parameters and Assumptions Used for the Adult Swimmer" Parameter
Point value
5-95th% tile; mean
Distribution Lognormal
Water concentration in pool, mg/l Air level at pool, mg/m 3
0.70
0.02--0.70; 0.15
6.40
0.04-6.40; 1.0
Lognormal
Permeability coefficient, cmlh
0.30
0.10-0.30; 0.20
Triangular
Hours in pool
3.00
0.25- 3.00; 1.00
Lognormal
Surface area for dermal, m2
2.20
1.50--2.20; 1.80
Normal
Inhalation rate - swimming, m 3/h
2.50
0.50-2.50; 1.25
Lognormal
100.0-250.0; 170.0
Triangular
Water intake - swimming, mllh
250.00
Oral absorption (default)
100%
Inhalation absorption (default)
100%
From swimming, flg/person
62,385.00
(see Eq. 8, Table 19.1 )
Dermal dose
13,860.00
Inhalation dose
48,000.00
(see Eq. 9, Table 19.1)
525.00
(see Eq. 10, Table 19.1)
Oral dose (ingestion) From drinking water, flg/person Water concentration, flg/l Water intake, l/day Oral absorption (default) From ambient air, flg/person away from pool, ng/m 3 Inhalation volume, m 3/20--22 h Ambient air -
Inhalation absorption (default)
Aggregate dose , flg/person
125.00
(see Eq. 3, Table 19.1)
50.00
15.0--50.0; 30.0
Lognormal
2.50
0.50--2.50; 1.50
Normal
100% 0.34
(see Eq. 2, Table 19.1)
20.00
5.0--20.0; 10.0
Lognormal
17.00
11.0--17.0; 13.0
Lognormal
100% 62,510.34
a Values in bold are sub- and grand totals, some of which are relisted in Table 19.7 under Point estimation; those in italics are for probabilistic analysis.
The Code of Federal Regulations (1997) requires that employees be trained to deal with potential hazards from working with pesticides. Where this type of employee safety training program is implemented, worker hand-to-mouth exposures and
those from eating produce picked off a field would be minimal and hence need not be considered in the third case. At other times, especially when the dermal dose is to be estimated from exposure rate (such as per pound of active ingredient handled),
486
CHAPTER 19
Coping With Aggregate Pesticide Exposure Assessment: An Integration Approach
.potential but not Included ] In this case study as [ considered inconsequential
Figure 19.4 field.
Multiexposure pathways via multimedia for the strawberry harvester who has worked in a treated
Table 19.4 Parameters and Assumptions Used for the Strawberry Harvester"
5-95th% tile; mean
Distribution
(see Eq. 12, Table 19.1) 0.02--0.10; 0.05 1,500-4,500; 2,750 4-12; 8
Lognormal Lognormal Triangular
16.50 3.00 11.00 50%
(see Eq. 2, Table 19.1) 1.0-3.0; 1.8 5.0-11.0; 7.5
Lognormal Lognormal
Hand-to-mouth, J.!g/person Hand portion of total dermal Residues on hand for oral Oral absorption (default)
324.00 90% 10% 100%
(see Eq. 7, Table 19.1) 0.60--0.90; 0.80 1%-10%; 5.5%
Triangular Triangular
Oral intake at field, J.!g/person Residues per strawberry, J.!g No. of berries picked and eaten Oral absorption (default)
320.00 8.00 40 100%
(see Eq. 11, Table 19.1) 0.3-8.0; 2.25 0-40;20
Lognormal Custom/discrete
From ambient air, J.!g/person Ambient air, ng/m3 Inhalation volume, m 3 /16 h Inhalation absorption (default)
0.63 90.00 14.00 50%
(see Eq. 2, Table 19.1) 10.0-90.0; 37.5 7.0-14.0; 10.0
Lognormal Lognormal
100.00 100.00 100%
(see Eq. 1, Table 19.1) 10.0-100.0; 34.0
Lognormal
Parameter Dermal dose, J.!g1person Foliar residues, J.!g/cm 2 Transfer factor, J.!g/h per J.!g1cm 2 Hours worked Dermal absorption (default) Inhalation dose, J.!g1person Air level at field, J.!g/m 3 Inhalation volume, m 3 /8 h Inhalation absorption (default)
Dietary intake, J.!g/person Dietary (precalculated), J.!g/person Oral absorption (default)
Aggregate dose, J.!g/person
Point value 1,800.00 0.10 4,500.00 8.00 50%
2,561.13
aValues in bold are sub- and grand totals, some of which are relisted in Table 19.8 under Point estimation; those in italics are for probabilistic analysis.
the fourth case will need to include several more parameters than presented here, such as the amount of active ingredient handled per day or per hour. As discussed by Hamey et al. in Chapter 22, PHED (1995) is a data base widely used in pesticide exposure assessment to generate surrogate exposure rates
for pesticide handlers, especially when chemical-specific exposure data are limited. In this chapter, the focus is on absorbed dose received in a single day. If chronic or lifetime exposure were considered, more parameters would be needed in each of the four cases.
19.4 Results and Discussion
487
oral intake
de rmal uptake
·potential but not included in this case] [ study as considered inconsequential
Figure 19.5 Multiexposure pathways via multimedia for the agricultural handler who has mixed. loaded. and applied a pesticide in a field. Table 19.5 Parameters and Assumptions Used for the Agricultural Handler" Parameter Dermal dose, flg/person Dermal residues, mg/person
Point value
56,250.00 225.00
5-95th% tile; mean
Distribution
(see Eq. 13, Table 19.1) 50.0-225.0; 120.0
Lognormal
15%-25%; 20%
Normal
Dermal absorption (rat)
25%
Inhalation dose, flg/person
605.00
(see Eq. 2, Table 19.1)
110.00
50.0-110.0; 76.0
Lognormal
5.0-11.0; 7.6
Lognormal
Air level at site, flg/m3 Inhalation volume, m 3/8 h Inhalation absorption (default) From ambient air, flg/person Ambient air, ng/m3 Inhalation volume, m 3/16 h Inhalation absorption (default) Dietary intake, flg/person Dietary (precalculated), flg/person
11.00 50%
0.35
(see Eq. 2, Table 19.1)
50.00
5.0-50.0; 20.0
Lognormal
14.00
7.0-14.0; 10.0
Lognormal
50%
3.00 3.00
Oral absorption (default)
100%
Aggregate dose, flg/person
56,858.35
(see Eq. 1, Table 19.1) 1.0-3.0; 1.8
Lognormal
aValues in bold are sub- and grand totals, some of which are relisted in Table 19.9 under Point estimation; those in italics are for probabilistic analysis.
Two such critical additional parameters are annual exposure frequency and the total years in which a person would be exposed to the pesticide. For chronic or lifetime exposure, a calendarbased model similar to the one described by Shurdut et at. (1998) may be warranted. The model by Shurdut et al. simulates the aggregate exposure for each of the 365 days first and then combines these 365 daily aggregate exposures into a single annual aggregate exposure distribution. Their model also differs from the integration exposure model presented in this chapter in that they take conditional probabilities of exposure into consideration in simulating the daily aggregate exposures. This chapter focuses primarily on the conceptual framework of aggregating exposure from multiple pathways, and hence more
cumbersome probabilistic methods to determine the conditional likelihood of receiving multiple exposures on a given day are not used. The conditional probability model asserts that there is a significant correlation between, for example, the exposure to the wood preservative received by children at the playground and the exposure that they would receive at home on that same day. The main difficulty of computing the conditional probabilities here is that these path-to-path daily correlations require a great deal of knowledge of daily, seasonal, and regional use frequencies and patterns. For the above four case studies, each aggregate dose from the potential sources was estimated separately using the three approaches described in Section 19.2 (and outlined graphically
488
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Table 19.6 Aggregate Dose Estimated for the 6-Year-Old Child, fig/person
Method
Playground
Outdoor
Indoor (day)
Indoor (night)
All day Deterministic accumulation
Point estimation
610.9
0.05
450.5
0.04
1,061.5 Probabilistic addition
Probabilistic
444.7
0.05
251.5
0.04
696.2 Probabilistic integration
Probabilistic
(8.0, 444.7)a
(0.01,0.05)a
(19.2,251.5)Q
(0.01,0.04)Q
542.2
Values presented in parentheses above for the various outdoor and indoor activities are the 5th and the 95th percentiles averaged over 10 Monte Carlo simulation trials each consisting of 10,000 runs performed under PROBABILISTIC ADDITION [see Dong et al. (1994a, b) for a general description of the simulation technique and the definitions used]; these percentiles, together with the assumption of a lognormal distribution, were reused to run the final 10 probabilistic simulation trials that in turn yielded a 95th percentile average for the all day aggregate dose as presented here under PROBABILISTIC INTEGRATION.
a
Table 19.7 Aggregate Dose Estimated for the Adult Swimmer, fig/person
Method
Dermal dose
Inhalation
Oral dose
Nonpool areaa
All day Deterministic accumulation
Point estimation
13,860.0
48,000.0
525.0
125.3
62,510.3 Probabilistic addition
Probabilistic
2,512.3
5,510.5
105.7
90.7
8,219.2 Probabilistic integration
Probabilistic
(49.2,2,512.3)b
(31.1,5,510.5)b
(2.2, 105.7)b
(13.4,90.7)b
6,853.1
Including doses from drinking water and ambient air, respectively: point estimation, 125.0, 0.3; probabilistic for addition, 90.5, 0.3; and probabilistic for integration, (0.3, 90.5), (0.06, 0.25). b Values presented in parentheses above for the various routes are the 5th and the 95th percentiles averaged over 10 Monte Carlo simulation trials (see footnote a in Table 19.6 for their use).
a
Table 19.8 Aggregate Dose Estimated for the Strawberry Harvester, fig/person
Method
Dermal dose
Inhalation
Hand/mouth
Oral at field
All daya Deterministic accumulation
Point estimation
1,800.0
16.5
324.0
320.0
2,561.1 Probabilistic addition
Probabilistic
1,132.6
12.0
141.4
157.6
1,538.6 Probabilistic integration
Probabilistic
(160.0, 1,132.6)b
(3.3, 12.0)b
(11.5, 141.4)b
(1.5, 157.6)b
1,355.0
Including doses from dietary intake and ambient air away from the field (i.e., during nonwork hours), respectively: point estimation, 100.0,0.6; probabilistic for addition, 86.7,0.6; and probabilistic for integration, (5.0,86.7), (0.07, 0.57). b Values presented in parentheses above for the various routes are the 5th and the 95th percentiles averaged over 10 Monte Carlo simulation trials (see footnote a in Table 19.6 for their use).
a
19.4 Results and Discussion
489
Table 19.9 Aggregate Dose Estimated for the Pesticide Handler, fLg/person Method
Dermal dose
Inhalationa
Ambient air
Dietary intake
All day Deterministic accumulation
Point estimation
56,250.0
605.0
0.35
3.0
56,858.4 Probabilistic addition
Probabilistic
44,944.4
475.0
0.24
2.9
45,422.5 Probabilistic integration
Probabilistic
(10,575.0, 44,944.4)b
(158.1, 475.0)b
BiomonitoringC
(0.02,0.24)b
(1.0,2.9)b
44,993.5 Urine analysis 7,040.0
a While working in the field. b Values presented in parentheses above for the various routes are the 5th and the 95th percentiles averaged over 10 Monte Carlo simulation trials (see footnote a in Table 19.6 for their use). C From a well-conducted biomonitoring study submitted in support of the reregistration of the pesticide.
in Fig. 19.1). The results from the three aggregate exposure approaches are summarized in Tables 19.6 through 19.9, with one case per table. Below are several important observations made from these results. Where only one random variable predominated, the 95th percentile dose simulated from the probabilistic analysis was found to be very close to the point estimate calculated for an extreme case scenario. This was the case with the nighttime indoor and nonplayground outdoor exposure scenarios for the 6-year-old child and with the aggregate dose for the pesticide handler. In all cases, the aggregate doses estimated by the probabilistic addition approach were somewhat higher than those simulated by the probabilistic integration approach. The swimmer case indicated that the conventional accumulation approach could yield an aggregate dose 10 times higher than those estimated by the probabilistic-based integration or addition approach. As indicated in Table 19.7, this lO-fold difference was primarily a result of the five- to eightfold lower dermal and inhalation doses simulated by the probabilistic approaches. In this swimmer case, the simulated dermal and inhalation doses were considerably lower because, as indicated in Table 19.3, both the water concentration in the pool and the air level at the pool were assumed to have a mean considerably lower than their 95th percentile. It is intuitive that where in a (lognormal) distribution the mean is far below the upper tail end (e.g., the 95th percentile), even a slight drop in percentile in the upper end will result in a significant drop in numerical value. Also of note is the fact that where two events are mutually exclusive, the probability of the joint event is simply the product of the probabilities of the two events occurring. In other words, a 95th percentile inhalation dose could result easily from an air level in the 70th or the 80th percentile range if the value used for breathing rate were also past the midrange. In each case, there were at least a few exposure sources that need not be included, as they did not affect the aggregate dose estimate much. Sensitivity analyses confirmed that the follow-
ing exposure parameters were most influential: for the first case, dislodgeable residues on treated playground structure surface; for the second case, water concentration in the pool; for the third case, foliar dislodgeable residues; and for the fourth case, dermal residues (those that had penetrated through the work clothes onto the skin). It is of note that the exposure parameters identified above do not necessarily represent the factors with the greatest variability, since sensitivity analysis is also used to discount those factors that have little (numerical) contribution to the outcome of interest (in this case the aggregate dose). In the last three cases, the background doses (i.e., those from dietary intake, ambient air, drinking water, etc.) appeared to contribute very little to the aggregate dose. On the other hand, the daily dose received by the 6-year-old child doubled after the child spent just 1 hour at a site where playground structures were treated. At least with the handler case here, biomonitoring seems to be the most convincing, if not the most effective, integration approach for calculating the aggregate dose. As indicated in Table 19.9, the high-end daily aggregate dose estimated by the biomonitoring approach for the pesticide handler was approximately 6 times less than the aggregate dose simulated by the probabilistic-based integration approach. This sixfold difference can be explained rather convincingly by the several factors inherent in the patch dosimetry data used in this chapter. According to a review on a handful of compounds tested and available, absorption in the rat was found to overestimate human dermal absorption by two- to ninefold (Wester and Maibach, 1993). Another potential factor is that the human outermost skin layer (i.e., the stratum corneum) is not as good an adsorbent to residues as clothing or cloth patch dosimeters. For one thing, the sandwiched residues that have penetrated through the (work) clothing onto the skin are subject to a greater degree of disturbance (or attrition) from body movement. Aggregate doses derived from well-designed biomonitoring studies are considered to be most convincing because such stud-
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Coping With Aggregate Pesticide Exposure Assessment: An Integration Approach
ies reduce the uncertainty with animal (versus human) dermal absorption or with other types of extrapolation, such as that from patch residues to dermal residues for an entire body region or from foliar residues to dermal residues. Biomonitoring studies also reduce the concern about events that may occur concurrently or those that cannot happen at the same time. Urine analysis is the most commonly used means for biological monitoring of exposure, as urine collection is noninvasive and easier to handle compared to the collection of other types of biological specimens from humans. Despite these advantages, routine collection of 24-hour urine samples in humans is often impractical. The measurement of the chemical and its major metabolite(s) in urine is accordingly often necessarily performed on spot specimens. Using physiologically based pharmacokinetic modeling, Dong et al. (1994c, 1996) provided examples demonstrating how spot urine sample results can be used to simulate the amount of a pesticide dermally absorbed in humans. It should be pointed out, though, that biomonitoring per se is not always a practical tool for measuring aggregate exposure. For many chemicals, the underlying pharmacokinetics are not readily available or well understood. In addition to such a deficiency and the difficulty of collecting urine or other biological specimens from humans, the lack of an acceptable analytical tool for quantifying the biomarker of interest often invalidates biomonitoring as an effective integration approach. It is these drawbacks of biomonitoring that have given the chapter's integration exposure model a good place in aggregate exposure assessment. Where food was considered as a potential source in the case studies, dietary intake was estimated using the dietary exposure analysis software Exposure 4, a microcomputer package developed by Technical Assessment Systems, Inc. (TAS, 1990) to assess dietary intake as part of the pesticide risk characterization process. The distribution of a dietary intake can also be simulated using the framework of Higley and Strenge (1993) or that described by Petersen and Tomerlin in Chapter 17. The case studies in this chapter did not use either of these latter frameworks to fine tune any of the dietary intake distributions, as their purpose was illustrative in nature. Furthermore, as shown in Tables 19.6, 19.8, and 19.9, the dietary intakes as estimated were inconsequential compared to the aggregate dose of which they were a part. Ingestion of drinking water can be a significant exposure source compared to other types of background or residential exposures, especially when the daily aggregate exposure is relatively low. As illustrated in the first case study, the dose received from drinking water ingestion by the 6-year-old child accounted for as much as one-third of the total daily exposure. The water concentration for this case study was presumed to follow a lognormal distribution (see Table 19.2) and to originate from an unspecified water source. The impact on the concentration of drinking water source being ground water, surface water, or a blend of the two has been discussed at some length in ILSI's workshop on aggregate exposure assessment (1998).
In terms of probabilistic analysis, no noticeable difference was found in the simulation results between using the conventional Monte Carlo sampling and the more accurate Latin hypercube sampling method, both of which are used for generation of random numbers. Both random number sampling methods yielded similar simulation results primarily because each of the simulation trials consisted of a rather large number of random samplings (i.e., 10,000 runs). For the four case studies, the 5th and the 95th percentiles were used to describe a normal or lognormal input distribution. Other alternative pairs of distribution parameters that can be used to yield the same simulation effect include, but are not limited to the 10th and the 90th percentiles, the (geometric) mean and the 95th percentile, and the mean and the standard deviation.
19.5 CONCLUSIONS The four case studies suggest that although a person could be exposed to pesticides through numerous complex routes of entry, some exposure pathways play a more predominant role than others. Unless more than one predominant route is involved, a probabilistic simulation for route- and medium-specific exposures or for the aggregate dose may not be warranted. For example, unless the chemical of interest has a high vapor pressure, the contribution from inhalation to absorbed dose will be negligible. In that case, the task of performing an aggregate exposure assessment may be reduced to estimating the dose received from the predominant exposure route or medium. In some other use scenarios, the conventional accumulation approach could yield an aggregate dose 10 times higher than those simulated by the probabilistic addition or integration approach, as illustrated in the swimmer case here. Where feasible, well-designed as well as well-conducted human biomonitoring studies should offer the most effective estimate for the aggregate dose in question.
REFERENCES Code of Federal Regulations (1997). "Worker Protection Standard," Part 170, 40 CFR. U.S. Gov. Printing Office, Washington, DC. Copeland, T. L., Holbrow, A. M., Otani, J. M., Connor, K. T., and Paustenbach, D. J. (1994). Use ofprobabilistic methods to understand conservatism in California's approach to assessing health risks posed by air contaminants. J. Air Waste 44, l399-14l3. Crystal Ball (1996). Crystal Ball™ Forecasting and Risk Analysis for Spreadsheet Users, version 4.0. Decisioneering, Inc., Aurora, CO. Dang, W. T. (1996). "The Swimmer Exposure Assessment Model (SWIMODEL) and Its Use in Estimating Risks of Chemical Use in Swimming Pools," U.S. EPA internal guidance document. Occupational and Residential Branch, Health Effects Division, U.S. EPA Office of Pesticide Programs, Washington, DC. Dong, M. H., Haskell, D., Ross, J. H., Schneider, F., Hernandez, B. Z., and Benson, C. (1994a). "Preliminary Human Pesticide Exposure Assessment Diazinon (for Use on Residential Turf and Soil)," HS-1694. Worker Health and Safety Branch, CaIJEPA Dept. of Pesticide Regulation, Sacramento. Dong, M. H., Ross, J. H., Schneider, F., Hernandez, B. Z., Haskell, D., Thongsinthusak, T., and Sanborn, J. R. (1994b). "A Probabilistic Approach
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to Estimating Exposure Potential for Children Playing on Diazinon-Treated Residential Soil," HS-1690. Worker Health and Safety Branch, CalIEPA Dept. of Pesticide Regulation, Sacramento. Dong, M. H., Draper, W. M., Papanek, P. J., Ross, J. H., Woloshin, K. A., and Stephens, R. D. (1994c). Estimating malathion doses in California's medfly eradication campaign using a physiologically based pharmacokinetic model. In "Environmental Epidemiology: Effects of Environmental Chemicals on Human Health" (W. M. Draper, ed.), ACS Advances in Chemistry Series No. 241, pp. 189-208. Am. Chem. Soc., Washington, DC. Dong, M. H., Ross, J. H., Thongsinthusak, T., and Krieger, R. I. (1996). Vse of spot urine sample results in physiologically based pharmacokinetic modeling of absorbed malathion doses in humans. In "Biomarkers for Agrochemicals and Toxic Substances" (J. N. Blancato, R. N. Brown, C. C. Dary, and M. A. Saleh, eds.), ACS Symposium Series, Vol. 643, pp. 229-241. Am. Chem. Soc., Washington, DC. Finley, B., and Paustenbach, D. (1994). The benefits of probabilistic exposure assessment: three case studies involving contaminated air, water, and soil. Risk Anal. 14, 53-73. Higley, K. A., and Strenge, D. L. (1993). Vse of a Monte Carlo modeling approach for evaluating risk and environmental compliance. In "Effective and Safe Waste Management: Interfacing Sciences and Engineering with Monitoring and Risk Analysis" (R. L. Jolley and R. G. M. Wang, eds.), pp. 337-347. Lewis Publishers, Boca Raton, FL. International Life Sciences Institute (ILSI). (1998). "Aggregate Exposure Assessment: An ILSI Risk Science Institute Workshop Report." International Life Sciences Institute, Washington, DC. National Research Council (NRC). (1983). "Risk Assessment in the Federal Government: Managing the Process." Natl. Acad. Press, Washington, DC. Office of Environmental Health Hazard Assessment (OEHHA). (1996). "Public Review Draft: Air Toxics Hot Spots Program Risk Assessment Guidelines Part IV - Technical Support Document for Exposure Assessment and Stochastic Analysis." CalIEPA Office of Environmental Health Hazard Assessment, Sacramento. Pesticide Handlers Exposure Database (PHED). Version 1.1. (1995). Prepared for Health Canada, V.S. EPA, and American Crop Protection Association by Versar Inc., Springfield, VA. Ross, J. H., Powell, S., Thongsinthusak, T., Haskell, S., Dow, M. I., and Worgan, J. (1997). "Harmonization of Issues Involving Pesticide Exposure Assessment in North America," HS- I 763. Worker Health and Safety Branch, CaIlEPA Dept. Pesticide Regulation (presented at the 213th American Chemical Society National Meeting jointly with U.S. EPA and Health Canada, Poster No. 112, April 13-17, San Francisco, CA). Shurdut, B. A., Barraj, L., and Francis, M. (1998). Aggregate exposures under the Food Quality Protection Act: An approach using chlorpyrifos. Reg. Toxieol. Pharmaeol. 28, 165-177. Technical Assessment Systems (TAS). (1990). TAS Exposure 4™ - Detailed Distributional Dietary Exposure Analysis. Technical Assessment Systems, Washington, DC. Thompson, K. M., Burmaster, D. E., and Crouch, E. A. C. (1992). Monte Carlo techniques for quantitative uncertainty analysis in public health risk assessments. Risk Anal. 12, 53-63.
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Thongsinthusak, T., Ross, J. H., and Meinders, D. (1993). "Guidance for the Preparation of Human Pesticide Exposure Assessment Documents," HS-1612. Worker Health and Safety Branch, CalIEPA Dept. of Pesticide Regulation, Sacramento. V.S. Environmental Protection Agency (VSEPA). (1994). "Notice to Manufacturers, Producers, Formulators, and Registrants of Pesticide Products Announcing the Formation of Two Industry-Wide Task Forces: Agricultural Reentry Task Force and Outdoor Residential Exposure Task Force," Pesticide Regulation (PR) Notice 94-9. V.S. EPA Office of Pesticide Programs, Washington, DC. V.S. Environmental Protection Agency (VSEPA). (1996). "Descriptive Statistics Tables from a Detailed Analysis of the National Human Activity Pattern Survey (NHAPS) Data," EPN600/R-96/148. V.S. EPA Office of Research and Development, Washington, DC. V.S. Environmental Protection Agency (VSEPA). (1997a). "Guiding Principles for Monte Carlo Analysis," EPN630/R-97/001. V.S. EPA Risk Assessment Forum, Washington, DC. V.S. Environmental Protection Agency (VSEPA). (l997b). "Exposure Factors Handbook," EPN600/P-95/002. V.S. EPA Office of Health and Environmental Assessment, Washington, DC. V.S. Environmental Protection Agency (VSEPA). (1998a). "Pirimiphos-Methyl (Chemical ID No. 108102IList B Reregistration Case No. 2535) - HED Human Health Risk Assessment and Supporting Documentation for the Reregistration Eligibility Decision Document (RED). No MRID #. DP Barcode Nos. D240741 and D241203," memo from Christina B. Swartz of Health Effects Division (HED), dated October 23. V.S. EPA Office of Pesticide Programs, Washington, DC. U.S. Environmental Protection Agency (VSEPA). (l998b). "Tetrach1orvinphos (Chemical ID No. 083701IList A Reregistration Case No. 0321) - Addendum to the HED Human Health Risk Assessment and RED Chapter dated 4/1/98. DP Barcode No. D249577," memo from Christina B. Swartz of Health Effects Division (HED), dated November 2. V.S. EPA Office of Pesticide Programs, Washington, DC. V.S. Environmental Protection Agency (VSEPA). (1999). "Methidathion Revised Short Format HED Chapter of RED. Chemical Number 100301. DP Barcode D252049," memo from Robert Travaglini of Health Effects Division (HED), dated January 13. V.S. EPA Office of Pesticide Programs, Washington, DC. Wang, R. G. M., Franklin, C. A., Honeycutt, R. c., and Reinert, J. C., eds. (1989). "Biological Monitoring for Pesticide Exposure: Measurement, Estimation, and Risk Reduction," ACS Symposium Series, Vol. 382. Am. Chem. Soc., Washington, DC. Wester, R. c., and Maibach, H. I. (1993). Animal models for percutaneous absorption. In "Health Risk Assessment: Dermal and Inhalation Exposure and Absorption of Toxicants" (R. G. M. Wang, J. B. Knaak, and H. I. Maibach, eds.), pp. 89-103. CRC Press, Boca Raton, FL. Whitmyre, G. K., Driver, J. H., Ginevan, M. E., Tardiff, R. G., and Baker, S. R. (1992). Human exposure assessment I: Vnderstanding the uncertainties. Toxieol. Ind. Health 8, 297-320.
CHAPTER
20 Occupational Exposure Data Bases/Models for Pesticides Gary K. Whitmyre risksciences, LLC
John H. Ross infoscientific.com, Inc.
Curt Lunchick Aventis CropScience
20.1 INTRODUCTION The use of pesticides to control insect pests, fungal diseases, and undesired plant pests (i.e., weeds) is an integral part of modem agricultural practices. Because pesticides are biologically active agents, it is essential to determine the human health risks incurred by the workers involved in use of pesticide products. To quantify the risks associated with pesticide use, it is important to understand not only the toxicity of the particular active ingredient, but also the actual magnitude of the anticipated exposure. Human exposures to pesticides may occur during worker contact involving principally the dermal and inhalation exposure routes. Worker populations that are routinely exposed to pesticides include agricultural handlers (or "operators" in Europe) involved in treatment of field crops, greenhouse crops, vineyards, and orchards, professional grounds applicators (e.g., parks and roadsides), lawn care professionals, structural and commercial applicators (e.g., for factories, food processing plants, hotels, hospitals, other institutions, offices, and residences), and field workers, e.g., during harvesting or canopy management of treated crops (Driver and Whitmyre, 1997; Maddy et aI., 1990; U.S. EPA, 1984). The requirement to quantify worker exposures to active ingredients during use of pesticide formulations is an integral part of risk assessments associated with the pesticide registration process in the United States, Canada, the European Union (EU), and other countries (e.g., Australia). The risk assessment process enables regulatory agencies and the agrochemical industry to predict the extent of risk of adverse human health effects associated with the use of a given pesticide under specific use conditions. Because the evaluation of risk requires knowledge of both exposure and toxicity, exposures to the active ingredient Handbook oJ Pesticide Toxicology
Volume 1. Principles
associated with a given pesticide formulation must be assessed. This chapter provides a summary of the most commonly used worker exposure data bases and models for mixing/loading and application of pesticides. Further, efforts to develop a generic data base for field worker exposures during reentry activities (e.g., for scouting or harvesting) following application are described, in general.
20.2 TIERED APPROACH As with other applications of the risk assessment process, a tiered approach to estimating exposures is often taken. Although there is no universal tiered approach to risk assessment of worker exposures, one commonly used system can be described as follows. The first tier (tier I) of the risk assessment process for worker exposures to pesticides can involve the use of generic exposure data from data bases such as the Pesticide Handlers Exposure Database (PHED) developed and used in North America (U.S. EPA, 1995a), the Predictive Operator Exposure Database (POEM) developed in the United Kingdom (Martin, 1986), and EUROPOEM (AIR, 1996). PHED provides actual measured dermal and inhalation exposure data that can be retrieved for specific subseUing conditions (e.g., open mixing or open cab groundboom application). POEM provides an estimate of exposure derived from actual exposure monitoring data based on the type of formulation, application method, application rate, etc. Because EUROPOEM was not formally adopted by EU member states for assessment of operator exposures under the 911414IEC legislation, industry submissions to the EC for registration or reregistration are still based largely on the U.K.'s POEM model and the German model (Lundehn
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CHAPTER 20 Occupational Exposure Data BasesIModels
et aI., 1992; FBRC, 1993). The national pesticide registration procedures in the Netherlands use a published Dutch operator exposure model (van Goldstein Brouwers et aI., 1996); submissions are to follow an exposure assessment protocol published by TNO, specifically TNO Report V 96.120 (van Goldstein Brouwers et aI., 1996). In France, the U.K.'s POEM model is used to estimate worker exposures, but only for unprotected operator scenarios. All of these data bases/models are founded on similar principles in that they rely on measured exposure data from various studies that are combined into data sets for each worker exposure scenario. The estimated exposure from the appropriate surrogate data base/model is then used in combination with the key toxicological benchmarks [e.g., no observed adverse effectlevels (NOAELs)] to determine whether an adequate margin of safety or margin of exposure exists for the worker exposure scenario of interest. If exposure data of adequate quality and quantity exist, a tier I assessment using a generic exposure data base may be refined further using a stochastic (probabilistic) approach. If an unacceptable risk is obtained with a tier I assessment, it may be necessary to proceed with a more refined exposure assessment or to collect formulation-specific exposure data (tier 11). In the latter case, tier 11 typically involves collection of external dosimetry data for workers on the formulation of interest, for specific use patterns of interest, thus circumventing the need to use surrogate monitoring data (for example, from PHED or POEM) because compound-specific exposure data are collected. The exposure monitoring studies have been conducted on a variety of pesticides using commercial applicators, farmers, field workers, pest control operators, professional lawn care operators, and homeowners engaged in normal work activities. Several general approaches to quantifying exposures to pesticides for these individuals have been used, including the use of (1) patch (e.g., gauze pad) dosimetry, (2) glove dosimeterlhand rinse techniques, and (3) whole body dosimetry. Hand washes, patch dosimetry, or whole body dosimeters are methods for quantifying the amount of pesticide that contacts the skin or clothing of a worker and thus, when adjusted for protection of specific body parts by clothing, provides a measure of external dermal exposure. The use of whole body dosimeters, which are usually sectioned into standard body part areas (e.g., upper legs or lower legs) before extraction and analysis, prevents the need for extrapolation from a small patch size to the whole body part. In addition, fixed location or personal air monitoring devices have been used to characterize exposures via the inhalation route by collecting a known volume of air in the breathing zone of the worker and analyzing for the mass of pesticide of interest present. Because dermal exposures represented by the dosimetry data are external exposures, it is necessary to apply assumptions on clothing protection factors and dermal absorption to estimate the absorbed dermal dose, unless a dermal toxicity study is available that provides the NOAEL for an applied dose. If the tier II assessment still does not indicate acceptable risk, the absorbed dose can be measured directly by means of a
biomonitoring study (tier Ill). Biomonitoring, also known as biological monitoring, typically uses the amount of pesticide (or its metabolites) detected in the urine of exposed individuals to obtain an accurate measurement of the total amount of pesticide actually absorbed by the worker via all routes (inhalation, dermal, and incidental oral ingestion). Although biomonitoring provides total absorbed doses (i.e., pesticide levels in the body) as a result ofthe exposure at the body's boundaries (e.g., skin or lungs), it does not explain the contributions of each specific exposure pathway. However, biomonitoring data prevent the need for extrapolation from external dosimetry to internal dose; thus, biomonitoring data provide a less conservative and more meaningful measure of exposure.
20.3 BASIS FOR USE OF GENERIC EXPOSURE MODELING A number of field studies of varying quality have been conducted in the United States, Canada, Europe, and elsewhere that provide surrogate data for estimation of worker exposures to pesticides. Occupational exposure data for selected pesticides can be found in scientific publications, registration standards, and special review documents published by the U.S. Environmental Protection Agency (EPA) and evaluations by California's Department of Pesticide Regulation, Worker Health and Safety Branch (e.g., Curry et aI., 1995; Ecobichon, 1999; Fong and Krieger, 1988; Honeycutt et aI., 1985; Krieger et aI., 1990; Mehler et aI., 1991; Nutley and Cocker, 1993; Plimmer, 1982; Rech et aI., 1988; Saleh et aI., 1994; Thongsinthusak et aI., 1993; U.S. EPA, 1997; van Hemmen, 1992; Wang et aI., 1989). The basic premise of the generic modeling approach is that worker exposures are a function more of the work activity, application equipment, formulation type, packaging type, level of clothing, total amount of active ingredient handled, and individual work practices than of the specific physical-chemical properties of the active ingredient. Thus, measured dermal and inhalation exposures from a given set of studies on surrogate active ingredients can be used to approximate the worker exposures to a given active ingredient under similar use conditions. Although formulation-specific agricultural worker exposure studies (e.g., mixer/loader, applicator, or harvester) may sometimes be necessary for pesticide registrations and reregistrations, exposure data on surrogate compounds for a given worker exposure scenario are often accepted. Selected exposure monitoring data form the basis for data bases and models that allow estimation of exposures for specific application methods, use conditions, and formulation types. A generic exposure data base provides several advantages to registrants, in that (1) they can use the data base as a tool to estimate potential exposures and associated risks early in the product development stage and (2) they can use the surrogate data in the data base to support registration and reregistration submissions to regulatory agencies, thus preventing expenditure of resources for generating compound-specific exposure data, and reducing time to registrationlreregistration approval. Because of the larger pool
20.4 Pesticide Handlers Exposure Database (PHED)
of data that a generic data base provides, greater reliability of the exposure estimates results and a better understanding of variability in the exposure estimates is obtained. In addition, a generic data base allows examination of the physical parameters and use conditions/work practices that affect exposure. Furthermore, a generic data base provides a common basis for both industry and the regulatory agencies to arrive at the same exposure estimates using a common data source and method. In the European and North American generic data bases/ models, the exposure data are provided in generic form and are incorporated into the exposure model or summarized in normalized units [e.g., )l.g/lb or kg of active ingredient (a.i.)]. The North American regulatory agencies (U.S. EPA, California EPA, and Health Canada) and European regulators have long endorsed this policy of using "surrogate" or "generic" exposure data for estimating worker exposures for the same work activity and work conditions, because it is impractical to conduct a field exposure monitoring study for every pesticide formulation. European data bases such as POEM (Martin, 1986) and the German model (FBRC, 1993) have been available for several years; these models are based on broad generic default values. The Europeans recognized the limitations of these data bases for applicability to EU-wide conditions and have focused their efforts on developing a scenario-specific mixer/loader/applicator exposure data base. These concerns and resulting refinements are reflected in the more recent tool known as EUROPOEM (AIR, 1996). Efforts are also underway through industry task forces in the United States to develop exposure data bases for different types of reentry activities. Each type of activity (e.g., weeding, thinning, reaching, and harvesting) and crop type combination are associated with different rates of contact with treated foliage, known as the transfer coefficient (cm21h). The Agricultural Reentry Task Force (ARTF) is attempting to develop through existing field studies and through commissioning of new studies, generic transfer coefficients that can be used in conjunction with compound-specific dislodgeable foliar residue (DFR) data to estimate dermal exposures of reentry workers for specific work activities. Similarly, the Outdoor Residential Exposure Task Force has performed a number of studies to characterize exposures to lawn chemicals on treated turf; these studies will be used by the member companies and by the regulatory agencies.
20.4 PESTICIDE HANDLERS EXPOSURE DATABASE (PHED) 20.4.1 HISTORICAL BACKGROUND
In 1983, the Public Health and Toxicology Committee of the National Agricultural Chemicals Association (NACA) held a workshop on pesticide worker exposure. Following this workshop, the Subcommittee on Field Exposure Assessment was formed for the dedicated purpose of developing a standard protocol for conducting field studies of mixer/loader/applicator
495
exposures (Day, 1991). At the first meeting of the Subcommittee, Drs. Hackathorn and Eberhart from Mobay presented findings from a literature search of relevant exposure studies. An analysis of the data from relevant literature studies indicated that where data were available for a given use pattern (e.g., mixing/loading and groundboom application of liquid formulations), the magnitude of exposure was independent of the physicochemical nature of the active ingredient (Day, 1991). Thus, they conjectured that if there were enough data available for a given use pattern, it should be possible to use these data as surrogate data that would be universally applicable to estimation of exposures for that use pattern (Day, 1991). Therefore, new field studies would not need to be conducted on every formulation that was presented to the U.S. EPA for registration. The use of generic exposure data and the development of a worker exposure data base were proposed and discussed at the 187th meeting of the American Chemical Society in April 1984. Three important papers were presented (Hackathorn and Eberhart, 1985; Honeycutt, 1985; Reinert and Severn, 1985), which proposed how a generic exposure data base could be used in the risk assessment process (Lunchick et aI., 1994). As a result of these proposals to develop a generic data base for applicator exposures, a task force consisting of representatives from Health and Welfare Canada (now Health Canada), the U.S. EPA, and the NACA [now the American Crop Protection Association (ACPA)] was eventually formed to oversee the design and development of what would ultimately become PHED. Industry agreed to allow proprietary exposure data to be included in the data base as long as the identity of the specific chemicals was kept confidential. Because the exposure studies would represent a variety of study designs, there was a need to anticipate the various data set types and the fields that would be necessary in PHED to hold the data. Furthermore, there was discussion as early as 1985 regarding the need to rank the data based on adequacy of quality assurance procedures, so that users would be able to create subsets from the data based on quality assurance ranking or grade. A joint EPA-industry task force was formed, headed initially by Joe Reinert of the U.S. EPA. A generic data base concept paper was presented at the Sixth International Congress of Pesticide Chemistry sponsored by IUPAC in Ottawa, Canada, August 10-15, 1986. Requests were made to industry members through the NACA (1987) for submission of studies to the Task Force (through the U.S. EPA) for entry into the data base. Registrants were asked to sign a data compensation waiver and to enter key data from their studies on a form that would later be used by a U.S. EPA contractor as a data entry form. A number of years later, a data entry diskette was developed by the U.S. EPA contractor (Versar, 1991), which would provide a mechanism for registrants to submit their data to the task force in electronic form. 20.4.2 OVERVIEW OF PHED
PHED was developed to provide dermal and inhalation exposure data for mixing/loading and application of pesticides.
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PHED was developed as a Jomt effort between the D.S. EPA, Health Canada, and NACA (now ACPA). PHED was first released for general use in 1992 as a tool for assessing mixerlloader and applicator exposures. PHED is being used by registrants and government agencies to supplement or replace worker exposure studies and as a validation tool for field exposure data. PHED can be useful for registrants for supplementing or combining with field exposure studies, estimating exposures during the early stages of product development (e.g., registrability assessments for alternative types of formulations and levels of active ingredient), and for fulfilling data requirements imposed by the regulatory agencies (D.S. EPA, Health Canada, and the California EPA). PHED was initially released (version 1.0) in May 1992 (Lunchick et aI., 1994). PHED, version 1.1 (D.S. EPA, 1995a), which was released in 1995, contains more data than version 1.0 (over 1700 records) on measured dermal and inhalation exposures and on various parameters that may affect the magnitude of exposures. Each data record represents one replicate for one worker involved in one work day or less of a given activity. There are four separate data files in PHED where the worker exposure data reside, depending on which work activities were monitored, as shown in Fig. 20.1. These include the MixerlLoader (556 replicates), Applicator (715 replicates), MixerlLoader/Applicator (349 replicates), and Flagger (92 replicates) files. PHED can be used to develop unit exposures [J.Lg/lb active ingredient (a.i.) handled] for a specific worker exposure scenario. PHED provides a useful tool for mode ling and predicting potential pesticide exposures based on consideration of numerous factors, such as application rate, formulation type and packaging, mixing/loading methods, application methods and equipment, and the type of clothing or protective equipment used. Guidelines have been developed for proper use and reporting of PHED data (D.S. EPA, 1995b).
92 Replicates 6% MxerJloacEr 556 Replicates
32"10
Applicator 7) 5 RepOCates 42"10
Figure 20.1
Distribution of PHED records.
20.4.3 PHED GRADING CRITERIA One of the aspects of data contained within PHED is that each data replicate contained within it is classified with regard to quality. The data-grading scheme developed by the D.S. EPA, Health Canada, and ACPA assigns letter grades (A through E) based on analytical quality assurance procedures. By basing the grading exclusively on analytical quality assurance, all valid data points, regardless of the completeness of the study or the quality of its parts, are available for use. The guidelines for using PHED (D.S. EPA, 1995b) specify that grades A and B data should be used. Class A data are those data for which (1) the associated laboratory recovery is 90-110% (with a coefficient of variation of less than or equal to 15%), (2) the associated field recovery is 70-120%, and (3) the field recovery samples have experienced the same environmental conditions for the same duration as the field monitoring samples.) Dsually the worker exposure monitoring data will have been corrected based on field recovery unless the field recovery value was 90% or more. Class B data are those data for which (1) the associated laboratory recovery is 80-110% (coefficient of variation of 25% or less) and (2) the field recovery data are present and in the range of 50-120%. Class C data are those data for which (1) the associated laboratory recoveries are 70-120% (coefficient of variation of 33% or less) and (2) either the associated field recovery data are present and in the range of 30-120% or field recovery data are absent and the storage stability data are in the range of 50-120%. Class D data are those data for which (1) laboratory recoveries are 60-120% (coefficient of variation of 33% or less) and (2) the field recovery or storage stability data are either present or missing. Data not meeting the criteria for classes A through D (e.g., if laboratory recovery data were not reported) are assigned to class E. 20.4.4 GUIDANCE FOR USE OF PHED The D.S. EPA has published guidance for proper and appropriate use of PHED (D.S. EPA, 1995b). It is first of all necessary for exposure assessors to accurately define the exposure scenario based on the pesticide product label. Second, it is important for the exposure assessor to subset the data in a consistent and logical way. To achieve this end, the D.S. EPA guidance document (D.S. EPA, 1995b) has indicated the key sub setting parameters for each of the 4 different data files. For example, mixer/loader data would typically be subset based on formulation type (e.g., wettable powder), packaging type (e.g., watersoluble packets), mixing procedures (e.g., open versus closed), and in some cases based on a specific range of total lb a.i. mixed. Applicator data would typically be subset by formulation type (e.g., granulars or all liquids), application method, and cab type (e.g., "open" versus "closed"). In essence, the selected subset should mimic as much as possible the actual use scenario 1Field recovery samples that are spiked with active ingredient and then placed immediately in a cooler are not field recovery samples; rather they measure storage stability.
20.4 Pesticide Handlers Exposure Database (PHED) intended for analysis. The acceptability of a data subset from PHED should be determined by 6 factors: sample size, data quality, duration of sampling period, key body regions, clothing scenarios, and likelihood of incidental contact. Although Subdivision V specifies that a minimum of 15 replicates be present in an exposure data set, the minimum acceptable number should be determined on a case-by-case basis. Sample size requirements are affected by the purpose of the assessment, the quality of the data, and the anticipated variability in exposure data for the worker exposure scenario. The quality of the data, which is assigned separately to airborne, hand, and other dermal exposures for each worker replicate would normally be limited to data quality grades A and B only (see above), although exceptions may be made for some use scenarios in which the data subset is expanded to include C grade data to obtain a data set of adequate size. The sampling period for each worker replicate corresponds to the duration of a given work activity (e.g., mixing/loading or application). Exposure data should be collected over a period of time that is representative of typical work practices. For some work activities, typical durations may be short (e.g., 20 min for mixing/loading) or represent most of the work day (e.g., 4-6 h for application). Data sets from PHED involving longer durations of time are preferred to those associated with short sampling periods. Key body regions relevant to the work activity must be represented in the data set. For example, an exposure data set for mixing/loading that contains none or just a few replicates of data for hand exposure would not be considered a valid data set because hand exposures typically make up the majority of total exposures for mixing/loading activities. The clothing scenario for which the data set is obtained from PHED as reflected in the dosimeter location (e.g., outside of clothing, under normal clothing, or under normal and protective clothing) should either be consistent with pesticide label requirements, or the data (e.g., for total deposition on the skin) should be adjusted to reflect the anticipated clothing scenario using standard clothing protection factors. One or a few replicates of data that produce a body part-specific exposure that is dramatically higher than the central tendency value for that body part may indicate that incidental contact has occurred (e.g., a spill on clothing). It requires careful judgment to determine whether specific worker replicates be deleted from the data subset before analysis. 20.4.5 SUBSETTING STRATEGY AND USEOFPHED
497
ENTER USER DATA OR
USE SYSTEM DATA FILES: ~ Applicator ~ MixerlLoader ~ MixerlLoaderlApplicator ~ Flagger
Select Exposure Algorithm: ~ Inhalation ~ Dermal ~ Inhalation and Dermal
Select Exposure Scenario: ~ Exposed Skin ~ Normal Work Clothing ~ Protective Clothing
Statistical Analysis: » Univariate )- Correlation ~ Regression
1-----+1 SUMMARY EXPOSURE REPORT
Figure 20.2 PHED system overview.
wand)]. The user indicates the way in which the exposure values should be normalized (by time, total lb. a.i., lb a.i./time, or lb a.i./acre/time). Typically, the user selects normalization by totallb a.i. to obtain normalized exposure values from PHED. The user can select the exposure pathway (inhalation, dermal, or both) and can indicate the clothing scenario that can be selected (e.g., total deposition on bare skin, normal work clothing, or protective clothing). Once the type of data normalization is selected, PHED calculates summary statistics as follows: (1) PHED selects dosimeter data of the appropriate type and location (i.e., outside of clothing or on bare skin, under normal clothing, or under normal and protective clothing); (2) PHED assigns a value of 1/2 the quantification limit to non-detect exposure data; (3) PHED performs a distributional fit test on exposure data in the subset for each body part and for total exposure; (4) PHED calculates key central tendency statistics (mean, median, and geometric mean) adjusted from dosimeter area to standard adult body part surface areas (total ng collected in the case of airborne data); and (5) PHED calculates measures of variability in the data. Thus, the standard output report from this module contains the file name, subset name, normalized exposures for 11 body areas, the type of distribution for each body area, central tendency statistics, the coefficient of variation, the best fit total exposure by the selected pathway(s), and the 95% confidence intervals for total exposure. 20.4.6 EXAMPLES OF PHED RETRIEVALS
Most of the practical use of PHED takes place in the Data Analysis Module. The steps involved in using the Data Analysis Module are shown in Fig. 20.2. The user selects which major data file will be used [Applicator, MixerILoader, MixerILoader/ Applicator (i.e., for combined work activities), or Flagger]. A data subset is formed by the user by specifying various conditions [e.g., open or closed mixing, formulation type, packaging type, open or closed cab application, or application method (e.g., broadcast, groundboom, airblast, or low-pressure hand
An example of a summary dermal exposure statistics output report from PHED is provided in Table 20.1 for the case of open mixing/loading of liquid formulations using standard work clothing (i.e., long pants, long-sleeved shirt, and protective gloves) when the PHED data are subsetted for only those worker records associated with adequate quality assurance per U.S. EPA guidance (V.S. EPA, 1995b), specifically PHED data quality grades A and B. The high degree of variability in worker
498
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Table 20.1 PHED Summary Statistics for Dermal Exposure: Open MixinglLoading of Liquid Formulations for the Scenario Long Pants, Long-Sleeved Shirt, Gloves Normalized exposure statistics (flg/lb a.i. mixed) Patch
Distribution
Arithmetic
Coefficient
Geometric
No. of
location
type
Median
mean
of variation
mean
observations
Head
Other
2.99
128.9568
493.8357
4.0992
121
Neck front
Lognormal
1.695
23.2318
360.9199
1.74
103
Neck back
Lognormal
0.341
15.7106
381.7060
0.5427
109
Upper arms
Other
0.582
157.6735
903.2036
1.4925
90
Chest
Other
3.905
19.2219
262.7404
3.4337
89
Back
Other
0.8875
11.009
221.7177
1.8891
88
Forearms
Other
0.6655
Thighs
Lognormal
3.82
Lower legs
Other
Feet
Lognormal
Hands
Lognormal
3.5883
4.4266
211.9821
0.8927
84
16.8134
196.8466
4.0237
71
0.952
38.271
819.5203
1.1162
81
5.371
346.998
180.1404
19.5296
25
316.3227
3.5782
80
34.7596
Note: Best fit dermal exposure is 39.3962 flglIb a.i. handled (35.818 flg/lb a.i. handled excluding hands); 95% confidence interval on the mean is - 12,060.5932 to 13,654.7376; total number of records = 137. Data file: mixerlIoader.
exposures across the multiple studies (roughly 20) in the specific subset represented by the data shown in Table 20.1 is reflected in the high coefficient of variation for the body partspecific exposures and the wide confidence interval for total best fit exposure, as noted in the example in Table 20.1. The variability may decrease somewhat for other exposure scenarios when the data subset represents a fewer number of separate studies, as in the case of dermal exposures associated with open cab groundboom application (derived from 6 separate studies), for which a lO-fold narrower confidence interval is associated with worker exposures. As seen in Table 20.1, the exposure data in PHED are analyzed by separate body regions (e.g., hands, forearms, chest) and exposure statistics provided in PHED for each body region include the median, arithmetic mean, geometric mean, and the coefficient of variation. A best fit total dermal exposure is also presented, which is the sum of the most appropriate measure of central tendency for each body part. 2 Thus, the central tendency estimate of exposure is the sum of the central tendency values for the individual body part exposures derived from all appropriate data for a given worker exposure scenario within the data base (i.e., based on separate exposure data for head, neck, upper arms, chest, back, forearms, thighs, lower legs, hands, and feet, with each body part exposure defined by a distribution and central tendency value). One can also obtain statistical output reports that include univariate, regression, or correlation statistics. The univariate analysis provides percentile estimates of exposure. The exposure estimates for dosimeters under personal clothing, under protective clothing, under both protective and personal clothing, or outside of all clothing, if selected, are 2PHED assigns the following as the most appropriate measure of central tendency: geometric mean if the exposure data are lognormally distributed, arithmetic mean if the exposure data are normally distributed, and median if the distribution of exposure data is neither normal or lognormal (Lunchick et aI., 1994; U.S. EPA, 1995a, b).
presented based on the available data for those dosimeter locations. No attempt is made in PHED to adjust outside dosimeter data using clothing protection factors or other exposure mitigation defaults to represent "inside clothing" exposures. Thus, if no "inside" dosimeter data are available for covered body parts for a given exposure scenario, no data will appear in the summary output report for the corresponding clothing scenario. Table 20.2 shows the best-fit normalized exposures for a variety of use scenarios and clothing scenarios to illustrate the different central tendency values that occur based on application method, formulation type, and extent of protective clothing. The D.S. EPA typically uses their PHED surrogate exposure tables (D.S. EPA, 1998) for development of worker exposure estimates for the Reregistration Eligibility Decision documents. 20.4.7 LIMITATIONS OF PHED
Because PHED was developed in the late 1980s and early 1990s, some of the studies in PHED were not strictly conducted according to good laboratory practices (GLP), and some of the more recent studies are not in the data base. PHED also suffers from a high percentage of nondetects, and the minimum detection limits for the studies range over several orders of magnitude; thus, for some subsets, the exposures are mainly driven by the limits of quantification. Because of the wide differences in study designs and detection limits, extremely high variability occurs in the aggregated data in the exposure data subsets from multiple studies. Thus, exposure statistics other than the central tendency may have limited meaning (van Hemmen, 1992). Many replicates are of such short duration or involve handling of such small amounts of material that they are not generally applicable to exposures associated with more typical worker task durations or more typical amounts of active ingredient handled per day, respectively. Because chemical-specific properties
20.5 POEM
499
Table 20.2 Example PHED Surrogate Exposure Data for Various Use Scenarios Normalized dermal exposure for clothing scenario (mgllb a.i. handled)
Normalized inhalation
Total deposition
Normal clothing,
Normal clothing,
Coveralls, normal
exposure
(no clothing)
no gloves
gloves
clothing, gloves
(mg/lb a.i. handled)
0.0034
0.0017
Use scenario
Formulation type
Open mixlload
Granular
0.032
0.0084
0.0069
Wettable powder
6.7
3.7
0.17
0.043
All liquids
2.2
0.36
0.24
0.0045
All liquids
0.046
0.014
0.014
0.00074
All liquids
0.010
0.0050
0.0051
0.000043
Granular
0.0021
0.0021
0.0020
0.00022
All liquids
1.9
1.3
0.39
0.0039
All liquids
0.0050
0.0050
0.0022
0.000068
9.3
0.062
0.012
0.00035
Airblast application (open cab) Groundboom application (open cab) Groundboom application (closed cab) Broadcast spreader (closed cab) Right-of-way application Aerial fixed wing application Open pour/belly grinder
Granular
210
10
(mixlload/application) Flagger
All liquids
0.053
0.011
Source: Adapted from U.S. EPA (1998).
(e.g., vapor pressure) may affect penetration of worker clothing, the protection provided by personal protective equipment (PPE) or engineering controls may be underestimated by the use of PHED data. Despite these limitations, PHED currently represents the major data base tool used in developing estimates of mixerlloader and applicator exposures in North America.
20.4.8 PHED SUBMISSIONS TO REGULATORY AGENCIES There are a number of key items to include in a submission of PHED data to the U.S. EPA, the California EPA, or Health Canada's Pest Management Regulatory Agency (PMRA). These include PHED output reports, specifications of subsetting parameters and conditions, completed PHED data reporting forms (available from the U.S. EPA), and ancillary data printouts that facilitate review. In addition, it is recommended that the registrant provide a worker exposure and risk assessment report that clearly states the assumptions used (e.g., use rate and area treated/day), identifies the no observed effect levels (NOELs) and other key toxicity benchmarks, describes the surrogate PHED data sets, presents the algorithms used to estimate exposures, and provides calculation and discussion of the margins of safety (margins of exposure) and excess lifetime cancer risk (if relevant). For PMRA submissions, it is recommended that a summary in the PMRA format be provided to supplement the risk assessment and to expedite review (Personal communication, 2000).
20.5 POEM 20.5.1 OVERVIEW AND HISTORY OF POEM In May 1985, the Joint Medical Panel of the Scientific Subcommittee on Pesticides and the British Agrochemical Association Toxicology Committee decided to review the available pesticide worker exposure data to determine the extent to which they would support a generic exposure model (Martin, 1986). The result of the effort that followed this decision is POEM, which was created in the United Kingdom. The European community first had an opportunity to turn to POEM in the late 1980s to estimate the levels of exposure likely to be incurred by operators (mixerlloader/applicators) applying pesticides in the United Kingdom under U.K. conditions. This model relies on default values for key parameters that are selected from lists by the user (see SC8001, 1986). It is notable that POEM was available to the risk assessment community in Europe a number of years before PHED was released to the public in the United States. Although France has adopted it (in modified form for operators not wearing protective clothing), the exposure data in POEM may not be universally relevant to application conditions in every European country. A complete discussion of POEM has been provided by Martin (1986, 1990) and Hamey (1992) provides the user's guide for the model. POEM is based on limited generic exposure monitoring data extended to cover a variety of formulation types and use scenarios. Variables that describe the use con-
500
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ditions are utilized in POEM to predict daily exposure. These variables include formulation type, formulation concentration, container size, contamination potential associated with pouring, number of hectares treated, and the total volume of material applied (Lunchick et aI., 1994). As with PHED, defining the use conditions properly is crucial; familiarity of the POEM user with pesticide handling practices and exposure study protocols is recommended for proper interpretation of the exposure estimates and the limitations under which the exposure estimates are used (Lunchick et al., 1994). The application methods addressed in POEM include vehicle-drawn hydraulic spray equipment and controlled droplet (CDA) equipment, low-level hand-held outdoor sprayers, low-level and high-level hand-held CDA sprayers, and very low-volume, low-volume, and highvolume orchard sprayers (Lunchick et aI., 1994). The model contains algorithms and structured calculation procedures in spreadsheet format (Hamey, 1992). Product-specific considerations (e.g., application rate) and technique-related considerations can be incorporated in the calculations. The impact of protective clothing and formulation differences are also addressed by POEM. The point estimate of exposure obtained from POEM is based on the upper bound value of the range that contains the 75th percentile of the exposure data evaluated by the Working Party on Pesticide Operator Exposure. POEM has been approved by the Scientific Subcommittee and the Advisory Committee on Pesticides and is used for regulatory purposes by the Pesticides Safety Directorate in the United Kingdom. 3 20.5.2 ESTIMATION OF OPERATOR EXPOSURES USING POEM
The estimation of operator exposures using POEM is divided into two parts: (1) estimation of exposures associated with handling the concentrated formulation (e.g., during mixing and loading procedures) and (2) estimation of exposures associated with actual application of the diluted formulation. For exposures during the mixing/loading operation, POEM assumes that only the hands are contaminated; generic data in POEM used for this activity are based on exposures during container pouring tests rather than on exposures during actual spray tank loading. These data indicate the dependence of operator exposure during mixing/loading on the volume of the pesticide formulation container and the neck aperture width of the container. Thus, for liquid formulations, hand contamination values for mixing/loading are 0.01 ml of formulation per operation for a l-liter bottle, 0.20 ml/operation for a 5-liter bottle, and 0.50 ml/operation for a 10- or 20-liter container. Bottles with wide necks are assigned hand exposure values ranging from 0.01 to 0.1 ml/operation, depending on the neck width. POEM assigns the following exposure values for mixing/loading for 3An alternative procedure has been accepted by the UK Pesticide Safety Directorate-namely the "absorption rate model" developed and proposed by ICI/Zeneca (Chester, 1988). This approach uses biological monitoring data in a generic fashion to predict the dermal absorbed dose, using percutaneous absorption data.
other formulation/packaging types: 0.01 g/operation for small packs (approximately 100 g) of wettable powders, waterdispersible granules, or tablets; 0.1 g/operation for larger packs of solid formulations (approximately 1 kg), and no exposure (i.e., complete protection) for solid formulations packaged in water-soluble packets. The transfer of formulation from protective gloves to hands (including glove penetration and incidental contact of the contaminated gloves with other parts of the body) is assigned a value of 10% of potential contamination for emulsifiable concentrates (ECs) and other liquid formulations containing organic solvents, 5% for aqueous-based liquid formulations, and 1% for solid formulations. For the application of pesticide formulations, both dermal and inhalation exposures are considered in POEM. A number of factors are considered in calculation of dermal exposures occurring during pesticide application-these include, for example, the work rate (i.e., area treated per day), the volume of pesticide contamination on the worker (mllh), the distribution of pesticide contamination on the worker (percentage of total contamination on hands, trunks, and legs), the anticipated penetration of clothing (assigned separately for hands, trunks, and legs), the duration of exposure to the spray (h), and the percutaneous absorption of the active ingredient. It is assumed that vehicle-mounted hydraulic-boom applications are made to 50 ha/day.4 For air-assisted applications to orchards, the area treated is assumed to be 30 ha/day. Hand-held applications are associated with a maximum work rate of 1 ha/day or a volumetric application rate of 4000 liters/day. The duration of exposure to spray, corresponding to duration of the application work task is typically assumed to be 6 h/day, at least for vehicle-mounted applications and hand-held applications. Where either in vitro or in vivo percutaneous absorption data exist, the dermal absorption can be classified as either low (1 %), medium (10%), or high (100%). In the absence of such data, the usual default assumed in POEM is 10%. Potential inhalation exposures to formulations ranging from 0.005 to 0.05 mllh are assigned in POEM depending on the type of application method, protective measures (e.g., use of cab), and application volume. POEM combines the dermal absorbed dose and the inhalation exposure to obtain a total absorbed dose [mgikg body weight (bw)/day], based on a 60-kg adult body weight. The total absorbed dose is then compared to an acceptable operator exposure level (AOEL). The AOEL is the exposure level that is unlikely to be associated with adverse health effects, based on a compound-specific toxicological benchmark such as a NOEL. 20.5.3 A COMPARATIVE CASE STUDY-POEM VERSUS PHED
Lunchick et al. (1994) reported the results of a comparative case study in which worker exposures to a synthetic pyrethroid were obtained using both POEM and PHED. Both models were used for estimation of airblast applicator exposures because by itself 4This is reasonable for grain crops, but a lower value for the area treated per day may be justified for other crops (Hamey, 1992).
20.6 EUROPOEM
PHED version 1.0 did not contain the requisite 15 replicates of grade A and B data for each body part nonnally required for regulatory purposes by the U.S. EPA and Health Canada (since then, PHED version 1.1 has been released, which does contain more than 15 replicates for each body part for dennal exposures for this worker exposure scenario). Both models were fun under comparable use conditions, as shown in Table 20.3. Predicted worker exposures were generated from POEM for the low-volume and high-volume orchard applications. The combined mixer/loader/applicator daily exposures for a pesticide handler wearing long pants and a long-sleeved shirt (and also protective gloves during the mixinglloading operation) were 1.0 mg/kg/day for the low-volume spray and 0.49 mg/kg/day for the high-volume spray. Table 20.4 presents the POEM printout for the low-volume spray. The separate mixer/loader best-fit estimate based on PHED data was 0.018 mg/lb a.i. for the same clothing scenario. Based on best-fit exposure data for application equipment similar to that relevant for the POEM model run (airblast application of liquids, excluding grapes as a crop), the dermal exposure for an applicator wearing long pants and a long-sleeved shirt was 1.0 mgllb a.i. Based on both POEM and PHED, the inhalation exposures were negligible compared to the dermal exposure (i.e., less than 1% of total exposure). Rounding the combined mixerlloader and applicator best-fit normalized exposure (0.018 mgllb a.i. + 1.0 mg/lb a.i. = 1.018 mg/lb a.i.) to 1 mgllb a.i., a total daily exposure can be calculated for a 60-kg worker applying 0.14 Ib a.i./acre to 70 acres/day (per Table 20.3) as follows: Daily exposure (mg/kg/day) = (1 mgllb a.i.) x(0.141b a.i./acre) x(70 acres/day)/(60 kg) = 0.16 mg/kg/day
Although the PHED estimate is 7.5-fo1d lower than the POEM estimate for the low-volume spray and 3-fold lower than the POEM estimate for the high-volume spray, these differences (1) are less than the variability typically encountered among replicates in individual worker exposure studies and (2) may be
due in part to the fact that the PHED estimate represents a central tendency estimate and the POEM estimate represents the 75th percentile of exposure. When the 75th percentile nonnalized data are retrieved using the univariate statistics feature of PHED for mixing/loading and airblast application, a calculated daily exposure of 1.2 mg/kg/day is obtained, which is virtually identical to the results from POEM. Although worker exposure assessments submitted to the U.S. EPA typically focus on the best-fit exposure, when the same exposure statistics (i.e., the 75th percentile) are compared for selected exposure scenarios POEM and PHED may yield similar results for some use scenarios.
20.6 EUROPOEM 20.6.1 OVERVIEW EUROPOEM is a data base/model that has been developed through an expert group for the European Commission as a harmonized model for prediction of operator exposures to plant protection products in Europe (AIR, 1996). This effort was a concerted action under the AIR-specific program (agriculture and agro-industry, including fisheries) of the European Community'S Third Framework Programme for Research and Technological Development. Approval of a plant protection product follows demonstration that the operator exposure associated with handling and use in the field based on EUROPOEM does not exceed the AOEL. In developing EUROPOEM, the EU considered 3 worker exposure models developed by its member states; these included the POEM developed in the United Kingdom (Martin, 1986), the German model (FBRC, 1993), and the Dutch model (van Goldstein Brouwers et aI., 1996). The EUROPOEM expert group has requested studies from academia, industry, and government; these studies address a variety of uses and worker exposure scenarios and are generally not publicly available. Review of these studies has been based on a set of acceptance criteria that consider the adequacy
Table 20.3 Summary of Use Information for Comparative Case Study Use pattern
501
Input parameter values for surrogate exposure models POEM
PHED
Formulation
EC
EC, solution, and suspension
Concentration
380 mg/ml
3.21b a.i.lgallon (U.S.)
Container size
41iters
I gallon (U.S.)
Container neck design
Narrow neck
Not applicable
Application dose
0.42 liters productlha
0.14 Ib a.i.!acre
Spray rate = high volume
2850 liters/ha
300 gallons (U.S.)/acre
Spray rate = low volume
300 liters/ha
30 gallons (U.S.)lacre
Work rate
28 ha/day
70 acres/day
Duration of exposure
6h
Not applicable
Source: Adapted from Lunchick et al. (1994).
502
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Table 20.4 Summary of POEM Low-Volume Orchard Sprayer Exposure A.
B.
C.
D.
E.
Product data
1.
Product name
2a.
Active ingredient
2b.
Concentration
380 mg/ml
3.
Formulation type
EC
4a.
Main solvent
Aromatic
4b.
Concentration of solvent
NA
5.
Maximum in-use a.i. concentration
0.532 mg/ml
Syn Pyreth
Exposure during mixing and loading la.
Container size
4liters
lb.
Hand contamination/operation
0.2ml
2.
Application dose
0.42 liter product/ha
3.
Work rate
28 ha/day
4.
Number of operations
3/day
5.
Hand contamination
0.6 mllday
6.
Protective clothing
Gloves
7.
Transmission to skin
10%
8.
Dermal exposure to formulation
0.06 mllday
9.
Concentration of a.i.
380 mg/ml
10.
Dermal exposure to a.i.lperson
22.800 mg/day
11.
Dermal exposure to a.i.lkg bw
0.38 mglkg bw/day
Exposure during spray application upward air-blast, no cab, low volume
1.
Application technique
2.
Application volume
300 literslha
3.
Volume of surface contamination
50 mllh
4.
Distribution
Hands
Trunk
Legs
10%
65%
25% Permeable
5.
Clothing
None
Permeable
6.
Penetration
100%
15%
20%
7.
Dermal exposure
5mllh
4.875 mllh
2.5 mllh
8.
Duration of exposure
6hr
9.
Total dermal exposure to spray
74.25 mllday
10.
Concentration of a.i.
0.532 mg/ml
11.
Dermal exposure to a.i.lperson
39.501 mg/day
12.
Dermal exposure to a.i.lkg bw
0.658 mglkg bw/day
Inhaled exposure during spray application 1.
Inhalation exposure
0.02 mllh
2.
Duration of exposure
6h
3.
Concentration of a.i.
0.532 mg/ml
4.
Inhalation exposure to a.i.
0.064 mg/day
5.
Percent absorbed
100%
6.
Absorbed dose
0.001 mglkg bw/day
Predicted exposure 1.
Inhalation exposure
0.001 mglkg bw/day
2.
Dermal exposure
1.038 mglkg bw/day
3.
Total potential exposure
1.039 mglkg/day
Source: Lunchick et al. (1994).
20.6 EUROPOEM
of experimental design, extent of documentation, number of replicates, and the adequacy of the quality assurance procedures. Although it was recognized that gaps would remain in the initial harmonized data base, data needs for future studies would be delineated as a result, which would allow expansion of the data base. The project initially proceeded with the option of adapting the North American PHED data base platform for use with the European data sets. Thus, the types of data fields in EUROPOEM are essentially identical to the data fields in PHED. The EUROPOEM data base compilation to date has been developed in the Microsoft Excel spreadsheet format, similar to the form of POEM, which allows flexibility in manipulation and statistical handling of the data. Each study in EUROPOEM has been summarized in a standardized generic format. Separate data bases have been prepared for each specific operator exposure scenario; these data bases contain the data points retrieved from various studies that contribute to modeling the operator exposure for standard scenarios. In part due to the high variability in the surrogate exposure data between different studies for a given use scenario, a single statistical value between the 75th percentile and the rounded maximum exposure is selected as the combined surrogate exposure value, depending in part on the quantity of data available in EUROPOEM for the operator scenario of interest. Typical formats for the normalized exposure in EUROPOEM are mg/kg of active substance (a.s.) handled, and ml of spray contacted per hour. The exposures are obtained based on application rate, volumetric application rate, concentration of active substance in the formulation, spray dilution, and total amount of plant protection product or active substance handled. Protection factors (i.e., exposure reduction coefficients) based on the use of PPE are selected for use in the tier II of the model. Typically, reduction coefficients used in EUROPOEM (e.g., in the United Kingdom) are in the range of 0.02-0.2 for garments (i.e., 5to 50-fold reduction in exposure compared to uncovered skin) and 0.01-0.1 for gloves (10- to lOO-fold reduction in exposure compared to bare hands), depending in part on the nature of the garments and type of gloves (AIR, 1996). Other European countries using EUROPOEM may assign different reduction coefficients. The EUROPOEM expert group has recommended focusing on the 75th percentile of the exposure data when a large number of data points (e.g., 50-100) are available from at least 10 studies that represent a wide range of active substances, uses, and climatic conditions. The reason for selecting the 75th percentile for large data sets is that the arithmetic mean exposure (most meaningful for comparison with chronic toxicity benchmarks) is close to the 75th percentile for lognormally distributed data, which is often the manner in which operator exposures are distributed. The maximum exposure value may be selected when only 15-20 data points are available from several studies (AIR, 1996). For intermediate cases (i.e., between 20 and 50 data points), the 90th percentile exposure may be selected. This yields a distinctly more conservative approach than the central tendency statistics currently used from PHED in North America, which often favor the geometric mean. How-
503
ever, the toxicological profile of the active ingredient can affect the selection of the exposure statistic to use as the surrogate exposure value. For example, if repeated exposure is required to cause the effect of concern, and anticipated use patterns suggest that exposure will occur only infrequently, then a central tendency value may be more appropriate. Furthermore, if an active ingredient has a fairly benign toxicological profile (e.g., low mammalian toxicity and/or reversible health effects), then a central tendency exposure value may also be appropriate, particularly if the exposure data set consists of a large number of studies for which the exposure ranges and variability are similar across the different studies. A key aspect of the EUROPOEM approach is the need to ensure sufficient conservatism in predictions of exposure across the highly variable conditions (e.g., in equipment used, climatic conditions, and work habits) in EU member states. Although point value exposure estimates are typically obtained with EUROPOEM, recent efforts indicate the usefulness of stochastic (e.g., Monte Carlo) analyses of relevant distributional data for worker exposures (Ross and Dong, 1996). 20.6.2 CASE STUDY
To illustrate the use of EUROPOEM, a case study involving estimation of exposures experienced by air-assisted broadcast sprayer operators is provided. The usable airblast sprayer data consists of 8 studies with a combined total of 51 replicates, involving use of tractors without cabs. These studies were conducted between 1984 and 1990 in either Germany or the United Kingdom. The field parameters for these studies are shown in Table 20.5. A significant number of these studies were conducted on small areas, with 21 of the replicates representing treated areas of less than 2 ha. As is typical with many of these studies, not all of the body parts have been monitored for exposure in some of the studies. Head exposures, which represent between 7 and 18% of total body exposures for this worker scenario, were not measured in 4 of the studies. The geometric mean (GM), geometric standard deviation (GSD), and percentile exposure data for the complete data set are shown in Table 20.6. Exposure data for unprotected hands were measured in 6 studies for a total of 35 data points. The hand exposure data cover a range of 3 orders of magnitude, with a 75th percentile value of 11 mg/kg a.s. handled. Data for body exposures (excluding hands) were measured in 6 studies for a total of 47 data points. The body exposure data span a range of approximately 2 orders of magnitude, with a rounded 75th percentile value of 63 mg/kg a.s. handled. When the data are subset to include hands with the whole body exposures, a rounded 75th percentile of 76 mg/kg a.s. handled is obtained. Inhalation exposure data for airblast operators were measured in 7 studies, for a total of 44 data points. The inhalation exposure data range across slightly more than 2 orders of magnitude, with a rounded 75th percentile value of 0.03 mg/kg a.s. handled. These normalized exposure values are then used to calculate exposures to the specific active substance of interest in mg/kg/day, based on the formulation use rate and percent a.s.
504
CHAPTER 20
Occupational Exposure Data BaseslModels
Table 20.5 EUROPOEM Field Parameter Data for Air-Assisted Broadcast Sprayer Operators
Statistic
Volumetric
Total a.s.
Total area
Ground
Dermal
Use rate
use rate
applied
treated
Final spray
No. tanks
speed
sampling
height
(kg a.s./ha)
(liters/ha)
(kg)
(ha)
conc.
applied
(km/h)
time (min)
(m)
Crop
Minimum
0.02
38
0.01
0.9
0.04
2
4.5
20
25th percentile
0.1
103
0.4
1.6
Ll
2
6.1
47
2.5
50th percentile
0.4
439
0.8
3.0
1.8
2
6.4
60
2.5
75th percentile
0.5
500
1.9
3.0
3
5
6.6
60
2.5
90th percentile
1.3
500
3.9
4.0
3
5
7
123
2.5
1.8
1000
Maximum No. of
51
5.6
51
51
14
3
7
8
311
51
51
19
51
51
2.5
3.5 47
observations
20.7 REENTRY EXPOSURE DATABASE 20.7.1 GENERAL PURPOSE Unlike in the case of mixerlloader/applicator exposures where PHED has been available to obtain normalized exposures to pesticides for specific use scenarios (U.S. EPA, 1995a), a completed publicly available data base on worker reentry exposures does not exist at this time. The general concepts of such a data base and how the information in it would be applied have been described by others (Honeycutt, 1985; Nigg et aI., 1984; Zweig et aI., 1985). The ARTF, which consists of member companies who manufacture and/or distribute pesticides, is developing a worker reentry data base for use by its member companies. As of the writing of this chapter, a portion of the worker reentry data base is completed, and the remainder of the data base is under construction. Such a data base could contain generic transfer coefficients that are representative of specific crop types and worker activities, based on actual field studies sponsored by the ARTF
or purchased by the ARTF. These generic transfer coefficients could then be used to estimate worker reentry exposures. In theory, such a data base could allow sub setting of the data by key variables that may affect the level of exposure. A worker reentry exposure data base could contain data on worker dermal and inhalation exposures, dislodgeable foliar residues, site location, meteorological conditions, and other ancillary information (e.g., formulation type, method of application, and reentry times and conditions). Such a data base could provide output reports on transfer coefficients on a whole-body and a body part-specific basis. The latter would provide guidance for exposure mitigation methods once the body part-specific transfer coefficients are used to calculate body part-specific exposures for the worker scenario of interest for a specific formulation of interest. The EUROPOEM H project is the European effort to develop a worker reentry exposure data base for use under 91/414 EC legislation. This will at least initially be a relatively small data base, and the hope is that a U.S. reentry exposure data base can be used to supplement data in EUROPOEM H.
Table 20.6 Normalized Exposure Data from EUROPOEM for Air-Assisted Broadcast Sprayer Operators for All Spray Volumes Normalized Exposure type
Geometric
exposure
No. of
Geometric
standard
units
observations
mean
deviation
Unprotected hands
mg/kg a.s.
35
Body (without hands)
mg/kg a.s.
47
27 33
Total dermal
mg/kg a.s.
47
Inhalation
mg/kg a.s.
44
3.7
0.018 1.6
Percentile 10th
25th
50th
75th
90th
5.4
0.5
11
25
3.3
6.9
10
26
63
110
12
30
76
116
2.2
3.2
7.8
2.8
0.007
0.007
3.8
0.02
5.2
0.07 16
Unprotected hands
ml spray/h
35
6.4
0.3
0.6
Body (without hands)
ml spray/h
47
14
4.6
1.6
5.7
12
34
93
Total dermal
ml spray/h
47
18
4.4
2.4
8.2
17
35
116
Inhalation
ml spray/h
44
3.8
0.0017
0.005
0.01
1.2
0.03
0.01
0.03
0.05
20.8 Summary
20.7.2 POSSIBLE DATA SUB SETTING CRITERIA
Examples of the basic data subsetting criteria that could be included in a reentry exposure data base are crop type, worker activity, growth stage of the crop (related to crop height and degree of foliage), geographic region, and level of clothing and protective equipment. Secondary subsetting criteria could include physical state as applied, application method used, application rate, season of the year, and meteorological conditions (typical versus atypical, wind speed ranges, precipitation, temperature, and presence or absence of dew during DFR sample collection). Transfer coefficient data could be subsetted from a reentry data base based on the anticipated label-specified clothing or the appropriate transfer coefficient data for specific body areas could be adjusted by a generic clothing penetration factor. Thus, it would be desirable to adjust or subset the available transfer coefficient data based on the relevant clothing scenario for the worker. 20.7.3 STATISTICAL CONSIDERATIONS
Worker reentry exposures result from contact with the foliage of treated crops and should be a function of DFR levels. Worker reentry exposures for a given day are a function of the contact with the foliage over the entire work day and at a number of locations in the field. If comparison of exposures to longterm toxicological benchmarks is appropriate, then amortizing exposures over time may bring workers into contact with different levels of DFRs in the same field or in different fields. When this sort of environmental averaging occurs, the resulting exposure measurements often follow a normal distribution. However, there may be cases where the empirical distribution of the exposure values resemble a more complex distribution form. To the extent that DFR values exhibit spatial heterogeneity across a field, heterogeneity of worker exposures may occur. Even if DFR values are spatially uniform across a treated field, personal work habits or personal characteristics may result in statistically significant differences in exposure between workers. Climatic or meteorological factors may also play a contributing role in heterogeneity of exposures across studies or across reentry dates. Insights concerning the distributional form of the worker reentry exposure data can provide guidance for selection of the most appropriate tools for screening and analyzing these data and may help guide the development of statistical models that can provide ways of investigating the effect of particular factors on worker reentry exposures. The actual choice of statistical tools would depend, in part, on the distributional form of the data. Statistical tools can be used, for example, to investigate data grouping and comparison issues, such as whether there are regional and seasonal differences in transfer coefficients or whether there are statistically significant differences in transfer coefficients between studies or between reentry times. The extent to which statistically significant differences are verified determines the nature and extent
505
of data grouping that can be made in a scientifically justifiable way. For mitigation purposes, one can also perform statistical comparisons of patterns of exposure (i.e., distribution of total exposure or transfer coefficients across specific body regions, such as hands, forearms, or legs, for the same clothing and PPE scenario) to see if there are differences from crop to crop, from one growth stage to another, and from one region to another, for example.
20.8 SUMMARY The development of PHED in North America, POEM in the United Kingdom, and EUROPOEM within the European Union provides powerful predictive tools for estimating worker exposures for specific pesticide use scenarios. Use of these widely available tools will continue to provide guidance to regulatory agencies and the agrochemical industry in regulation, product development, and product stewardship. Detailed knowledge of the workings of POEM and PHED, appropriate selection of pesticide use information, and familiarity with the features and limitations of pesticide worker exposure studies are critical to the effective use of both models. When comparing the same statistical measure of exposure (i.e., the 75th percentile) is compared, POEM and PHED may yield similar results for some exposure scenarios. The eventual development of a worker reentry exposure data base may provide an important predictive tool for selected worker activity/crop combinations, and, thus, fill an important gap for exposure assessors.
ACKNOWLEDGMENTS The authors would like to extend their appreciation to Graham Chester of Syngenta and Pierre-Gerard Pontal of Aventis CropScience for providing guidance, comments, and source materials for the sections on POEM and EUROPOEM. Their assistance was invaluable.
REFERENCES Agriculture and Agro-Industry including Fisheries (AIR). (1996). "The Development, Maintenance, and Dissemination of a European Predictive Operator Exposure Model (EUROPOEM) Database. A EUROPOEM Database and Harmonised Model for Prediction of Operator Exposure to Plant Protection Products." Draft Final Report, December, Publication No. AIR3 CT93-1370, concerted action under the AIR specific programme of the Community'S Third Framework Programme for Research and Technological Development and managed by DGVLFII.3. Chester, G. (1988). Pesticide applicator exposure-Towards a predictive model for the assessment of hazard. Asp. Appl. Bioi. 18,331-343. Curry, P. B., Iyengar, S., Maloney, P. A., and Maroni, M. (1995). "Methods of Pesticide Exposure Assessment." Plenum, New York. Day, E. W. (1991). "Historical Development and Status of the Pesticide Users Exposure Data Base" (A historical summary dated July 19, 1991). DowElanco, Indianapolis. Driver, J. H., and Whitmyre, G. K. (1997). Pesticide regulation and human health: The role of risk assessment. In "Fundamentals of Risk Analysis and Risk Management" (Molak, v., ed.), pp. 143-162. CRC Lewis, New York.
506
CHAPTER 20
Occupational Exposure Data BaseslModels
Ecobichon, D. J. (ed.). (1999). "Occupational Hazards of Pesticide Exposure: Sampling, Monitoring, Measuring." Taylor & Francis, Philadelphia. Federal Biological Research Centre (FBRC). (1993). "Guidelines for the Examination of Plant Protection Products in the Authorization Procedure. Part I, 3rd ed. Labeling of Plant Protection Products-Health Protection. Instructions for the Protection of Operators and Other Persons in the Directions for Use." Braunschweig Federal Biological Research Centre for Agriculture and Forestry, Federal Republic of Gennany. Fong, H. R., and Krieger, R I. (1988). "Estimation of Exposure of Persons in California to Pesticide Products that Contain Dinocap (Karathane) and Estimation of Effectiveness of Exposure Reduction Measures." Publication No. HS-1469, Department of Pesticide Regulation, Worker Health and Safety Branch, Sacramento. Hackathorn, D. R, and Eberhart, D. C. (1985). "Data-Base Proposal for Use in Predicting Mixer-Loader-Applicator Exposure." In "Dennal Exposure Related to Pesticide Use" (R C. Honeycutt, G. Zweig, and N. N. Ragsdale, eds.), ACS Symposium Series, Vo!. 273, pp. 341-355. Am. Chem. Soc., Washington, DC. Hamey, P. Y. (1992). "Predictive Operator Exposure Model (POEM): A User's Guide." MAFF Pesticides Safety Division. Honeycutt, R. C. (1985). Field worker exposure: The usefulness of estimates based on generic data. In "Dennal Exposure Related to Pesticide Use" (R C. Honeycutt, G. Zweig, and N. N. Ragsdale, eds.), ACS Symposium Series, Vo!. 273. pp. 369-375. Am. Chem. Soc., Washington, DC. Honeycutt, RC., Zweig, G., and Ragsdale, N. N. (eds.). (1985). "Dennal Exposure Related to Pesticide Use." ACS Symposium Series, Vo!. 273. Am. Chem. Soc., Washington, DC. Krieger, R, Biewett, c., Edmiston, S., Fong, H., Gibbons, D., Meinders, D., O'Connell, L., Ross, J., Schneider, F., Spencer, J., Thongsinthusak, T. (1990). Gauging pesticide exposure of handlers (mixerlloaders/applicators) and harvesters in California agriculture. Med. Lav. 81(6), 47~79. Lunchick, c., Hamey, P., and Iyengar, S. (1994). The use of the North American (PHED) and United Kingdom (POEM) worker exposure models in pesticide registration. In "Proceedings of the Brighton Crop Protection Conference: Pests and Diseases-1994," Vo!. 3. British Crop Protection Council, Famham, Surrey, U.K. Lundehn, J. R., Westphal, D., Kieczka, H., Krebs, B., Locher-Bolz, S., Maasfeld, w., and Pick, E. D. (1992). "Unifonn Principles for Safeguarding the Health of Applicators of Plant Protection Products: Unifonn Principles for Operator Protection." Braunschweig Federal Biological Research Centre for Agriculture and Forestry, Federal Republic of Gennany. Maddy, K. T, Edmiston, S., and Richmond, D. (1990). Illnesses, injuries, and deaths from pesticide exposures in California, 1949-1988. Rev. Environ. Contam. Toxieol. 114,57-123. Martin, A. D. (1986). "Estimation of Exposure and Absorption of Pesticides by Spray Operators." Scientific Subcommittee on Pesticides and British Agrochemical Association Joint Medical Panel Paper Number PS 4221/SC 8001. Martin, A D. (1990). A predictive model for the assessment of dermal exposure to pesticides. In "Proceedings of a Workshop Entitled 'Prediction of Percutaneous Penetration' " (R C. Scott, R. Guy, and J. Hadgraft, eds.). IBC Technical Services, London. Mehler, L., Thongsintusak, T., and Haskell, D. (1991). "Estimation of Exposure of Persons in California to Pesticide Products That Contain Benomy!." Publication No. HS- 1557, Department of Pesticide Regulation, Worker Health and Safety Branch, Sacramento. National Agricultural Chemicals Association (NACA). (1987). Jan. 2 letter from Dr. John McCarthy, NACA Director of Scientific Affairs to the NACA Exposure Assessment Subcommittee, regarding "Submission of data to the generic mixer/loader/applicator exposure data base." Nigg, H. N., Stamper, J. H., and Queen, R M. (1984). The development and use of a universal model to predict tree crop harvester pesticide exposure. Am. Ind. Hyg. Assoe. 1. 45, 182-186. Nutley, B. P., and Cocker, J. (1993). Biological monitoring of workers occupationally exposed to organophosphorous pesticides. Pestic. Sei. 38, 315-322.
Personal communication (2000). June 30 communication between Gary Whitmyre, risksciences, LLC, and Christine Nonnan, PMRA, Health Canada. Plimmer, J. (ed.). (1982). "Pesticide Residues and Exposure." ACS Symposium Series, Vo!. 182. Am. Chem. Soc., Washington, DC. Rech, c., Bissell, S., and del Valle, M. (1988). "Potential Dennal and Respiratory Exposure to Abamectin during Greenhouse Applications," Publication No. HS-1491, Department of Pesticide Regulation, Worker Health and Safety Branch, Sacramento. Reinert, J. c., and Severn, D. J. (1985). Dennal exposure to pesticides. The Environmental Protection Agency's viewpoint. In "Dennal Exposure Related to Pesticide Use" (R C. Honeycutt, G. Zweig, and N. N. Ragsdale, eds.), ACS Symposium Series, Vo!. 273, pp. 357-368. Am. Chem. Soc., Washington, DC. Ross, J. H., and Dong, M. H. (1996). The use of probabilistic modeling to detennine reentry intervals. Paper presented at the 35th Society of Toxicology Annual Meeting, Poster No. 1299, March 10-14, 1996, Anaheim, California. Saleh, M. A., Blancato, J. N., and Nauman, C. H. (eds.). (1994). "Biomarkers of Human Exposure to Pesticides," ACS Symposium Series, Vo!. 542. Am. Chem. Soc., Washington, DC. SC800l (PS 4221). (1986). "Estimation of Exposure and Absorption of Pesticides by Spray Operators." Scientific Subcommittee on Pesticides and British Agrochemical Association Joint Medical Pane!. Thongsinthusak, T, Biewett, T C., Ross, J., and Krieger, R. I. (1993). "Estimation of Exposure of Persons in California to Pesticide Products That Contain Chlorothaloni!." Publication No. HS-1475, California Department of Pesticide Regulation, Worker Health and Safety Branch (WHSB), Sacramento. U.S. Environmental Protection Agency (U.S. EPA). (1984). "Pesticide Assessment Guidelines. Subdivision K. Exposure: Reentry Protection." Office of Pesticide Programs, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA). (1995a). "Pesticide Handlers Exposure Database." Occupational and Residential Exposure Branch, Office of Pesticide Programs, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA). (1995b). "Pesticide Handlers Exposure Database (PHED) Evaluation Guidance, PHED version 1.1." Occupational and Residential Exposure Branch, Office of Pesticide Programs, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA). (1997). "Series 875Occupational and Residential Exposure Test Guidelines, Group B-PostApplication Exposure Monitoring Test Guidelines." Office of Pesticide Programs, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA). (1998). "PHED Surrogate Exposure Guide. Estimates of Worker Exposure from the Pesticide Handlers Exposure Database, version 1.1." Office of Pesticide Programs, Washington, DC. van Goldstein Brouwers, Y. G. c., Marquat, J., and van Hemmen, J. J. (1996). "Assessment of Occupational Exposure to Pesticides in Agriculture. Part VI. Protocol for the Use of Generic Exposure Data." TNO Report V 96.120, TNO Nutrition and Food Research Institute, The Netherlands. van Hemmen, J. J. (1992). Agricultural pesticide exposure data bases for risk assessment. Rev. Environ. Contam. Toxieo!. 126, 1-85. Versar. (1991). Pesticide Handlers Exposure Database (PHED). Data Entry Diskette User's Guide, version 1.0. Report prepared by Versar Inc., Springfield, VA for the U.S. EPA, Office of Pesticide Programs, Health Effects Division, Occupational and Residential Exposure Branch. Wang, R G. M., Franklin, C. A, Honeycutt, R. c., and Reinert, J. (eds.). (1989). "Biological Monitoring for Pesticide Exposures," ACS Symposium Series, Vo!. 382. Am. Chem. Soc., Washington, DC. Zweig, G., Leffingwell, J. T, and Popendorf, W. J. (1985). The relationship between dennal pesticide exposure by fruit harvesters and dislodgeable foliar residues.1. Environ. Sei. Health B20(l), 27-59.
CHAPTER
21 Factors That Affect Pesticide Metabolism and Toxicity Emest Hodgson North Carolina State University
21.1 INTRODUCTION Pesticides have been implicated as potential causative agents in the expression of many toxic endpoints (Hodgson and Levi, 1996). Acute toxicity is relatively easy to demonstrate and investigate, and toxic outbreaks have resulted from misuse of almost every type of pesticide: organochlorines such as DDT, lindane, and chlordecone; chlorinated camphenes such as toxaphene; the cyclodienes aldrin and dieldrin; organophosphate and carbamate acetylcholinesterase inhibitors; organomercury fungicides; inorganics; and others. Although the incidence of both individual and large scale acute poisoning episodes has fallen dramatically in developed countries, they are still common in third world countries. In the case of chronic toxicity, cause and effect relationships are more difficult to discern. However, based on studies with experimental animals and some epidemiological evidence, various pesticides have been implicated as at least potential causes of carcinogenesis and reproductive and developmental effects, as well as various manifestations of neurotoxicity. Although the toxicity of pesticides, as with that of any toxicants, must involve the interaction of the pesticide or one or more of its metabolites with a target macromolecule, toxicity cannot be thought of as a single defining molecular event. On the contrary, the expression of a toxic endpoint is the final event in a cascade of events that begins with exposure. As shown in Figure 21.1, this cascade involves absorption, distribution, metabolism to reactive metabolites, further distribution, and interaction with target molecules. At the same time, pesticides may be metabolic ally detoxified or excreted or the potentially toxic lesion may be repaired. Factors that affect any of these interactions may affect the ultimate expression of toxicity, although, in general, metabolism and interaction with target molecules play the most important roles. Thus, any of the factors that affect pesticide metabolism may affect the ultimate toxicity of the pesticide. Most, if not all, of the factors that affect pesticide metabolism and toxicity are discussed in detail elsewhere in this edition of Handbook of Pesticide Toxicology Volume I. Principles
the Handbook of Pesticide Toxicology. The following summary is an attempt to provide an overall integration of the factors that affect pesticide metabolism and toxicity, with appropriate examples.
21.2 EXPOSURE It is axiomatic that no matter how hazardous a pesticide may be, exposure must occur before toxicity can be manifested. The importance of exposure becomes obvious when occupational poisoning cases are correlated with particular occupations (Kilgore, 1988), where it is clear that those occupations that have the highest potential for exposure, such as ground applicators, also generate the highest number of poisoning cases. Considerable attention is now being paid to the measurement of exposure, primarily in connection with regulation of pesticides under the Federal Insecticide, Fungicide and Rodenticide Act or Food Quality Protection Act or with major epidemiological studies such as the Agricultural Health StUdy. Pesticide exposure is discussed in detail in Chapters 16-20.
21.3 ABSORPTION Following exposure, pesticides can be absorbed through the skin, the alimentary canal, or the respiratory system and it is apparent that both the rate and the extent of absorption may affect the expression of toxicity. Which system is quantitatively most important varies with the type of exposure as well as the chemical and physical properties of the pesticide. In general, however, in humans the skin is probably the most important portal of entry for pesticides because residues in food are usually small. Penetration of pesticides through the skin of humans or experimental animals can be estimated by a number of in vivo, ex vivo, or in vitro methods. Whereas all of these methods have advantages and disadvantages, taken together the results provide
507
Copyright © 200 1 by Academic Press. All rights of reproduction in any form reserved.
508
CHAPTER 21
Pesticide Metabolism and Toxicity
21.3.2 AGE
, ,
Exposure
I
Absorption at Portals of Entry
I
-'I--1
L-._ _ _ _ _D_i_str_ib_ut_io_n_to_B_O_dY_ _ _ _ _ _
,
,t
,t
Metabolism to More Toxic Metabolites
Metabolism to Less Toxic Metabolites
Metabolism 10 Conjugation Products
,
ExcretIon
!
21.3.3 SPECIES VARIATION
I
,
Although it is generally assumed that skin absorption in the young will be greater than that in older animals, this is not always the case. Thus, although absorption of atrazine, carbaryl, chlorpyrifos, and hexachlorobiphenyl was greater in the young rat, absorption of carbofuran, captan, dinoseb, disodium methanearsonate, monosodium methanearsonate, nicotine, and parathion was greater in the adult (Shah et al., 1987).
Distribution
, , I
Interaction with Macromolecules (Proteins. DNA. RNA. Receptors. elc.)
I
Variations in rate and extent of absorption may occur as a result of anatomical and functional differences between species. Thus rodent skin is generally not a good model for human skin, whereas pig skin, which is structurally similar to human skin, has proven to be a useful model for human pesticide absorption studies.
Toxic ENects (Genetic. Carcinogenic. Reproductive. Immunotoxic. etc.)
21.3.4 CHEMICAL PROPERTIES Figure 21.1
Sequence of events following exposure of an animal to a toxicant.
considerable insight into the factors that affect the rate and the extent of dermal absorption of pesticides in mammals. In recent years, the isolated perfused porcine skin flap has proven to be invaluable as a model for human dermal absorption. This and other methods, the findings derived from them, and the factors that affect penetration are all discussed in detail by Baynes and Riviere in Chapter 22. Some of the more important factors that affect dermal penetration are the following.
Properties such as molecular weight, reactive groups, lipophilicity, and ionization constants can all affect dermal absorption. The various constituents of pesticide formulations can also affect the rate of uptake of the active ingredient. In experimental studies, the use of acetone as a solvent clearly enhanced penetration more than other solvents such benzene and corn oil. The insect repellent DEET has been shown to enhance penetration in some cases (e.g., 2,4-D amine; Moody et aI., 1992) and inhibit it in others (e.g., permethrin; Baynes et al., 1997; Baynes and Riviere, 1998).
21.3.1 ANATOMICAL SITE
21.3.5 ENVIRONMENTAL FACTORS: TEMPERATURE, HUMIDITY, AND OCCLUSION
Variations in skin thickness may be large (e.g., between palmar and forearm) and would be expected to affect penetration (Feldman and Maibach, 1974), but regional variations in vasculature and blood flow may also be important (Monteiro-Riviere, 1991; Qiao et aI., 1993). The absorption of parathion in pigs (Qiao et aI., 1993), pyrethrin in humans (Wester et al., 1994) and lindane, N,N -diethyl-m-toluamide (DEET), permethrin, aminocarb, and fenitrothion in rhesus monkeys (Moody and Franklin, 1987; Moody and Ritter, 1989; Moody et al., 1987, 1989; Sidon et aI., 1988) have all been shown to vary with anatomical site. However, despite differences in dermal thickness and structure, in some cases anatomical site may have little or no effect on penetration as, for example, in the absorption of the acid and amine forms of 2,4-D in rhesus monkeys (Moody et aI., 1990, 1992).
These factors may all affect the rate of penetration of pesticides. Studies to date are summarized in Chapter 22. It is important to note that although many factors affect the rate and extent of dermal penetration, these factors may have little or no effect on the expression of toxicity. For example, increasing the rate of penetration of a pesticide that is rapidly detoxified would have no effect until the rate of uptake exceeded the capacity of the detoxifying enzymes, a concentration that, in practical terms, may never be attained. On the other hand, increasing the penetration rate of a pesticide that is rapidly metabolized to a more toxic metabolite might well increase the expression of toxicity. Carbaryl is an excellent example of the former. Although of low mammalian toxicity, it is known to penetrate human and animal skin much more readily than most other pesticides.
21.5 Metabolism
21.4 DISTRIBUTION Following absorption, pesticides and other toxicants circulate throughout the body, either to sites of metabolism or sites of action. Similarly, active metabolites formed at sites other than the site of action must also circulate to cause toxicity. Finally, circulation to organs of storage or excretion is necessary for either temporary or permanent clearance. Pharmacokinetics, the study of absorption, distribution, metabolism, and elimination of chemicals, relies heavily on mathematical models to describe the rate and the extent of these processes. Although much has been learned of the pharmacokinetic parameters involved in the action of clinical drugs and of how modification in these parameters may affect efficacy or toxicity, little is known about how these parameters and their modification affect the ultimate expression of toxicity from pesticides. The fundamentals of pharmacokinetics as they may be applied to pesticide toxicity are described by Dix in Chapter 24.
21.5 METABOLISM Pesticides are metabolized by numerous Phase I and Phase 11 xenobiotic-metabolizing enzymes and in addition to serving as substrates, they may also function as inhibitors and/or inducers. These enzymes, the reactions they catalyze, and their inhibition and induction, insofar as they involve pesticides, are discussed in detail by Hodgson and Levi in Chapter 23. Whereas metabolism of pesticides may bring about either detoxication or activation of the parent molecule, the balance between these two outcomes is often the determinant of toxicity. An extensive review of the metabolic activation of agrochemicals was published in 1995 (Hollingworth et aI., 1995). Thus any of the many factors that affect xenobiotic metabolism may affect the ultimate expression of toxicity. Some examples follow. 21.5.1 STRAIN AND SPECIES DIFFERENCES Species-specific differences in pesticide toxicity are frequently apparent and are usually the result of differences in metabolism --differences that may explain the specificity between target and nontarget species that is a characteristic of some important pesticides. One of the better known examples is the selectivity of malathion due to the presence of a malathion-hydrolyzing carboxylesterase in mammals that is seldom found in insects. Strain and species differences in the metabolism of toxic ants have been summarized by Rose and Hodgson (2000). The term tolerance as applied to toxic chemicals such as pesticides is used in two ways: (1) for individuals that acquire the ability to resist the toxic effects of a chemical as a result of prior exposure or (2) strains and/or populations that demonstrate differences in toxic response to a particular chemical without prior exposure. Resistance, on the other hand, involves a change in the genetic constitution of the population following exposure to a toxic chemical in such a way that a greater number of
509
individuals survive exposure than would have survived in the unexposed population. Thus resistance involves selection of the population by the chemical followed by inheritance by the generations that follow. In the case of microorganisms, mutations are often involved, with higher organisms, genes (often genes for xenobiotic-metabolizing enzymes) already present at low frequency, are selected for due to the higher mortality in individuals that do not possess these genes. Resistance is much better understood in target organisms than in nontarget species such as mammals. 21.5.2 CHEMICAL EFFECTS Chemical effects on pesticide metabolism and/or toxicity are the results of the interaction of two or more chemicals, one of which is the pesticide under consideration. The principal mechanisms behind such interactions are induction and inhibition. Some pesticides that act as substrates for xenobioticmetabolizing enzymes may also serve as inhibitors and/or inducers of these enzymes. These interactions have been summarized by Hodgson and Meyer (1997) and Rose and Hodgson (2000). 21.5.2.1 Inhibition The effects of inhibition may vary, depending upon the effect of the metabolic reaction being inhibited. For example, a reduction of the toxicity of azinphosmethyl to rats occurs when rats are pretreated with SKF-525A, a cytochrome P450 (P450 or CYP) inhibitor, presumably as a result of inhibition of ox on production. An increase in toxicity may result from inhibition of a detoxication reaction as in the case of the cocarcingenicity of piperonyl butoxide and the carcinogenic freons, 112 and 113. Whereas the mode of action of the methy lenedioxypheny1 synergists such as piperonyl butoxide is inhibition of toxicant metabolism by P450 (Dahl and Hodgson, 1979; Hodgson and Philpot, 1974), they have considerable potential to affect the toxicity of pesticides to nontarget as well as target species. Organophosphorus insecticides such as parathion are activated by the P450-dependent monooxygenase system to the corresponding oxons. During the course of this reaction highly reactive sulfur is released, which binds to the heme iron of P450 isoforms, effecting an irreversible inhibition of these enzymes (Halpert et aI., 1980; Kamataki and Neal, 1976), an inhibition that, at least in the case of fenitrothion, is isoform specific (Levi et aI., 1988). The effects of nonpesticidal inhibitors on pesticide metabolism and toxicity may also be of importance. For example, it has been shown (Agyeman and Sultatos, 1998) that treatment of mice with cimetidine, a clinically used H2-histamine antagonist, can effect a moderate increase in parathion toxicity, apparently due to its inhibitory effect on various P450 isoforms. 21.5.2.2 Synergism, Potentiation, and Antagonism The terms synergism and potentiation have been used and defined in several ways (Hodgson, 1997), but in all cases they
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involve toxicity that is greater when two or more compounds are given sequentially or in close sequence than would be expected from a consideration of their toxicities when administered alone. Only one of the two compounds may have appreciable toxicity when given alone. This is the case in the increase in toxicity of insecticides to insects and mammals brought about by cotreatment with methylenedioxyphenyl synergists such as piperonyl butoxide, tropital, or sexamex, the latter compounds having little toxicity when given alone (Hodgson and Levi, 1998; Jones, 1998). The term potentiation is often used when both compounds have appreciable toxicity alone, but the combined toxicity is still greater than additive. This is the case with malathion and EPN; the latter inhibits the carboxyl esterase responsible for the low toxicity of malathion to mammals. The herbicide synergist, tridiphane, owes it effectiveness to its ability to inhibit plant glutathione transferases. At the same time, tridiphane is a peroxisome proliferator and inducer of epoxide hydrolase in rodents (Moody and Hammock, 1987) and may serve both as an inhibitor (Moreland et aI., 1989) and as an inducer (Levi et aI., 1992) of mammalian P450s. The many complex interactions and their potential effects on toxicity that are possible with biologically reactive pesticides such as tridiphane are difficult to predict, but, particularly from a regulatory and safety viewpoint, should always be considered. Antagonism is that circumstance when the combined toxicity of two toxicants is less than that expected from consideration of their separate toxicities. Apart from that due to induction of xenobiotic-metabolizing enzymes (see subsequent text), antagonism is not known to occur to any extent with combinations of pesticides.
21.5.2.3 Induction Induction of xenobiotic-metabolizing enzymes may cause either an increase or a decrease in pesticide toxicity. Induction of P450 has been studied more intensively than induction of other xenobiotic-metabolizing enzymes, but, nevertheless, induction of almost all Phase I and Phase 11 enzymes has been seen. Effects on pesticide toxicity may be seen when either a pesticide inducer or a nonpesticide inducer causes changes in the metabolism of a pesticide. At the same time, pesticides may function as inducers, thus having the potential to affect the metabolism, not only of pesticides, but also of non pesticide substrates such as clinical drugs. Because of the relatively long period of time necessary for induction, a decrease in toxicity caused by induction is not usually classified as antagonism. Inducers of P450 fall into several different classes based on their mechanism of induction and the isoform specificity of the induced protein. These classes and some examples are outlined in Table 21.1, whereas many pesticide inducers are listed in Table 23.5 in Chapter 23. Although most of these studies were carried out before methods were available to determine which isoform of P450 was induced, it is apparent that induction can affect in vivo endpoints. For example, piperonyl butoxide increased parathion toxicity in the mouse after 1 hour (inhibition), but caused a decrease after 48 hours (induction)
Table 21.1 Induction of Cytochrome P450 Isofonns by Pesticides and Other Xenobiotics Inducer category Ah receptor-dependent
Example
Isofonn induced
3-MethyIcholanthrene
IAI,IA2
TCDD Piperonyl butoxide Unknown, non-Ah
Piperonyl butoxide
IA2,IBI
Phenobarbital
2BI,2B2
receptor-dependent Phenobarbital
DDT Mirex PCN/glucocorticoid
Dexamethasone
3AI
Erythromycin Ethanol
Ethanol
2EI
Acetone Imidazole Peroxisome proIiferator
Clofibrate
4AI
Tridiphane Fenvaleric Acid
(Kamienski and Murphy, 1971). Dieldrin induction can increase the in vitro metabolism of organophosphates such as chlorfenvinphos (Wright et aI., 1972). Although pesticides from many chemical and use classes are known to be inducers, including synergists such as piperonyl butoxide, chlorinated hydrocarbon insecticides such as DDT and DDT analogs, chlordane and dieldrin, herbicides such as monuron, diuron, and atrazine, and herbicide synergists such as tridiphane, the significance of such induction is not always apparent. The effects of non pesticide inducers on pesticide metabolism are also listed in Chapter 23. Although such effects have commonly been observed, their significance in terms of toxic endpoints in vivo are not always readily apparent. Because it is not possible to know the mechanism of induction or isoform specificity in most of these early studies, it is probable that many pesticide inducers act in a manner similar to phenobarbital. More recent studies (Dai et aI., 1998) have shown this to be true of mirex and chlordecone, whereas piperonyl has been shown not only to induce CYP 2b lOin the mouse (like phenobarbital), but also CYP lA2 and CYP lEl by both Ah receptor-dependent and Ah receptor-independent mechanisms (Hodgson and Levi, 1998; Ryu et aI., 1996; Ryu and Hodgson, 1999). It should be noted that induction by pesticides may, in fact, be due to their metabolites. For example, Morisseau et al. (1999) showed that the (R) enantiomer of fenvaleric acid is a peroxisome proliferator, which induces CYP 4Al, whereas the more active (S) enantiomer induces CYP2B and microsomal epoxide hydrolase activities, thus raising the possibility that high exposure to pyrethroid insecticides could interact with metabolism and/or toxicity of other xenobiotics. The effects of pesticides on hepatotoxicity may be due to induction of P450 isoforms as in the effect of mirex and
21.6 Excretion chlordecone on acetaminophen toxicity (Fouse and Hodgson, 1987), but this is not always the case. The potentiation of carbon tetrachloride hepatotoxicity by chlordecone (but not mirex) apparently is due to an effect that involves a suppression of tissue repair in the recovery phase of carbon tetrachloride toxicity (Mehendale, 1994). The role of mirex (but not chlordecone) as a potent tumor promoter is also independent of the role of mirex as an enzyme inducer (see Hodgson and Meyer, 1997 for references). 21.5.2.4 Effects on Multienzyme Systems Enzyme systems such as the P450-dependent monooxygenase systems and the flavin-containing monooxygenase system (FMO) may have many substrates in common. However, the products may vary and have different toxic potencies. Thus it is important to know the relative contributions of different metabolizing systems toward the same pesticide substrate as well as the factors that may change the relative contribution. The insecticide phorate provides an excellent model for investigation of the role of multi enzyme systems in pesticide metabolism. It undergoes a complex series of oxidations and both P450 and FMO are involved. Whereas the products are generally more toxic than the parent compound, this complex reaction sequence is, therefore, an activation sequence. FMO forms only one product from phorate, namely phorate sulfoxide, whereas P450 yields phorate sulfoxide as well as additional products. The sulfoxidation reaction is stereospecific: FMO produces the (-) sulfoxide and several P450 isoforms produce the (+) sulfoxide. Although both (+) and (-) sulfoxides are substrates for further metabolism by all P450 isoforms tested, the ( +) sulfoxide is always prefered to the ( - ) sulfoxide. The relative contribution of FMO to sulfoxide is gender dependent, being higher in female mice than in males, and although overall phorate oxidation is higher in liver than in any other tissue, the activity of FMO relative to that of P450 is higher in lung, kidney, and skin tissue. The FMO contribution may be as high as 90% in renal microsomes from female mice. The contribution of P450 relative to FMO is increased by treatment in vivo with inducers of P450 such as phenobarbital or decreased by P450 inhibitors such as piperonyl butoxide (Hodgson and Levi, 1996; Kinsler et aI., 1988, 1990; Tynes and Hodgson, 1983, 1985). 21.5.3 GENETIC POLYMORPHISMS Polymorphisms are known to exist in the genes that code for many xenobiotic-metabolizing enzymes, including several P450 isoforms, FMO, alcohol dehydrogenase, and paraoxonase. The extent, nature, and significance of polymorphisms in xenobiotic-metabolizing enzymes has been summarized by Hodgson and Goldstein (2000). They are defined as inherited monogenetic traits that exist in the population in at least two genotypes (two or more variant alleles) that are stably inherited. The importance of polymorphisms in the metabolism of clinical drugs has been well documented. Populations can frequently
511
be separated into extensive metabolizers or poor metabolizers on the basis of polymorphisms in drug metabolisms, and these polymorphisms may have critical importance for both efficacy and toxicity of many drugs. Polymorphisms may exist as variations in the base sequence in the coding region of the gene, in which case different polymorphic genes produce slightly different enzyme proteins that may have different substrate specificities. On the other hand, they may exist as variations in the regulatory sequences of the gene and control the expression of the enzyme protein. Pharmacogenetics has been defined as study of the hereditary basis of differences in response to drugs (Nebert, 1997). By analogy, toxicogenetics may be defined as the study of hereditary differences in response to toxicants. Although it is clear that polymorphisms in xenobioticmetabolizing enzymes could be of considerable importance in pesticide metabolism and possibly in defining populations at risk for pesticide poisoning, knowledge of isoform specificity for pesticide metabolism is sparse. It has been demonstrated (Coleman et aI., 1999,2000; Hodgson et aI., 1998; Tang et aI., 2000) that the metabolism of alachlor and related herbicides, as well as phorate, chlorpyrifos, and other organophosphorus insecticides, is CYP-isoform specific in humans and that several of the isoforms involved (CYP2Cs and CYP3A4) are known to be polymorphic. The known human polymorphism in paraoxonase (Costa et aI., 1999) would appear to be of considerable potential importance in the human toxicity of organophosphorus compounds. The two polymorphic forms, Q and R, contain either a glutamine (Q) or an arginine (R) at position 192. The type Q allele shows higher activity toward sarin, soman, and diazoxon, whereas the R allele has higher activity toward paraoxon. Both alleles have similar activity toward chlorpyrifos oxon.
21.6 EXCRETION Although the use of urinary excretory products as biomarkers of exposure has received attention, mechanisms of excretion of pesticides and their metabolites in vertebrates have not been studied extensively. However, mechanisms for the excretion of pesticides and their metabolites do not seem to vary in any marked way from those for other xenobiotics. There is little evidence that factors that affect the rate of excretion have any significant effect on the ultimate expression of pesticide toxicity.
REFERENCES Agyeman, A. A., and Sultatos, L. G. (1998). The actions of the H2-blocker cimetidine on tbe toxicity and biotransformation of the phosphorothioate insecticide parathion. Toxicology 128(3), 207-218. Baynes, R. E., and Riviere, J. E. (1998). Influence of inert ingredients in pesticide formulations on dermal absorption of carbaryL Am. 1. Vet. Res. 59, 168-175. Baynes, R. E., Hailing, K. B., and Riviere, J. E. (1997). The influence of dietbylm-toluamide (DEET) on the percutaneous absorption of permethrin and carbaryL Toxicol. Appl. Pharmacol. 144, 332-339.
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Coleman, S., Liu, S., Linderman, R., Hodgson, E., and Rose, R. L. (1999). In vitro metabolism of alachlor by human liver microsomes and human cytochrome P450 isoforms. Chem.-Bio!. Interact. 122,27-39. Coleman, S., Linderman, R., Hodgson, E., and Rose, R. L. (2000). Comparative metabolism of chloroacetamide herbicides in human and rat liver microsomes. Environ. Health Perspect. 108, 1151-1157. Costa, L. G., Li, W. F., Richter, R. J., Shih, D. M., Lusis, A., and Furlong, C. E. (1999). The role of paraoxanase (PONI) in the detoxication of organophosphates and its human polymorphism. Chem.-Biol. Interact. 119120, 429-438. Dahl, A. R., and Hodgson, E. (1979). The interaction of aliphatic analogs of methylenedioxyphenyl compounds with cytochromes P450 and P420. Chem.-Bio!. Interact. 27, 163-175. Dai, D., Rose, R. L., and Hodgson, E. (1998). Toxicology of environmentally persistent chemicals: Mirex and chlordecone. Rev. Toxicol. 2, 477-499. Feldman, R. J., and Maibach, H. I. (1974). Percutaneous penetration of some pesticides and herbicides in man. Toxico!. App!. Pharmacol. 28, 126--132. Fouse, B. L., and Hodgson, E. (1987). Effect of chlordecone and mirex on the acute hepatotoxicity of acetaminophen. Gen. Pharmacol. 18, 623--630. Halpert, J., Hammond, D., and Neal, R. A. (1980). Inactivation of purified rat liver cytochrome P450 during the metabolism of parathion (diethyl p-nitrophenyl phosphorothionate). 1. Bio!. Chem. 255, 1080--1089. Hodgson, E. (1997). Modification of xenobiotic metabolism. In "A Textbook of Modem Toxicology" (E. Hodgson, and P. E. Levi, eds.). Appleton and Lang, Stamford, CT. Hodgson, E., and Goldstein, J. E. (2000). Metabolism of toxicants: Phase I reactions and pharmacogenetics. In "Introduction to Biochemical Toxicology" (E. Hodgson, and R. C. Smart, eds.), 3rd ed. WiIey, New York. Hodgson, E., and Levi, P. E. (1996). Pesticides: An important but underused model for the environmental health sciences. Environ. Health Sci. 104, 97106. Hodgson, E., and Levi, P. E. (1998). Interactions of piperonyl butoxide with cytochrome P450. In "Piperonyl Butoxide, the Insecticide Synergist" (D. G. Jones, ed.), Chap. 3. Academic Press, San Diego. Hodgson, E., and Meyer, S. A. (1997). Pesticides. In "Comprehensive Toxicology" (I. G. Sipes, C. A. McQueen, and A. J. Gandolfi, series eds.), Vo!. 9. Hepatic and Gastrointestinal Toxicology (McCuskey and D. L. Earnest, volume eds.), pp. 369-387. Pergamon, Elsevier Science, New York. Hodgson, E., and Philpot, R. M. (1974). Interaction of methylenedioxyphenyl (1,3-benzodioxole) compounds with enzymes and their effects on mammals. Drug Metabo!. Rev. 3, 231-301. Hodgson, E., Cherrington, N., Coleman, S., Liu, S., Falls, J. G., Cao, Y., Goldstein, J. E., and Rose, R. L. (1998). Flavin-containing monooxygenase and cytochrome P450 mediated metabolism of pesticides: from mouse to human. Rev. Taxicol. 2, 231-243. HoIIingworth, R. M., Kurihara, N., Miyamoto, J., OltO, S., and Paulson, G. D. (1995). Detection and significance of active metabolites of agrochemicals and related xenobiotics in animals. Pure App!. Chem. 67, 1487-1532. Jones, G. D., ed. (1998). "Piperonyl Butoxide, the Insecticide Synergist." Academic Press, San Diego. Kamataki, T., and Neal, R. A. (1976). Metabolism of diethyl p-nitrophenyl phosphorothionate (parathion) by a reconstituted mixed-function oxidase enzyme system: Studies of the covalent binding of the sulfur atom. Mo!. Pharmacol. 12, 933-944. Kamienski, F. X., and Murphy, S. D. (1971). Biphasic effects of methylenedioxyphenyl synergists on the action of hexobarbital and organophosphate insecticides in mice. Toxicol. Appl. Pharmaco!' 18, 883-894. Kilgore, W. (1988). Human exposure to pesticides. In "International Toxicology Seminar: Environmental Toxicology" (P. M. Newberne, R. C. Shank, and M. Ruchirawat, eds.). Chulabhorn Research Institute and Mahidol University, Bangkok. Kinsler, S., Levi, P. E., and Hodgson, E. (1988). Hepatic and extrahepatic microsomal oxidation of phorate by the cytochrome P450 and FMOcontaining monooxygenase systems in the mouse. Pestic. Biochem. Physiol. 31,54-60. Kinsler, S., Levi, P. E., and Hodgson, E., (1990). Relative contributions the cytochrome P450 and FMO-containing monooxygenases to the microsomal
oxidation of phorate following treatment of mice with phenobarbital hydrocortisone, acetone and piperonyl butoxide. Pestic. Biochem. Physiol. 32, 174-181. Levi, P. E., Hollingworth, R. M., and Hodgson, E. (1988). Differences in oxidative dearylation and desulfuration of fenitrothion by cytochrome P-450 isozymes and in the subsequent inhibition of monooxygenase activity. Pestic. Biochem. Physiol. 32, 224-231. Levi, P. E., Rose, R. L., Adams, N. H., and Hodgson, E. (1992). Induction of cytochrome P450 4AI in mouse liver by the herbicide synergist tridiphane. Pestic. Biochem. Physiol. 44, 1-19. Mehendale, H. M. (1994). Amplified interactive toxicity of chemicals at nontoxic levels: Mechanistic considerations and implications to public health. Environ. Health Perspect. 102(Supp!. 9), 139-149. Monteiro-Riviere, N. A. (1991). Comparative anatomy, physiology and biochemistry of mammalian skin. In "Dermal and Ocular Toxicology: Fundamentals and Methods" (D. W. Hobson, ed.), Chap. I, pp. 3-71. CRC Press, Boca Raton, PL. Moody, D. E., and Hammock, B. D. (1987). The effect of tridiphane (2(3,5-dichlorophenyl)-2-(2,2,2-trichloroethyl) oxirane) on hepatic epoxide metabolizing enzymes: Indications of peroxisome proliferation. Toxico!. App!. Pharmacol. 89, 37-46. Moody, R. P., and Franklin, C. A. (1987). Percutaneous absorption of the insecticides fenitrothion and aminocarb in rats and monkeys. J. Toxicol. Environ. Health 20, 209-218. Moody, R. P., and Ritter, L. (1989). Dermal absorption of the insecticide lindane (1a,2a,3b,4a,5a,6b-hexachlorocyclohexane) in rats and rhesus monkeys: Effect of anatomical site. 1. Toxicol. Environ. Health 28, 161-169. Moody, R. P., Riedel, D., Ritter, L., and Franklin C. A. (1987). The effect of DEET (N,N -diethyl-m-toluamide) on dermal persistence and absorption of the insecticide fenitrothion in rats and monkeys. J. Toxico!. Environ. Health 22,471-479. Moody, R. P., Benoit, F. M., Riedel, D., and Ritter, L. (1989). Dermal absorption of the insect repellant DEET (N,N-diethyl-m-toluamide) in rats and monkeys: Effect of anatomical site and multiple exposure. J. Toxico!. Environ. Health 26, 137-147. Moody, R. P., Franklin, C. A., Ritter, L., and Maibach, H. I. (1990). Dermal absorption of the phenoxy herbicides 2,4-D, 2,4-D amine, 2,4-D isooctyl and 2,4,5-T in rabbits, rats rhesus monkeys and humans: A cross-species comparison, 1. Taxica!. Environ. Health 29, 237-245. Moody, R. P., Wester, R. c., Melendres, J. L., and Maibach, H. I. (1992). Dermal absorption of the phenoxy herbicide 2,4-D dimethylamine in humans: Effect of DEET and anatomic site. J. Toxicol. Environ. Health 36, 241-250. Moreland, D. E., Novitsky, W. P., and Levi, P. E. (1989). Selective inhibition of cytochrome P450 isozymes by the herbicide tridiphane. Pestic. Biochem. Physial. 35, 42-49. Morisseau, c., Derbel, M., Lane, T. R., Stoutamire, D., and Hammock, B. D. (1999). Differential induction of hepatic drug-metabolizing enzymes by fenvaleric acid in male rats. Toxico!. Sci. 52, 148-153. Nebert, D. W. (1997). Polymorphisms in drug metabolizing enzymes: What is their clinical relevance and why do they exist? Am. J. Hum. Genet. 60, 265271. Qiao, G. L., Chang, S. K., and Riviere, J. E. (1993). Effects of anatomical site and occlusion on the percutaneous absorption and residue pattern of 2,6[ring_ 14 C]parathion in viva in pigs. Toxico!. App!. Pharmacol. 122, 131138. Rose, R. L., and Hodgson, E. (2000). Adaptation to toxicants. In "Introduction to Biochemical Toxicology," 3rd ed. (E. Hodgson, and R. C. Smart, eds.). WiIey, New York. Ryu, D.-Y., and Hodgson, E. (1999). Constitutive expression and induction of CYPIBI mRNA in the mouse. 1. Biochem. Molec. Toxicol. 13,249-251. Ryu, D.-Y., Levi, P. E., Fernandez-Salguero, P., Gonzalez, F. G., and Hodgson, E. (1996). Piperonyl butoxide and acenaphthylene induce CYPIA2 and CYPIBI mRNA in Ah receptor knockout (AHR-I-) mouse liver. Mo!. Pharmacal. 50, 443-446. Shah, P. v., Fisher, H. L., Sumler, M. R., Monroe, R. J., Chernoff, N., and Hall, L. L. (1987). Comparison of the penetration of 14 pesticides through the skin of young and adult rats. J. Toxico!. Environ. Health 21, 353-366.
References
Sidon, E. W., Moody, R. P., and Franklin, C. A. (1988). Percutaneous absorption of cis- and trans-perrnethrin in rhesus monkeys and rats: Anatomical site and interspecies variation. Toxico!. Environ. Health 23, 207-216. Tang, l., Rose, R. L., and Hodgson, E. (2000). Unpublished observations. Tynes, R. E., and Hodgson, E. (1983). Oxidation ofthiobenzamide by the FADcontaining and cytochrome P4S0-dependent monooxygenases of liver and lung microsomes. Bioehem. Pharmaeol. 32, 3419-3428.
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Tynes, R. E., and Hodgson, E. (198S). Magnitude of involvement of the mammalian flavin-containing monooxygenase in the microsomal oxidation of pesticides. J. Agrie. Food Chem. 33, 471-479. Wester, R. c., Bucks, D. A., and Maibach, H. 1. (1994). Human in vivo percutaneous absorption of pyrethrin and piperonyl butoxide. Food Chem. Toxieol. 32, SI-S3. Wright, A. S., Potter, D., Wooder, M. E, Donninger, c., and Greenland, R. D. (1972). Effects of dieldrin on mammalian hepatocytes. Food Cosmet. Toxieol. 10,311-322.
CHAPTER
22 Pesticide Disposition: Dermal Absorption Ronald E. Baynes and Jim E. Riviere North Carolina State University
The skin is a complex tissue with a large surface area whose primary function is to protect the body from physical or chemical insult, to thermoregulate and to simultaneously prevent water loss from the body. Dermal absorption of any chemical requires movement from the environment across this barrier, which is a biochemical milieu of complex lipids and proteins. It is therefore important to understand these unique anatomical and biochemical features of skin and the mechanisms by which chemicals cross this barrier. Experimentally, there are several in vitro, ex vivo, and in vivo models that have been be used to estimate dermal absorption of pesticides in humans. Although in vivo methods are the "golden standard," each of these methods has its respective weaknesses and strengths for accurately predicting dermal absorption of pesticides. Dermal absorption assessment is further complicated by species, age, and sex differences, and differences between anatomical sites within a species. More importantly, dermal absorption in rodent skin is not always equivalent to that in human skin. Finally, dermal absorption is dependent on the physicochemical properties of the pesticide, the formulation, and the environmental conditions. The pesticide applicator is often clothed and operating in extreme environments, not standard laboratory conditions.
22.1 INTRODUCTION Percutaneous absorption is reported as the possible route of entry in 65-85% of all cases of occupational exposure with pesticides (Galli and Marinovich, 1987). Spray or dusting of pesticides can result in disposition of 20-1700 times the amount deposited in the respiratory tract (Feldmann and Maibach, 1974). Epidemics of pesticide poisoning following cutaneous exposure have been reported for nonoccupational uses (Ferrer and Cabral, 1993). These cases often involved accidental contamination of infant clothing or exposure to talcum powder with pesticides (Martin-Bouyer et aI., 1983). These anecdotal case reports, coupled with dermal exposure estimated from various Handbook of Pesticide Toxicology
Volume 1. Principles
direct and indirect dosimetric experiments are often the only available human data with which to perform dermal absorption assessment. In spite of such limited data, it is possible to estimate dermal absorption by extrapolation from dermal exposure data. Algebraic equations that take into account exposure time and the chemical nature of the compound (lipophilicity and molecular weight) have been presented for estimating dermal absorption (Cleek and Bunge, 1993; Potts and Guy, 1992). However, the biological variability inherent in skin complicates the quantitative prediction of the rate and extent of penetration (into skin) and absorption (through skin) into the blood stream. Furthermore, the diversity of protocols and methods for calculating dermal exposure and ultimately dermal bioavailability makes dermal absorption assessment a difficult process. This chapter will initially focus on mechanisms and pathways of dermal absorption and experimental models used to assess dermal absorption. The effects of biological variability, pesticide chemistry and formulations, and environmental variables that influence dermal absorption will then be discussed.
22.2 PATHWAYS OF DERMAL ABSORPTION The skin is relatively impermeable to aqueous solutions and ions, but it may be permeable in varying degrees to a large number of drugs or xenobiotics. Drug or xenobiotic delivery pathways can hypothetically involve intercellular and intracellular passive diffusion across the epidermis and dermis and/or transappendageal routes via hair follicles and sweat pores. Transappendageal pathways are currently considered to contribute very little to the dermal transport of most drugs compared to transport across the epidermis (Barry, 1991). It is possible for very small and/or polar molecules to penetrate through these appendages or shunts, but very unlikely for many classes of highly lipophilic pesticides. The stratum corneum cell layer in humans (10-50 !-lm) and pigs (15 !-lm) is nonviable and is considered to be the rate-limiting barrier in per-
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CHAPTER 22
Pesticide Disposition: Dermal Absorption
cutaneous absorption of many drugs and pesticides (MonteiroRiviere et aI., 1990). Most available research has concentrated on the stratum corneum as the primary barrier to absorption, although the viable epidermis (ca. 80 !-lm in humans and 60 !-lm in pigs) and dermis (3-5 mm in humans) are now thought to contribute significantly to the percutaneous penetration of drugs and ultimately their bioavailability. Scheuplein (1972) proposed that polar drugs diffused through the hydrated keratin of the dead cells in the stratum corneum, whereas nonpolar drugs traversed the intracellular lipid. The accepted hypothesis is that the dominant pathway for polar molecules resides in the aqueous region of the intercellular lipid with the hydrophobic region of the lipid chains providing the nonpolar route (Elias, 1981). The intercellular region as depicted in the "brick and mortar" model of the stratum corneum, now considered the most likely path for absorption of lipophilic drugs, is filled with neutral lipids (complex hydrocarbons, free sterols, sterol esters, free fatty acids, and triglycerides), which makes up 75% of the total lipids, and polar lipids, such as phosphatidylethanolamine, phosphatidy1choline, lysolecithin, ceramides, and glycolipids (Magee, 1991). Successive tape stripping, delipidization techniques, and use of epidermis from heat or chemical separation techniques have been used by investigators to demonstrate the influence of the stratum corneum and the lipid domain on penetration of hydrophilic and lipophilic chemicals. In preparation of intact skin sections for in vitro studies, efforts should be made to ensure that hair clipping or any other chemical or mechanical cleaning of the skin surface does not alter the stratum corneum. Percutaneous absorption through the intercellular pathway is by passive diffusion and it is often correlated to the partition coefficient. The rate of absorption of the penetrant and can be described by Fick's law of diffusion: fl
_ UX -
diffusion coefficient x surface area x partition coefficient x conc. gradient skin or membrane thickness
Theoretically, the determinants of flux or diffusion rates for a given skin section may be altered experimentally through manipulation of drug or pesticide formulations. If lipid solubility increases, the penetrant may remain in the stratum corneum and form a reservoir. Some compounds can also form a reservoir in the dermis. These scenarios can prolong absorption half-life, which can also prolong the body burden of the penetrant. Determinants other than those described by Fick's law of simple diffusion of drugs should be considered in understanding percutaneous absorption mechanisms. For example, lipid content and lipid composition are not described in the equation other than how they affect the magnitude of the diffusion or partition coefficient. One study demonstrated that differences in the thickness and number of cell layers in the stratum corneum (SC) are insufficient to account for differences in percutaneous transport of water and salicylic acid across leg and abdominal skin (Elias et aI., 1981). The data demonstrated that total lipid concentration might be the more critical factor governing skin permeability. The inverse correlation expected from Fick's law did not apply because more water and salicylic acid penetrated
thicker skin and had less lipid by weight. Whereas there were no differences in percent composition of phospholipids, neutral lipids, and sphingolipids between leg skin and abdominal skin, total lipid content may be a determinant in drug transport. Alternatively, the actual intercellular "path length" of the lipid domain may be the critical factor, rather than just skin thickness. This study was not conclusive because very few chemicals were evaluated. In the presence of a simple cosolvent such as water, polar lipids (e.g., phosphoglyceride) associate into various types of aggregates that resemble micelles formed with soaps. In these structures, the hydrocarbon chain is hidden from the external aqueous environment, but the hydrophilic heads are exposed on the surface. Under these conditions, hydrophilic drugs and pesticides may traverse the Se. It has been proposed that solvents and surfactants may also structurally alter the proteinaceous and lipid components of the stratum corneum; thus altering the absorption pathway of co-administered toxicants. These interactions will be further described when mixture and formulation effects are discussed.
22.3 METHODS OF STUDYING DERMAL ABSORPTION 22.3.1 IN VITRO METHODS In vitro technology can be used to obtain accurate measurements of dermal penetration, absorption, and metabolism of topically applied drugs and pesticides. The choice of in vitro systems is dependent on experimental objectives, and limitations inherent to any in vitro system ought to be recognized. Static cell diffusion systems with two-chamber cells were initially used to quantitate percutaneous absorption of drugs and toxicants. The two-chambered cell can be used when steady state kinetics is required to study mechanisms of absorption and to quantitate permeation through skin using Fickian diffusion (Bronaugh and Collier, 1993). However, skin is excessively hydrated with this two-chamber system and the donor chamber does not mimic environmental conditions. The onechamber static diffusion cell is now more often used than the two-chamber static diffusion cell system. More recently, flow-through diffusion cell systems (a one-chamber system) as described by Bronaugh and Stewart (1985) have contributed significantly to quantitating the rate and extent of dermal absorption of drugs and pesticides in humans, rodents, and pigs. In many instances, permeation of lipophilic and hydrophilic chemicals into these in vitro systems was comparable to permeation in vivo. The major benefit of the flow-through diffusion system over one-chamber static diffusion systems is that the skin is more likely to be viable and the penetrating drug or pesticide molecule is continually removed from the dermal reservoir, which is analogous to the continual perfusion of blood in the in vivo situation (Bronaugh and Stewart, 1985; Grummer and Maibach, 1991). Furthermore, lipophilic chemicals are
22.3 Methods of Studying Dennal Absorption
more likely to be absorbed into a saline receptor fluid with a continuous flow system than with a static cell diffusion system (Wester et aI., 1985). The receptor volume of the diffusion cell is usually small ( <0.5 ml) so that it can be completely flushed out during sample collection intervals. Flow rates are usually set at 5-10 times the receptor volume (e.g., 2.5-5 ml/h) to allow for adequate removal of absorbed drug molecules. Other studies have demonstrated that changes in perfusion rate (Chang and Riviere, 1991; Crutcher and Maibach, 1979), receptor fluid constituents (Bronaugh and Stewart, 1984), environmental conditions (Chang and Riviere, 1991), and skin section thickness (Scott et aI., 1991) dramatically alter penetration of the marker chemical. A review of current literature suggests that the latter variable is often overlooked in the design of in vitro percutaneous studies. Furthermore, percutaneous absorption in vitro can underestimate in vivo absorption if the physical and biochemical properties of the receptor fluid are significantly different than that of blood and whole skin sections used in vitro. Recall that for some xenobiotics the availability of plasma protein for binding can influence systemic absorption. For many of the lipophilic pesticides, penetrating molecules are thought to enter the systemic circulation in vivo at the dermis/epidermis interface and do not necessarily traverse the full thickness of the dermis. With in vitro studies, the aqueous dermis can act as a significant, additional, artificial barrier for penetrating lipophilic chemicals such as DDT (Reifenrath et aI., 1991). Several studies have had difficulty demonstrating good agreement between in vitro and in vivo absorption with lipophilic chemicals because of the presence of excess dermis in final skin preparations (Bronaugh and Stewart, 1984; Bronaugh and Collier, 1993), whereas for more hydrophilic chemicals, good agreement between in vitro and in vivo absorption data is more often achieved. In one study, which used whole rat skin treated with carbaryl (Macpherson et aI., 1991a), considerably less penetration and absorption were observed when compared with data from in vivo studies (Shah etal., 1987a). Several other studies have demonstrated that use of skin sections that contain mostly epidermis and minimal dermis results in good agreement with in vivo data. For example, cypermethrin did not penetrate in vitro through rat whole skin, but penetrated epidermal membranes and predicted the in vivo absorption in rats (Scott and Ramsey, 1987). Furthermore, bovine serum albumen was used as the receptor fluid because it mimics blood and it facilitates partitioning of lipophilic chemicals from the skin sections into the receptor fluid. Grissom et al. (1987) also demonstrated that in vivo penetration of both hydrophilic and lipophilic compounds compared well with in vitro penetration when the hypodermis and most of the dermis are removed from mouse skin sections. This finding was consistent for isofenphos exposure for 24 hours in human skin (Wester et aI., 1992a), but not for DDT absorption in pig skin for 24 hours (Reifenrath et aI., 1991). In the latter study, greater levels of DDT were detected in the dermis in vitro than in vivo, again suggesting difficulty in DDT partitioning into perfusate.
517
22.3.2 EX VIVO METHODS There are several perfused skin preparations with an intact functional microvasculature. The perfused rabbit ear model, perfused pig ear model, in situ sandwich skin flap in athymic rats, and the hybrid rat-human sandwich flap have been developed, but each has severe limitations (Biren et aI., 1986; Pershing and Krueger, 1987). The isolated perfused porcine skin flap (IPPSF) is a unique ex vivo skin preparation that has many advantages over other ex vivo models and most in vitro systems. In addition to having an intact functional cutaneous microcirculation, predictions from IPPSF studies have correlated well with in vivo absorption data for several drugs and insecticides (Riviere et aI., 1986, 1995). In addition to the time and cost required to surgically prepare IPPSFs, the major drawback is that neither systemic-mediated immunological, physiological, neural feedback responses to topically applied toxicants can be observed. In spite of this, IPPSFs can be used to study skin distribution, transdermal drug delivery, and systemic targeting of anticancer drugs to tumor-bearing flaps (Vaden et aI., 1994; Williams and Riviere, 1990). IPPSFs are physiologically and biochemically viable and therefore can be used to assess cutaneous toxicity of topically applied chemicals (Monteiro-Riviere, 1993). The latter is most important because cutaneous toxicity as well as dermal absorption of various pesticide formulations can be assessed simultaneously. Our laboratory has repeatedly demonstrated the comparability of IPPSF absorption profiles and in vivo human absorption reported in the literature (Carver et aI., 1989; Chang et aI., 1994a, b; Riviere et aI., 1995; Srikrishna et aI., 1992). The compounds involved included the pesticides carbaryl, malathion, parathion, DFP, lindane, and paraquat. The comparison between IPPSF predicted and observed human absorption (percent dose) was specifically validated for other classes of chemicals: salicylic acid, 7.5 ± 2.6 vs. 6.5 ± 5.0; theophylline, l1.8 ± 3.8 vs. 16.9 ± 11.3; 2,4 dimethyl amine, 3.8 ± 0.6 vs. 1.1 ± 0.3; diethyl hexyl phthalic acid: 3.9 ± 2.4 vs. 1.8 ± 0.5; and p-aminobenzoic acid, 5.9 ± 3.7 vs. 15.3 ± 8.4 (Wester et aI., 1998). The discrepancy seen with the latter compound is related to the fact that the pKa of this compound is midway between the surface pH of porcine and human skin. Although in vivo studies can be used to estimate first-pass cutaneous biotransformation, it is limited in scope in accurately assessing cutaneous metabolism. With the IPPSF, there is no systemic metabolism to confound cutaneous metabolism, and our laboratory has demonstrated cutaneous metabolism for several pesticides, including parathion and carbaryl (Carver et aI., 1990; Chang et aI., 1994b). 22.3.3 IN VIVO METHODS As indicated earlier, there is an urgent need to establish whether there exists a relationship between in vitro and in vivo permeability of commonly used pesticides before in vitro data can be used for risk assessment purposes. Such relationships have been
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CHAPTER 22
Pesticide Disposition: Dermal Absorption
established to a limited extent for human and laboratory animal skin penetration for several pesticides (Chang et al., 1994b; Riviere et al., 1995; Scott et al., 1992; Wester et al., 1985). In many cases, there is better agreement between both methods for polar chemicals than for highly lipophilic chemicals. If no direct relationship exists to suggest a simple diffusion mechanism, it is also possible that factors such as sebum or sweat production rate, blood flow rate, and rate at which keratinized epithelium cells are shed may be involved in the transdermal delivery process. If this is the case, in vivo studies under various conditions must be conducted. If permeability data from the in vitro studies correlate to in vivo behavior, determination of the absorption mechanism will be easier. Unfortunately, most of the available in vivo dermal absorption and penetration data for pesticides in laboratory animals has many limitations. For example, it is more difficult to get consistent, reproducible data at early sample times using in vivo methods (Shah and Guthrie, 1983). However, a stronger correlation between in vivo and in vitro absorption of several pesticide formulations can occur over longer sample time periods (Scott et al., 1992). For many in vivo studies, urinary and fecal excretion and tissue residue data are the only information available regarding percutaneous absorption of pesticides. The major problem with this data base is that companion intravenous dosing studies have not been conducted and thus interpretation is hampered and quantitation of absolute bioavailability is impossible. These in vivo studies are usually unable to determine "first-pass" cutaneous metabolism for topically applied drugs or pesticides. For many of the pesticides, systemic metabolism may confound bioavailability data. Dermal absorption data may be inadequate for risk assessment purposes because of small sample size, unknown vehicle or solvent system, poor analytical method, and/or poor experimental design. These factors will be discussed in further detail in latter sections of this chapter. There is further conflicting in vivo and in vitro data in the literature regarding dermal absorption of permethrin, which in addition to its agronomic use, is impregnated in military uniforms and is applied to the scalp of children to treat for lice. The molecular weight and structure of permethrin suggest limited dermal absorption. However, previous in vivo studies have demonstrated 14-28% absorption of cis and trans isomers in the forehead and 5-12 % of both isomers in the forearm of rhesus monkey (Sidon et al., 1988) and as much as 63.8% absorption in 8 hours in mice (Shah et al., 1981) and 49-57% in 72 hours in rats (Shah et al., 1987a). Recent in vitro and ex vivo studies have demonstrated that absorption is minimal ( < 1%) in perfused rabbit, human, and pig skin (Bast and Kampffmeyer, 1996; Bast et al., 1997; Baynes et al., 1997; Franz et al., 1996). In addition to the obvious species differences, this discrepancy between in vivo and in vitro data may be related to radiochemical impurity, position of the 14C label, and in vivo permethrin metabolism, which can confound interpretation of cumulative urine data. In vitro absorption of fenvalerate in human newborn foreskin, which is very permeable, further demonstrated absorption of pyrethroid pesticides can be insignificant «5%) within 24 hours of exposure (Shehata-Karam et al., 1988).
For some pesticides, in vivo and in vitro data are comparable. For example, dermal absorption of malathion (6.8% at 24 hours and 8.2% dose at 5 days) in human volunteers (Feldmann and Maibach, 1974; Maibach and Feldmann, 1974) was comparable to in vitro absorption (8.77% dose at 24 hours) in human skin (Wester et al., 1996a) and in vivo absorption of malathion absorption in Yorkshire pigs (5.2% at 6 days; Carver and Riviere, 1989). IPPSF studies demonstrated a 10.8% bioavailability from in vitro absorption rates and pharmacokinetic modeling (Carver et al., 1989). One should also exercise caution with interpreting and comparing dermal pharmacokinetic data because different methods (direct vs. indirect) may be employed to determine skin penetration. The indirect method involves determination of radioactivity in excretory products over the course of the experiment and assumes that the fate of injected and topical doses is similar. The direct method involves sacrifice of animals at several time intervals and assay of tissues for labeled pesticides. One study determined percutaneous penetration of 4 J.Lg!cm 2 of carbaryl (naphthyl-1- 14C labeled) and parathion (U-ring labeled) in Dublin (SDD) male rats over a 5-day period by both methods (Shah and Guthrie, 1983). At 8 hours, skin penetration of carbaryl and parathion was 37 and 21 %, respectively, by the indirect method and 52 and 27%, respectively, by the direct method. At 120 hours, percent penetration was >91 % for carbaryl and parathion using direct and indirect methods. The penetration half-lives for carbaryl determined by the indirect and direct method were 10.34 and 4.75 hours, respectively. Neither method affected parathion penetration half-lives, which were 15 and 13 hours for indirect and direct methods, respectively. Clearly, the experimental method influenced determination of carbaryl penetration more than parathion penetration during the early time periods. The difference between the two methods for carbaryl absorption rates may be because carbaryl moves slowly from the carcass following initial penetration and into urinary pathways, and/or ring-Iabeled carbaryl requires substantially longer intervals to be excreted, because it would be expected to require conjugation. In addition, this study demonstrated that carbaryl penetrates rodent skin faster than parathion, although the extent of absorption was similar for both pesticides. Com!Jared to human data (Feldmann and Maibach, 1974), the rate and extent of absorption appears to be greater in rats than in humans following exposure to 4 J.Lg/cm2 of carbaryl for 120 hours. These differences are more likely due to rodent skin being thinner than human skin. 22.3.4 DERMAL BIOAVAILABILITY CALCULATIONS
In assessing bioavailability with in vivo studies, direct or indirect methods can be used. The direct method is a more massbalanced technique that requires animals to be sacrificed at specified time points to determine the extent of absorption. Bioavailability is the ratio of amount detected in viscera and body fluids to the amount applied to skin. Although this method accurately assesses bioavailability without need for a correc-
22.4 Factors that Affect Dennal Absorption
tion for metabolism and excretion, it is expensive and provides limited kinetic information. Indirect methods require measurement of pesticides in biological samples such as blood, urine, and feces. For accurate determination of dermal bioavailability, intravenous (IV) as well as dermal exposure studies should be conducted. Blood levels vs. time plots from both IV and dermal studies are used to determine the area under the blood concentration-time curve (AUC) and dermal bioavailability determined from the calculation dermal bioavailability
=
AUC(dennal) AUC(intravenous) x dose (intravenous) (x 100) dose(dennal)
Although using blood or plasma data is the most sensitive method for assessing dermal bioavailability, it may be inconvenient for small animals that have small blood volumes and for those situations where there are little or no detectable levels of the parent pesticide in blood or plasma. In these situations, cumulative urine concentration can be used to calculate the total mass of pesticide excreted in urine from time zero to infinity (Ae). Dermal bioavailability can be calculated via the formula dermal bioavailability
=
Ae (dennal) Ae (intravenous) X
dose (intravenous) (x 100) dOSe(dennal)
Whereas urine samples are convenient and inexpensive, this method assumes that tracer excretion into urine does not differ between IV and dermal routes. These studies may require creatinine data to correct for incomplete urinary excretion. This method is limited if a small fraction of the parent pesticide is found in urine. Confounding factors include variable urine pH, urinary flow rate, and inconsistent collection. Cutaneous metabolism can be determined from the difference between (a) total 14C excretion and parent bioavailability or (b) between metabolite 14C excretion and systemic metabolism. If blood or plasma data are available, the difference between total 14C bioavailability and parent bioavailability may be used to assess cutaneous metabolism. Determining the disappearance of the xenobiotic from the skin surface (Hall and Shah, 1990) can also assess dermal bioavailability. This method assumes there is no evaporative loss of xenobiotic and that which has penetrated below the surface is potentially bioavailable. Bioavailability is therefore the difference between topical dose and the amount remaining on the skin surface. For chemicals with slow and limited absorption, the error of measurement may be significant. Removal of the stratum corneum by successive stripping with cellophane tape (at least 12 tape strips) provides information about the relative disposition and sequestering of the xenobiotic and potential to be systemically absorbed. Although this is very important for very lipophilic pesticides, shedding of the stratum corneum reduces dermal bioavailability and this method will most likely
519
overestimate the risk of dermal absorption for chemically related pesticides.
22.4 FACTORS THAT AFFECT DERMAL ABSORPTION 22.4.1 ANATOMICAL SITE DIFFERENCES Regional variation in skin permeability in different body sites may be related to skin thickness, number of cell layers, cell size of the epidermis and stratum corneum, and distribution of hair follicles and sweat pores. Because of thick layers of stratum corneum, permeability in palmar and plantar skin is expected to be less than that in the scalp or forearm (Feldmann and Maibach, 1974). Data from several studies suggest that regional variation in vascular anatomy and blood flow should also be considered (Monteiro-Riviere et aI., 1990; Qiao et aI., 1993). Various studies have demonstrated regional variation in penetration of drugs and pesticides in pig skin (Qiao et aI., 1993; Qiao and Riviere, 1995), rat skin (Bronaugh, 1985), and rhesus monkey skin (Qiao et aI., 1993; Wester et aI., 1980a). These studies further demonstrated that parathion penetrates nonoccluded pig skin in the decreasing order back > shoulder > buttocks > abdomen; for occluded skin, the order is back > abdomen > buttocks > shoulder. Wester et al. (1994) also demonstrated that pyrethrin absorption through the human forearm is less than the predicted absorption in the human scalp. This anatomical difference is somewhat consistent with lindane absorption through the forearm (18%), forehead (34%), and palm (34%) of rhesus monkeys (Moody and Ritter, 1989). This anatomical range for lindane is similar to that for dermal absorption of DEET (diethyl-m-toluamide) in rhesus monkeys (Moody et aI., 1989). There is also significant data to suggest that dermal absorption of permethrin, aminocarb, DEET, and fenitrothion in monkey foreheads is twice that in monkey forearms (Moody and Franklin, 1987; Moody et aI., 1987; Sidon et aI., 1988). However Moody et al. (1990, 1992) demonstrated that there is no difference between the absorption of acid and amine forms of 2,4-D in rhesus monkey forearm and forehead and forearm and palm regions. The palmar absorption data conflict with the accepted dogma that absorption through palmar skin should theoretically be less than that in forearm skin because of the thickness of the stratum corneum in palmar skin (Maibach et aI., 1971). It is proposed that because of the hydrophilic nature of 2,4-D-amine, absorption can occur through polar routes such as eccrine glands, which are more frequent in the palmar skin than in the forearm skin. This anatomical difference does not explain the discrepancy with lindane, which is more lipophilic than 2,4-D and least likely to be absorbed via a polar route. Despite a threefold range in follicle area in the marmoset, no differences in absorption rates of paraquat, mannitol, water, and ethanol were observed between different body sites (Scott et aI., 1991). However, among the different species examined in this
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CHAPTER 22
Pesticide Disposition: Dermal Absorption
study, there was an 80-times range in follicle area, which correlated with observed differences in the rates of mannitol and paraquat absorption. The authors concluded that this correlation was only possible with relatively slowly absorbed test penetrants such as paraquat and mannitol. Further work is needed to determine the extent to which the unique anatomical features at different body sites play a role in absorption and penetration of both lipophilic and hydrophilic pesticides. 22.4.2 AGE-RELATED DIFFERENCES
Very little information is available on age-related absorption of pesticides. The generally accepted theory is that dermal bioavailability should be greater in the young than in adults because of the underdeveloped skin barrier in the young. Differences in dermal absorption between young and adult rat skin has been demonstrated for 11 out of 14 pesticides (Shah et aI., 1987a). Dermal absorption of four pesticides (atrazine, carbaryl, chlorpyrifos, and hexachloro-biphenyl) was greater in the young, whereas dermal absorption of seven pesticides [carbofuran, captan, dinoseb, disodium methanearsonate (DSMA), monosodium methanearsonate (MSMA), nicotine, and parathion] was greater in adults. In vitro studies with human skin further demonstrated that triclocarban absorption was greater in newborn foreskin (2.5%) than in adult abdominal skin (0.23%; Wester and Maibach, 1985). The latter comparison is confounded by anatomical site differences. Another study clearly indicated greater penetration of carbofuran in young rats as measured by both in vivo and in vitro methods (Shah et aI., 1987b). Carbofuran absorption in young rats was 34% for flow-through systems, 12% for static systems, and 34% for in vivo systems, whereas in older rats, absorption was 11.2, 8.8, and 11.8%, respectively. Although there were some differences between the systems used, there are clear agerelated differences for absorption of this carbamate insecticide. Lack of proper mixing and lipophilicity of carbofuran probably explain why absorption was low in the static cell system. In another in vitro study, carbofuran absorption through human newborn foreskin was 78% in 24 hours (Shehata-Karam et aI., 1988), which is greater than in previously described rodent models. This is not surprising because a low dose was used and dose-dependent effects on carbofuran absorption have been reported (Shah et al., 1987a, b). Dermal absorption studies with dinoseb exposure in rats have, however, demonstrated the opposite relationship, with absorption being greater in adults than in the young (Hall et al., 1992). In adult rats, 85% was absorbed with a half-life of 4.9 hours in one phase of the biphasic absorption process and the remaining 15% with a half-life of 34 hours in the second phase. In young rats, 59.7% was absorbed with half-life of 5.4 hours in the first phase and the remaining 40% with a half-life of 298 hours. Dermal absorption of captan was similar in adult (9.8%) and young (8.4%) rats in vivo after a 72-hour exposure (Fisher et al., 1992). Penetration also increased as applied dose decreased. However, in vitro studies utilizing static cells and flow-though diffusion cells provide conflicting results.
The static system gave penetration values twice those obtained for in vivo, while the flow-through diffusion system overestimated in vivo absorption by 43% in young and underestimated in vivo absorption in adults by 19%. This underscores the need to use a validated model prior to making such comparisons. 22.4.3 VARIATION BETWEEN AND WITHIN SPECIES
The use of laboratory animal models can sometimes lead to overprediction of pesticide absorption in human skin. Differences in permeability properties between human skin and those of laboratory animal skin can account for this overprediction. Furthermore, inherent structural differences in skin biology (e.g., skin thickness, sebaceous secretions) make speciesspecies extrapolation of dermal absorption data very difficult. These differences in epidermal and dermal anatomy and physiology have been well documented, although the basic architecture of skin is similar in all terrestrial mammals (Monteiro-Riviere, 1991). It is plausible that a high density of hair follicles attenuates the thickness of the interfollicular epidermis, which may promote absorption. Additionally, species differences in stratum corneum lipid composition may be the overriding factor in determining the rate and extent of absorption. Because of these most glaring differences that distinguish rodent skin from human skin, skin from rodents may not always serve as a suitable model. However, pig skin is functionally and structurally similar to that of human skin (Monteiro-Riviere, 1991) and, therefore, percutaneous absorption of toxicants through pig skin should mimic absorption through human skin. Recent studies have demonstrated that the range of percutaneous absorption of carbaryl, lindane, malathion, and parathion in pig skin in vivo (Carver and Riviere, 1989) or in vitro (Chang et aI., 1994b) is similar to that observed in humans (Feldmann and Maibach, 1974). However, this was not the case when the permeability of hydrophilic chemicals (mannitol, water, and paraquat) and lipophilic chemicals (carbaryl, aldrin, and fluazifop-butyl) in pig ear skin was compared with human abdominal skin and rat dorsal skin (Dick and Scott, 1992). This study demonstrated that for hydrophilic chemicals, pig ear skin and rat skin overestimated permeability in human skin. Although the permeability was generally higher in animal skin than in human skin for the lipophilic chemicals, the permeability of carbaryl in human and pig skin was almost identical. The permeability of lipophilic chemicals in pig skin correlated better with data from human skin compared to the permeability of hydrophilic chemicals. Bartek et al. (1972) demonstrated good agreement between human in vivo and pig in vivo dermal absorption data for lipophilic chemicals, such as butter yellow, DDT, haloprogin, lindane, and testosterone, and hydrophilic chemicals, such as caffeine, acetylcysteine, and cortisone. Further studies with nonhuman primates observed that lindane absorption from rhesus monkey forearm (18%) was about twice that for the ventral forearm in humans (9.3%) (Feldmann and Maibach, 1974) and for the ventral abdomen in pigs (7.7%) (Chang et aI., 1994b).
22.4 Factors that Affect Dermal Absorption
Finally, note that these species comparative data do not prove any specific anatomical route of penetration, only that there is a species difference. In summary, the rate of absorption of most chemicals can be ranked in the decreasing order rabbits > rodents > weanling pig
= rhesus monkey::::: human.
22.4.4 PHYSICOCHEMICAL PROPERTIES AND DOSE EFFECTS 22.4.4.1 Dose Effects Limited information is available in the literature regarding the effects of dose or multiple dose exposure on percutaneous absorption of pesticides. Most studies have demonstrated that absorption efficiency tends to decrease significantly as topical dose increases in simple mixtures. This makes predicting pesticide bioavailability very difficult. This decrease in absorption efficiency has been demonstrated with drugs (e.g., cortisone, testosterone, benzoic acid) in human and rhesus monkey skin (Scheuplein and Ross, 1974; Wester and Maibach, 1976), with lindane in humans (Maibach and Feldmann, 1974), and with parathion in pigs in vitro (Chang and Riviere, 1991). Other studies by Shah et al. (1987a) demonstrated that as the topical dose of 14 pesticides increased, the median penetration value decreased in both young and adult rats in vivo. Of the 14 pesticides tested, a highly significant effect of dose on skin permeability was observed with carbaryl, captan, folpet, and polychlorobiphenyl (PCB). More specifically, skin penetration expressed as a percentage of recovered carbaryl dose ranged from 30-37,12-20, and 4-5% in skin dosed with 31-37, 108, and 539-l-l-g/cm2 of carbaryl, respectively. Insecticides with skin penetration values higher than those for carbaryl were permethrin (16-57%), chlorpyrifos (59-90%), and parathion (58-82%). The range of skin penetration for these other three pesticides was associated with application of low, medium, and high doses. Note that this dose-dependent effect on absorption has not been demonstrated in human skin treated with parathion (Maibach and Feldmann, 1974) nor in rats treated with methylated arsenicals (Shah et aI., 1987a) and dinoseb (Hall et aI., 1992). Repeated or multiple exposure to malathion (15 I-l-g!cm2) for 15 days does not appear to alter absorption in guinea pigs (Bucks et aI., 1985; Courtheoux et aI., 1986) or in humans (Wester et aI., 1983). Therefore, malathion, does not appear to induce changes in barrier function as observed with multiple exposure to drugs such as hydrocortisone and salicylic acid (Roberts and Harlock, 1978; Wester et at., 1980b). However, be aware that some experimental protocols or real world occupational scenarios involve daily soap and water treatment prior to daily topical doses that can significantly decrease barrier function and can increase absorption (Wester and Maibach, 1989). One human in vivo 5-day study demonstrated significant dermal absorption (14%) of 2,4-D-amine, despite rigorous washing at 24 hours posttreatment (Moody et aI., 1992). However, only 0.5% absorption occurred after washing at 1 minute posttreatment.
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22.4.4.2 Pesticide Chemistry Physicochemical factors, such as molecular weight and structure, lipophilicity, pKa, ionization, solubility, partition coefficients, and diffusivity, can influence the dermal absorption of various classes of pesticides. In addition, the penetration of acidic and basic pesticides is influenced by the skin surface, which is weakly acidic (pH 4.2-5.6). Although several of these factors (e.g., molecular weight and partition coefficients) have been used to predict absorption of various drug classes (Bunge and Cleek, 1995; Cleek and Bunge, 1993; Potts and Guy, 1992), this approach has not been applied to pesticides. It is expected that these physicochemical factors influence the rate and extent of absorption for various classes of pesticides. Paraquat and diquat are hydrophilic pesticides that exist as fixed charged cations and remain dissociated at all pH values. Very little paraquat or diquat is, therefore, expected to be absorbed by skin, although percutaneous absorption of paraquat has resulted in systemic effects and deaths in humans (Smith, 1988). Dermal absorption studies in human volunteers demonstrated 0.29, 0.23, and 0.29% absorption in the leg, forearm, and forearm, respectively (Wester et aI., 1984). One in vivo study in rats, demonstrated that paraquat absorption (3.5%) was greater in rats because an occlusive dressing was used and also because of differences in skin thickness between species (Chui et al., 1988). Topical application of 3-,24-, and 200-mg doses of paraquat to IPPSFs for 8 hours resulted in penetration (skin and perfusate) of 0.91, 1.09, and 0.50%, respectively (Srikrishna et aI., 1992). These absorption data are comparable to the human in vivo data. Despite the limited amounts absorbed, they were sufficient to cause morphological and biochemical changes in the IPPSFs. Other studies have determined that the in vitro permeability constants for paraquat in various animal species (rat, hairless rat, nude rat, mouse, hairless mouse, rabbit, guinea pig) are 40-1600 times greater than for humans (Walker et aI., 1983). Like paraquat, very little diquat is absorbed (0.3%) in the human forearm in vivo (Maibach and Feldmann, 1974). Diquat absorption increased to 1.4% with occlusion and to 3.8% with damaged skin. Data from these in vivo and IPPSF studies suggest that paraquat- or diquatinduced dermatotoxicity is a highly probable mechanism, a priori, for dermal absorption of these hydrophilic and charged pesticides. The various forms of organic arsenicals used as herbicides include dimethylarsinic acid (DMA) and the sodium salts of methylarsinic acid, MS MA and DSMA. Because of the carcinogenic potential of these herbicides, there have been no human dermal absorption studies. Many of the commercial MSMA and DMSA formulations contain various surfactants [U.S. Environmental Protection Agency, (EPA), 1975], that alter skin barrier properties and may promote permeation of the arsenicals. Skin penetration of MSMA and DSMA in mice can range from 2 to 22 and 1.2 to 15% of dose, respectively, within 72 hours (Hall et al., 1989; Shah et aI., 1987a). Surprisingly, absorption was greater in adult mice than young mice for both arsenicals and there was no dose effect for either of these arsenicals. In similar studies, a constant fraction of the dose (12.4%)
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penetrated mice skin within 24 hours in vitro from aqueous vehicles over an wide dosage range (Rahman and Hughes, 1994). Of this amount, only 4% were absorbed into the receptor fluid. In the same study, short-term exposure (1 hour) resulted in 0.63% absorption in skin and 0.03% absorption in receptor fluid. For the remaining 23-hour perfusion, 0.54% of the dose remained in the skin, whereas 0.13% was still absorbed. Although the pH levels of MSMA and DSMA were at least 1.18-1.7 units higher than the pKa of these arsenica1s, indicating 93-98% of the arsenicals were ionized (mono- and dibasic forms), ionization had no effect on absorption though mouse skin. Topical application of aqueous solutions (20, 100, and 250 !-LI) of 10 !-Lg ofDMA in mice skin in vitro resulted in 5.1625.22% dose in receptor fluid and 2-16% dose in skin tissue within 24 hours (Hughes et al., 1995). This study also determined that lag times were about 1 hour, but within 4 hours, 50% of the total (24 hour) cumulative dose detected in the receptor fluid had penetrated the skin at doses ranging from 10 to 500 !-Lg. As demonstrated in previous studies with MSMA and DSMA, short-term exposure to DMA (1 hour) resulted in about 1% absorption, and the percentage of dose in perfusate and skin was unaffected by changes in applied dose or pH of the dosing solution. Many of the chlorinated hydrocarbon insecticides have a known tendency to accumulate in adipose tissue. Because of the high lipophilicity of these compounds compared to organophosphates, partitioning of absorbed compound from dermis to subcutaneous adipose tissue ultimately affects absorption flux profile and tissue distribution. For example, parathion and lindane have similar molecular weights, but the difference in partition coefficients is a very plausible explanation for why measured and predicted absorption of lindane was demonstrated experimentally to be less than that for parathion (Chang et aI., 1994b). In a 24-hour in vivo study in rabbits, greater quantities of DDT and dieldrin were retained on the skin surface, skin, and adipose tissue compared to carbaryl, parathion, and malathion, which are less lipophilic than the organoch10rines (Shah and Guthrie, 1977). Not surprisingly, blood concentration was significantly lower at 8 and 24 hours for the organochlorines. Carbaryl is the most studied of all carbamate pesticides and despite its low toxicity, appears to penetrate human and animal skin more readily than most other pesticides. This increased permeability compared to most other pesticides is most likely associated with its unique physicochemical characteristics. Almost complete penetration of carbaryl was observed when carbaryl was dissolved in acetone and applied to the forearm and jaw angle of six human volunteers (Feldmann and Maibach, 1974; Maibach et al., 1971). The data from these studies demonstrated that 74% of the 24-hour applied dose (4 !-Lg/cm2) was excreted in urine over 5 days. Utilizing deconvolution analysis of the same human data (skin to urine and blood to urine data), cumulative absorption over 5 days was estimated to be 63% of the applied dose, with 45% of this occurring 8 hours after onset of penetration (Fisher et aI.,
1985). When this analysis was performed with 11 other pesticides (e.g., parathion, aldrin, diquat), more of the carbaryl dose was absorbed within 120 hours compared to other pesticides (0.3-20%). Only carbaryl had a lag time (3.5 hours), which was followed by rapid absorption. About 50% of the 120-hour total absorption of parathion and dieldrin occurred in the first 4 hours. Therefore, although the absorption rate appeared to be less (due to the 3.5-hour lag time) with carbaryl compared to parathion in humans, the extent of absorption during 120 hours was greater for carbaryl. This 1ag period may be unique to carbaryl, although it can be formulation dependent or related to body clearance. Although dermal absorption usually follows Fick's law of diffusion, other studies have demonstrated that dermal absorption for carbaryl as well as dinoseb is a biphasic process. This phenomenon is probably related to the physicochemical properties of the penetrant (Hall et aI., 1992; Shah and Guthrie, 1983). 22.4.5 PESTICIDE FORMULATION AND MIXTURES
Insecticide efficacy, the stability of active ingredients, and programmed release of active ingredients from the vehicle/device are the most important characteristics controlled for when pesticides are formulated (Krenek and Rohde, 1988). EPA registration does not always require percutaneous absorption studies. For this reason, more efficacy data than dermal pharmacokinetic data are available in the literature. Furthermore, most of the available pesticide absorption data pertain to binary mixtures (pesticide + vehicle). Technical grade formulations are, however, complex mixtures of formulation additives and, therefore, risk assessment based on data from exposure to binary mixtures may be inappropriate. Pesticides are usually formulated to contain active and inactive or inert ingredients. The latter component(s) can enhance the rate and extent of absorption or slow the release of the active ingredient and thus reduce the rate and extent of absorption (Walters and Roberts, 1993). These "inert" ingredients are often classified as adjuvants, surfactants, preservatives, solvents, diluents, thickeners, and stabilizers. These pesticide additives were first covered by the Food and Drug Administration and now are covered by EPA regulation 40 CFR 180.1001 and also TSCA and FIFRA (Seaman, 1990). This increasing list of inerts as well as the prohibitive cost to obtain 40 CRF 180.1001 clearance of new inerts strongly support the need to evaluate the influence of current and novel inerts on the toxicology and dermal absorption of active ingredients in pesticide formulations. Several studies have demonstrated the penetration enhancing ability of acetone compared to water, ethanol, or other vehicles commonly used in dermal absorption studies. Early work by O'Brien and Dannelley (1965) showed that in comparison with benzene and corn oil, acetone was best at enhancing carbaryl absorption. More recent studies also have demonstrated the enhancing effect of acetone compared with other solvent systems on the absorption of carbaryl, p-nitrophenol, and 2,4-D (Baynes and Riviere, 1998; Brooks and Riviere, 1995; Moody et aI., 1992).
22.4 Factors that Affect Dennal Absorption
However, other studies have demonstrated that commercial formulations are more effective than acetone in enhancing pesticide absorption. Methyl parathion absorption in vitro in human skin at 24 hours was 1.3% in acetone, but was significantly increased to 5.2% in a commercial formulation (Sartorelli et aI., 1997). Likewise, in vivo dermal exposure studies of lindane in humans resulted in approximately 60% with a white spirits formulation and 5% with an acetone vehicle (Dick et aI., 1997a, b). In these latter experiments, more of the lindane dose (79%) remained on the skin surface at 6 hours with acetone than with the white spirits formulation (10.5%), and significant levels of lindane accumulated in the stratum corneum with white spirits (30%) and with acetone (14.3%) at 6 hours. These findings strongly suggest that the white spirits formulation enhanced lindane penetration with respect to acetone vehicle. The in vitro studies with human skin also demonstrated a similar pattern although only 18 and 0.3% of the dose was absorbed into the perfusate at 6 hours for the white spirits formulation and the acetone vehicle, respectively. Topical application of 1% commerciallotion of lindane in vitro in human and guinea pig skin resulted in absorption levels as high as 71.72 and 35.31 %, respectively, at 48-hour exposure (Franz et aI., 1996). Dermal absorption of alachlor as an emulsifiable concentrate and microencapsulated formulation was demonstrated to be 8.5 and 3.8%, respectively, in rhesus monkeys after a 12-hourexposure (Kronenberg et aI., 1988). About 88% of the systemically absorbed dose were excreted in urine within 48 hours. However, the differences between these two formulations were not statistically significant. Although dilution of either of these formulations (1 :29) slightly enhanced alachlor absorption, these effects were surprisingly not statistically significant. One in vitro study with human skin demonstrated similar absorption data (0.54%) after an 8-hour exposure and peak fluxes within 3-5 hours postapplication (Bucks et aI., 1989a). However, a significant effect of formulation dilution with water was observed in this study, even though the same mass of alachlor was applied to skin. Not surprisingly, a greater fraction of alachlor was present on the skin surface and skin tissue than in the receptor fluid, and the high capacity for stratum corneum binding demonstrated in this study is not unique for related chlorinated aromatic chemicals. Data from several studies have suggested that the insect repellent, DEET, enhances transdermal delivery of drugs and toxicants (Moody et aI., 1987; Windheuser et aI., 1982). Some studies have demonstrated that DEET can act as a transdermal accelerant of2,4-D-amine (Moody et aI., 1992). Recent studies in our laboratory have, however, determined that DEET blocked permethrin absorption and inhibited carbaryl absorption in acetone, but not in dimethyl sulfoxide DMSO mixtures (Baynes et al., 1997; Baynes and Riviere, 1998). The insecticide synergist, piperonyl butoxide, was also shown to enhance carbaryl absorption (Baynes and Riviere, 1998). These diffusion studies further demonstrated that piperonyl butoxide does not enhance absorption through inert latex membranes, but does so in porcine skin sections. This observation suggests that some chemical-biological interaction or other mechanisms (e.g., ir-
523
ritation) may occur in skin to enhance the absorption of pesticides. An expected, but important finding in these carbaryl experiments was that increased dilution of the carbaryl formulation with water, especially in the presence of the surfactant, sodium lauryl sulfate (SLS), enhanced carbaryl absorption. The penetration enhancing effect of SLS also was observed with parathion (Qiao et al., 1996). In addition to the formulation additives, these agrochemicals may contain isomers, homo logs, or breakdown products that form after synthesis and/or formulation and during storage (Chambers and Dorough, 1994). Although these impurities can potentially alter the toxicity and toxicokinetics of the pesticide, many toxicology and dermal absorption studies have ignored these impurities and used the pure rather than the technical grade pesticide. There is evidence that technical grade malathion can be more lethal (eightfold difference) in rats than the purified form (Umetsu et aI., 1977). Other studies have demonstrated that organophosphates such as malathion and fenitrothion can potentiate the toxicity of the carbamate insecticide carbaryl (Takahashi et aI., 1987). Previous metabolism studies in the IPPSF (Carver et aI., 1990) demonstrated a significant first-pass metabolism of parathion to p-nitrophenol and paraoxon, and that these metabolites may be present simultaneously during absorption of parathion. Environmental exposure to parathion is never to pure parathion because spontaneous degradation occurs during storage. When mixtures of parathion and its metabolites were dosed and then assayed for parathion and its two metabolites across pig skin in vitro, significant interactions were detected. In general, the nontoxic metabolites p-nitrophenol and I-naphthol can significantly enhance the absorption of the parent compounds, parathion and carbaryl, respectively (Baynes and Riviere, 1998; Chang et aI., 1994a). Surprisingly, p-nitrophenol did not enhance the absorption of paraoxon; this toxic metabolite of parathion and parathion appears to decrease the absorption of p-nitrophenol and paraoxon. In other related absorption studies, pretreatment with 3% fenvalerate decreased subsequent absorption of parathion, increased subsequent lindane absorption, and had no effect on subsequent fenvalerate or carbaryl absorption (Chang et aI., 1995). These results underscore the chemical specificity of these interactions and reinforce the concept that the percutaneous absorption of a mixture cannot be predicted from individual component studies. These data suggest that other mechanisms in addition to vehicle and surfactant effects must be operating simultaneously; hence further investigation is required. The data reinforce the concept that the permeability of a mixture cannot be predicted from individual component studies. Many of the mechanisms of pesticide mixtures interactions are not well understood and are not easy to model, although a biophysical model for parathion was attempted (Williams et aI., 1996). It should also be recognized that it is more often the formulation additives and other environmental factors rather than the active ingredient that compromise the skin barrier and eventually enhance pesticide absorption. There is epidemiological evidence that agricultural pesticides can cause dermatoses (Abrams et aI., 1991;
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Cellini and Offidani, 1994; Guo et al., 1996) and there is experimental evidence that UV irradiation can enhance skin reactions to topical agricultural chemical treatment (Kimura et al., 1998). In the latter study, significant reactions were observed for several herbicides. Maibach and Feldmann (1974) demonstrated that dermal absorption of pesticides such as parathion, azodrin, and diquat occurs more readily (ninefold) through damaged skin than through normal skin. It is, therefore, plausible to assume that the formulation additive can inflict local reversible or irreversible "damage" to the skin structure and physiology, and that it is these interactions that modulate dermal absorption of most pesticides. 22.4.6 ENVIRONMENTAL FACTORS 22.4.6.1 Temperature
Changes in ambient air temperature can alter lipid fluidity in the intercellular lipid domain of the stratum corneum. This alteration in the intercellular pathway can theoretically alter pesticide penetration through the stratum corneum. Previous in vivo studies have demonstrated that increased percutaneous absorption of a cholinesterase inhibitor (VX) was a function of skin temperature (Craig et aI., 1977). In humans topically exposed to parathion at different ambient temperatures (11, 25, and 40°C), the urinary excretion of the metabolite p-nitrophenol paralleled the increase in ambient temperature (Hayes et al., 1964). Sev~ral in vitro experiments with pig skin also demonstrated that increasing air temperature from 37 to 42°C significantly increased parathion absorption (Chang and Riviere, 1991). Increased ambient temperatures can also increase the evaporation of volatile pesticides from the skin, thereby reducing the topical dose available for absorption. Increasing air flow over the skin increases evaporative loss and significantly decreases dermal residues in the upper skin layer of pigs for DDT, malathion, parathion, and DEET (Reifenrath et al., 1991). Wester et al. (1992a) demonstrated that isophenfos concentrations on the human skin surface in vivo was less than 1% dose at 24 hours and that evaporation from the skin surface during absorption reduced the dose available for penetration and absorption. Finally, it should be recognized that skin surface conditions in vitro are more easily controlled than in vivo, and data from in vitro studies can significantly underestimate evaporation in vivo. 22.4.6.2 Humidity and Occlusion
Skin hydration can be increased by occlusion, with high relative humidity or immersion conditions (e.g., swimming or bathing). Although previously it was assumed that hydration changes only affect dermal absorption of polar compounds, there is significant data that suggest that at high relative humidity, this hydration effect becomes more important for nonpolar molecules such as pesticides and is most likely secondary to an increase in diffusivity of the penetrating molecule (Behl et aI., 1980). Under relative humidity conditions greater than 80%, parathion
absorption was significantly increased in pig skin in vitro by as much as 2-3 times the value under standard conditions of 60% relative humidity (Chang and Riviere, 1991, 1993). The practical application of occlusion is when pesticides get into and under the clothing of workers and form the ideal reservoir for penetration and absorption into the skin. Occlusion can change dermal absorption by various mechanisms, among which are reducing loss of evaporation from the skin surface, enhanced skin hydration, changes in cutaneous metabolism, dermal irritation, and altered cutaneous blood circulation (e.g., vasodilation). Occlusion can increase hydration of the stratum corneum from as little as 5-15% to as much as 50% (Bucks et aI., 1989b), thereby modulating the absorption profile for the pesticide. One in vivo study with pigs (Qiao et aI., 1997) demonstrated that occlusion significantly enhanced pentachlorophenol (PCP) absorption from 29.1 to 100.72% dose and changed the shape of the absorption profile in blood and plasma. The study also suggested that occlusion changed the local metabolism of PCP and as a result, the 14C partitioning between plasma and red blood cells. Occlusion was also kinetically related to modification of cutaneous biotransformation of topical parathion (Qiao and Riviere, 1995). Occlusion enhanced the cutaneous metabolism of parathion to paraoxon and to p-nitrophenol as well as the percutaneous absorption and penetration of both parathion and p-nitrophenol. Occlusion also reduced parathion and p-nitrophenollevels in the skin, but increased p-nitrophenol and p-nitrophenol-glucuronide in the blood. Other in vivo studies (Qiao et al., 1993) showed that dermal occlusion significantly enhanced the rate and extent of parathion absorption in pigs in the abdomen (43.94 vs. 7.47%), buttocks (48.47 vs. 15.60%), back (48.82 vs. 25.00%), and shoulder (29.28 vs. 17.41 %). Although significant anatomical site differences were observed with non occluded skin, these site differences were concealed with occluded skin. In vitro studies with parathion also demonstrated that occlusion increased absorption from 0.46-7.69 to 1.04-17.46% at doses ranging from 4 to 400 ).lg/cm2 (Chang and Riviere, 1993). Pesticides can be transferred from cotton fabric into and through human skin as demonstrated in several studies (Snodgrass, 1992; Wester et aI., 1996a), but it should be recognized that these studies were under occlusive conditions. Dermal absorption of malathion was 3.92% with ethanol wet fabric and 0.6% with 2-day-treated cotton sheets (Wester et aI., 1996a). However, malathion absorption was increased to 7.34% when the 2-day-treatedldried cotton fabric was wetted with aqueous ethanol. In the same study, absorption of glyphosphate was 1.42% in water solution, 0.74% when applied as wet cotton sheets, and 0.08% when applied as 2-day-treatedldried cotton sheets. Absorption increased to 0.36% when the 2-daytreatedldried cotton sheets were wetted with water to simulate sweating and wet conditions. Military uniforms are impregnated with permethrin as a defense against nuisance and disease-bearing insects. Application of fabric impregnated with permethrin to the backs of rabbits resulted in a 3.2% migration to the skin surface with 2% of the impregnant being absorbed
22.4 Factors that Affect Dennal Absorption
and 1.2% remaining on the skin surface after 7 days of continuous skin contact (Snodgrass, 1992). The implications of these interactions, especially for agricultural workers during pesticide application in humid climates or for military personnel under combat conditions in the desert, should not be underestimated.
22.4.6.3 Soil Pesticide adsorption to soil can alter the amount of pesticide available for dermal absorption. It should also be recognized that exposure conditions such as exposure time, pesticide concentration, soil load, and soil characteristics are important variables that can theoretically influence absorption (Bunge and Parks, 1997). Soil adherence to skin, for instance, can vary from 10- 3 to 102 mg/cm 2 and has been shown to be activity dependent (Kissel et al., 1996). Predicting dermal absorption of pesticides from contaminated soils is, therefore, not a simple process and becomes problematic because there are very few studies that have addressed many of these issues. For several pesticides (e.g., pep, 2,4-D, chlordane), percutaneous absorption in acetone vehicle appears to be slightly less or not significantly different from absorption from soil. However, for several other pesticides (e.g., DDT, organic arsenicals), soil appears to reduce percutaneous absorption of the pesticide. The interactions between soil and several of these pesticides are subsequently described in more detail, but note that in vitro skin models are, in general, not very predictive of in vivo absorption when exploring these interactions (Wester and Maibach, 1998). Although DDT is no longer widely used in the United States, residues in soil are still detectable and human contact with contaminated soil can result in DDT exposure. One study demonstrated that in vivo absorption of DDT in rhesus monkeys was significantly less from soil (3.3% dose) than from acetone vehicle (18.9%; Wester et aI., 1990). The absorption of DDT in acetone in rhesus monkey is not significantly different from DDT absorption in man (10.4% dose; Feldmann and Maibach, 1974). In vivo absorption from acetone or soil was not similar to in vitro absorption ( <0.1 %). However, in vitro experiments demonstrated that 18.1 % penetrated skin with acetone and 1.0% penetrated skin with soil. Less than 1% dose partitioned into the receptor phase, demonstrating that the skin barrier in addition to the soil is rate limiting and that in vitro skin models may not be useful for predicting DDT absorption in vivo. Unfortunately, only in vitro dermal absorption studies are available for organic arsenicals. One study demonstrated that as much as 12.4% dose of MSMA and DSMA penetrated mice skin within 24 hours from aqueous vehicles over an wide dosage range (Rahman and Hughes, 1994). Of this amount, only 4% were absorbed into the receptor fluid. In the presence of soil (690 ppm), penetration through mice skin was reduced to not more than 0.48 and 0.22% for MSMA and DSMA, respectively. Increasing MSMA and DSMA levels in soil from 690 to 6900 ppm increased skin content, but decreased the percentage of applied dose in skin. Whereas absorption into receptor fluid was very low for MSMA (0.01 %), it was not detectable for DSMA. Topical application of aqueous solutions (20, 100,
525
and 250 J.l.I) of 10 J.l.g of DMA to mice skin resulted in 5.1625.22% dose in receptor fluid and 1.95-15.67% dose in skin tissue within 24 hours (Hughes et aI., 1995). However, when DMA (690 ppm) was applied with soil, absorption was reduced to 0.08% in the receptor fluid and 0.45% in skin. The influence of soil was, however, not observed with inorganic arsenic. In vivo percutaneous absorption of arsenic as H3AS04 in water in rhesus monkeys (2.0-6.4%) was somewhat comparable to in vitro absorption (1.9%) in human skin (Wester et aI., 1993b). However, the soil vehicle did not influence absorption in rhesus monkeys (3.2-4.5%) or human skin in vitro (0.8%), although absorption in both skin models is not comparable. The relative similarities in partition coefficient of arsenic from water to stratum corneum and from water to soil probably explain why absorption from water was similar to absorption from soil. Interactions between soil and the phenoxy herbicides (e.g., 2,4-D acid, 2,4-D amine) are unique. One study demonstrated that dermal absorption of the herbicide, 2,4-D acid, is nonlinear with respect to soil load or skin contact time (Wester et aI., 1996b). Percutaneous absorption in acetone vehicle (8.6%) was not different from absorption of soil loads of 1 mg/cm2 (8.6%) and 40 mg/cm2 (15.9%) in rhesus monkey in vivo. Further in vitro experiments with human skin demonstrated that increasing the soil load from 5 to 40 mg/cm 2 did not affect 2,4D absorption, which ranged from 1.4 to 1.8%. During the first 24 hours of in vitro exposure, absorption was linear with respect to time for an acetone vehicle (3.2%); however, there was an apparent lag time of about 8 hours with absorption from a soil vehicle (0.03-0.05%). This early lag time may be related to chemical partitioning from soil and may be beneficial if the skin is decontaminated within 24 hours. The investigators proposed that because of complex interactive forces between pesticides and soil, dermal absorption calculations based on assumed linearity can incorrectly estimate the threat to human health. Mathematical extrapolation from high soil loads to low soil loads may significantly underestimate 2,4-D absorption. These studies also demonstrated that soil release kinetics may limit dermal absorption and that more data are needed to make valid predictions. In contrast to DDT, chlordane absorption in rhesus monkeys in acetone (6.0% dose) was similar to absorption in soil (4.2% dose) 6 days after exposure (Wester et aI., 1992b). Although human skin in vitro experiments demonstrated similar partitioning into receptor fluid for acetone (0.07%) and soil vehicles (0.04%), there was greater penetration into skin with acetone (10.8%) than with soil (0.34% dose) at 24 hours. It is possible that chlordane adsorption to soil delayed percutaneous absorption during the initial 24 hours and an extrapolation to 6 days would reveal no vehicle differences as demonstrated in the in vivo study. The octanol:water partitioning coefficients (log P) of chlordane and DDT are 5.58 and 6.91, respectively, and, therefore, dermal disposition should be similar. The high lipophilicity of these pesticides explains the higher proportion of pesticide in the skin than in the receptor phase, but it does not explain why the differences between acetone and soil for
526
CHAPTER 22
Pesticide Disposition: Dermal Absorption
DDT are greater than those for chlordane; it only suggests that factors other than lipophilicity influence absorption of these organochlorines. Various studies have demonstrated that the very ubiquitous pesticide, PCP, is very readily absorbed though human, monkey, and pig skin (Qiao et aI., 1997; Wester et aI., 1993a). In vivo absorption of PCP in rhesus monkeys with acetone vehicle (29.2%) was similar to absorption with soil vehicle (24.4%) after a 24-hour exposure period (Wester et aI., 1993a). However, in vitro absorption with human skin appears to underestimate in vivo absorption because only 0.6-1.5 and 0.01-0.07% dose were detected in receptor fluid with acetone and soil vehicle, respectively, at 24 hours. Skin concentrations were only 2.63.7 and 0.11-0.14% for acetone and soil vehicles, respectively, which is still not comparable to the in vivo data. The approximately 25% PCP absorption from nonocclusive soil in monkeys (Wester et aI., 1993a) compares favorably with the 29% PCP absorption from non occlusive soil in Yorkshire pigs (29.1 %) in vivo (Qiao et aI., 1997). Note that inhibition of soil and/or skin microorganisms can inhibit absorption of PCP, alter local and systemic distribution, and increased plasmalblood concentration ratios in pig skin in vivo (Qiao et aI., 1997). It is plausible to assume that skin or soil microorganisms and/or products of PCP microbial degradation may play a role in PCP absorption and disposition. 22.4.7 CUTANEOUS METABOLISM It is well known that monooxygenase activity exists in basement cells of the epidermis (Bicker et aI., 1982; Rettie et aI., 1986)
and that these activities are less than or equivalent to those in the liver if the epidermis and not the whole skin is considered (Noonan and Wester, 1987). Consideration of skin metabolism, however, becomes important when the product of metabolism pesticide is a toxic metabolite or when there is no metabolism or detoxification of the toxic pesticide to harmless metabolites as seen with oral routes of exposure. For example, the only barrier to hexachlorophene toxicity by the dermal route is the stratum corneum, whereas by the oral route, there is first-pass metabolic detoxification by the liver after oral administration (Wester and Maibach, 1985). Of further importance, pesticide metabolites of parathion and carbaryl can significantly increase absorption of the parent pesticide (Baynes and Riviere, 1998; Chang et aI., 1994a) as stated earlier. Previous studies in our laboratory have demonstrated significant first-pass bioactivation of parathion to paraoxon (68%) and p-nitrophenol (15%) in IPPSFs (Carver et aI., 1990). Pretreatment with a P450 suicide substrate, l-aminobenzotriazole, blocked 90% of the paraoxon production observed in the control IPPSFs. Carbaryl can be metabolized by hydrolysis and ring hydroxylation followed by conjugation to glucuronide or sulfate with liver postmitochondrial fraction (Macpherson et aI., 1991a). However, when skin postmitochondrial fraction was exposed to carbaryl, only hydrolysis and conjugation products were detected. No naphthol or naphthol conjugates were detected after topical application of carbaryl to rat whole skin in vitro.
The investigators proposed that insufficient chemical came in contact with enzymes for metabolism to be detected. However, other studies demonstrated that a significant percentage of the carbaryl dose (23%) was converted to I-naphthol in isolated perfused porcine skin flaps (Chang et aI., 1994b). Studies with another carbamate, propoxur, demonstrated that pig, rabbit, and human skin tissue was capable of phase I and phase 11 metabolism of this pesticide (Van de Sandt et aI., 1993). Conjugation of the metabolite 2-isopropoxyphenol was, however, species specific, with glucuronides and sulfates, sulfates only, and glucuronides being the major conjugated metabolites detected in porcine, rabbit, and human skin cultures, respectively. Mouse skin microsomes can metabolize triazine herbicides (atrazine, and simazine), parathion, phorate, and aldrin (Venkatesh et aI., 1992). The same study demonstrated that 66-69% of phorate sulfoxidation activity was due to flavincontaining monooxygenase and the remainder was due to P450. Ademola et al. (l993a) further demonstrated that atrazine was metabolized by N-dealkylation to harmless metabolites during percutaneous absorption and that the formation of metabolites decreased with increasing dose. This decrease in conversion may reflect saturation of skin metabolism enzymes or decreased access to metabolism enzymes during absorption. This study also demonstrated microsomal and as well as cystolic metabolism during transdermal transport, with greater amounts of metabolites present in the skin than in the perfusate. During the penetration and absorption of the chloroacetanilide, butachlor, 2.8 and 6.8% of the dose absorbed by skin were recovered as metabolites in the receptor fluid and skin homogenates, respectively (Ademola et aI., 1993b). In addition to hydroxylated derivatives of butachlor, glutathione and cysteine conjugates were also detected in the same study. Aldrin epoxidation and paraoxon hydrolysis can also occur in skin (Fredrickisson, 1964; Graham et aI., 1987). Other studies demonstrated that aldrin conversion to dieldrin can occur during percutaneous absorption in rat skin in vitro and in vivo (Graham et aI., 1987; Macpherson et aI., 1991b). The more recent study also confirmed that aldrin metabolism was primarily in the epidermis and dermal metabolism was minimal. Other investigators were unable to demonstrate cutaneous metabolism of DDT (Bronaugh et aI., 1989).
REFERENCES Abrams, K., Hogan, D. J., and Maibach, H. 1. (1991). Pesticide-related dermatoses in agricultural workers. Occup. Med. 6, 463-492. Ademola, J. 1., Sedik, L. E., Wester, R. c., and Maibach, H. 1. (1993a). In vitro percutaneous absorption and metabolism of 2-chloro-4-ethylamino6-isopropylamine-s-triazine (Atrazine). Arch. Toxico!. 67, 85-91. Ademola, J. 1., Wester, R. c., and Maibach, H. 1. (I 993b). Absorption and metabolism of 2-chloro-2,6-diethyl-N -(butoxymethyl)acetanilide (Butachlor). Toxico!. Appl. Pharmaco!' 121, 78-86. Andrawes and Bailey (1978). Barry, B. W. (1991). The LPP theory of skin penetration enhancement. In "In Vitro Percutaneous Absorption: Principles, Fundamentals, and Applications" (R. L. Bronough and H. 1. Maibach, eds.), pp. 165-185. CRC Press, Boca Raton, FL.
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CHAPTER 22
Pesticide Disposition: Dennal Absorption
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Shah, P. v., Monroe, R. v., and Guthrie, F. E. (1983). Comparative penetration of insecticides in target and nontarget species. Drug Chem. Toxieo!. 6, 155179. Shah, P. v., Fisher, H. L., Sumler, M. R., Monroe, R. J., Chernoff, N., and Hall, L. L. (1987a). Comparison of the penetration of 14 pesticides through the skin of young and adult rats. J. Toxieo!. Environ. Health 21, 353-366. Shah, P. v., Fisher, H. L., Month, N. J., Sumler, M. R., Chernoff, N., and Hall, L. L. (1987b). Dennal penetration of carbofuran in young and adult Fischer 344 rats. 1. Toxieo!. Environ. Health 22, 207-223. Shehata-Karam, H., Monteiro-Riviere, N. A., and Guthrie, F. E. (1988). In vitro penetration of pesticides through human newborn foreskin. Taxieo!. Left. 40, 233-239. Sidon, E. w., Moody, R. P., and Franklin, C. A. (1988). Percutaneous absorption of cis- and trans-pennethrin in rhesus monkeys and rats: Anatomic site and interspecies variation. Toxieo!. Environ. Health 23, 207-216. Smith, J. G. (1988). Paraquat poisoning by skin absorption: A review. Human Toxieo!. 7, 15-19. Snodgrass, H. L. (1992). Pennethrin transfer from treated cloth to the skin surface: Potential for exposure in humans. J. Toxiea!. Environ. Health 35, 91105. Srikrishna, v., Riviere, J. E., and Monteiro-Riviere, N. A. (1992). Cutaneous toxicity and absorption of paraquat in porcine skin. Taxieo!. App!. Phannaca!. 115,89-97. Takahashi, H., Kato, A., Yamashita, E., Naito, Y., Tsuda, S., and Shirasu, Y. (1987). Potentiations of N-methylcarbamate toxicities by organophosphorous insecticides in male mice. Fundam. App!. Taxieo!. 8, 139-146. Vmetsu, N., Grose, F. H., Allahyari, R., Abu-EI-Haj, S., and Fukuto, T. R. (1977). Effects of impurities on the mammalian toxicity of technical malathion and acephate. J. Agric. Food. Chem. 25, 946-953. V.S. Environmental Protection Agency. (1975). "Initial Scientific Review of Cacodylic Acid." Rep. EPA-540/l-75-021, Office of Pesticide Programs, Washington, DC. Vaden, S. L., Page, R. L., Williams, P. L., and Riviere, J. E. (1994). Effects of hyperthennia on cisplatin and carboplatin in the isolated perfused tumor and skin flap. Int. J. Hyperthermia 10, 563-572. Van de Sandt, J. J., Rutten, A. A., and van Ommen, B. (1993). Species-specific cutaneous biotransfonnation of the pesticide propoxur during percutaneous absorption in vitro. Taxieol. App!. Pharmaea!' 123, 144-150. Venkatesh, K., Levi, P. E., Inman, A. 0., Monteiro-Riviere, N. A., Misra, R., and Hodgson, E. (1992). Enzymatic and immunohistochemical studies on the role of cytochrome P450 and the flavin-containing monooxygenase of mouse skin in the metabolism of pesticides and other xenobiotics. Pest. Bioehem. Physial. 43, 53-66. Walker, M., Dugard, P. H., and Scott, R. C. (1983). In vitro percutaneous absorption studies: A comparison of human and laboratory species. Human Toxieo!. 2,561-568. Walters, K. A., and Roberts, M. S. (1993). Veterinary applications of skin penetration enhancers. In "Pharmaceutical Skin Penetration Enhancement" (K. A. Walters and J. Hadgraft, eds.). Dekker, New York. Wester, R. C., and Maibach, H. I. (1976). Relationship of topical dose and percutaneous absorption in rhesus monkey and man. J. Invest. Dennatal. 67, 518-520. Wester, R. c., and Maibach, H. I. (1985). Dennatopharmacokinetics: A dead membrane or a complex multifunctional viable process. In "Progress in Drug Metabolism" (J. W. Bridges and L. F. Chasseaud, eds.), Vo!. 9, pp. 95109. Taylor and Francis, Washington, DC. Wester, R. C., and Maibach, H. I. (1989). Dennal decontamination and percutaneous absorption. In "Percutaneous Absorption. MechanismsMethodology-Drug Delivery" (R. L. Bronaugh and H. I. Maibach, eds.). Dekker, New York. Wester, R. c., and Maibach, H. I. (1998). Percutaneous absorption of hazardous substances from soil and water. In "Dennal Absorption and Toxicity Assessment" (M. S. Roberts and K. A. Walters, eds.), pp. 697-707. Dekker, New York. Wester, R. C., Noonan, P. K., and Maibach, H. I. (1980a). Variations in percutaneous absorption of testosterone in the rhesus monkey due to anatomic site
530
CHAPTER 22
Pesticide Disposition: Dermal Absorption
application and frequency of application. Arch. Dermato!' Res. 267. 229235. Wester, R. c., Noonan, P. K., and Maibach, H. I. (l980b). Percutaneous absorption of hydrocortisone increases with long-tenn administration. Arch. Derm. Res. 116, 186-188. Wester, R. c., Maibach, H. I., Bucks, D. A w., and Guy, R. H. (1983). Malathion percutaneous absorption after repeated administration to man. Toxieo!. App!. Pharmaeo!' 68,116-119. Wester, R. c., Maibach, H. I., Bucks, D. A w., and Aufrere, M. B. (1984). In vivo percutaneous absorption of paraquat from hand, leg, and forearm of humans. 1. Toxieol. Environ. Health 14,759-762. Wester, R. c., Maibach, H. I., Surinchak, J., and Bucks, D. A. W. (1985). Predictability of in vitro diffusion systems. Effects of skin types and ages on percutaneous absorption of triclocarban. In "Percutaneous Absorption" (R. Bronaugh and H. I. Maibach, eds.), pp. 223-226. Dekker, New York. Wester, R. c., Maibach, H. I., Bucks, D. A. w., Sedik, L., Melendres, J., Liao, c., and DiZio, S. (1990). Percutaneous absorption of [I4C]DDT and benzo[a]pyrene from soil. Fundam. App!. Toxieo!. 15,510-516. Wester, R. C., Maibach, H.I., Melendres, 1., Sedik, L., Knaak, J., and Wang, R. (1992a). In vivo and in vitro percutaneous absorption and skin evaporation ofisofenphos in man. Fundam. App!. Toxieo!. 19,521-526. Wester, R. c., Maibach, H.I., Sedik, L., Melendres, J., Liao, C. L., and DiZio, S. (1992b). Percutaneous absorption of [14C]chlordane from soil. 1. Toxieol. Environ. Health 35, 269-277. Wester, R. C., Maibach, H. I., Sedik, L., Melendres, J., Wade, M., and DiZio, S. (1993a). Percutaneous absorption of pentachlorophenol from soil. Fundam. App!. Toxieo!. 20, 68-71. Wester, R. C., Maibach, H. I., Sedik, L., Melendres, J., and Wade, M. (1993b). In vivo and in vitro percutaneous absorption and skin decontamination of arsenic from water and soil. Fundam. App!. Toxieo!. 20,336-340.
Wester, R. c., Bucks, D. A, and Maibach, H.I. (1994). Human in vivo percutaneous absorption of pyrethrin and piperonyl butoxide. Food Chem. Toxieo!. 32,51-53. Wester, R. c., Quan, D., and Maibach, H. I. (1996a). In vitro percutaneous absorption of model compounds glyphosate and malathion from cotton fabric into and through human skin. Food. Chem. Toxieol. 34,731-735. Wester, R. c., Melendres, J., Logan, F., Hui, X., Maibach, H. I., Wade, M., and Huang, K. C. (1996b). Percutaneous absorption of 2,4dichlorophenoxyacetic acid from soil with respect to soil load and skin contact time: In vivo absorption in rhesus monkey and in vitro absorption in human skin. 1. Toxieol. Environ. Health 47, 335-344. Wester, R. c., Melendres, J., Sedik, L., Maibach, H.I., and Riviere, J. E. (1998). Percutaneous absorption of salicylic acid, theophylline, 2,4-dimethylamine, diethyl hexyl phthalic acid and p-aminobenzoic acid in isolated perfused porcine skin flap compared to man in vivo. Toxieol. Appl. Pharmaeo!' 1, 159-165. Williams, P. L., and Riviere, J. E. (1990). Efect of hyperthennia on cisplatin (CDDP) disposition to isolated perfused skin. Int. 1. Hyperthermia 6, 923932. Williams, P. L., Thompson, D., Qiao, G. L., Monteiro-Riviere, N. A., Baynes, R. E., and Riviere, J. E. (1996). The use of mechanistically defined chemical mixtures (MDCM) to assess component effects on the percutaneous absorption and cutaneous disposition of topically exposed chemicals. n. Development of a general dennatopharmacokinetic model for use in risk assessment. Toxieol. Appl. Pharmaeol. 141,487-496. Windheuser, J. J., Haslam, J. L., Caldwell, L., and Shaffer, R. D. (1982). The use of N,N-diethyl-m-toluamide to enhance dennal and transdennal delivery of drugs. 1. Pharm. Sci. 71, 1211-1213.
CHAPTER
23 Metabolism of Pesticides Emest Hodgson and Patricia E. Levi North Carolina State University, Raleigh
23.1 INTRODUCTION The word "metabolism" may be used to designate the sum of chemical reactions that serve to maintain life. Parts of this integrated whole are spoken of as protein metabolism, fat metabolism, nucleic acid metabolism, and the like. Such aspects, as they deal with the processing of the normal (endogenous) constituents of the body and the effect of pesticides on them, are dealt with in other chapters. The word "metabolism" may also be used to designate the effect of an organism, through its enzymes, on the chemical structure of foreign compounds (now more often referred to as xenobiotics). These effects, also called biotransformation, are the subject of this chapter. Given the enormous literature on pesticide metabolism, it is no longer possible to provide an exhaustive review of the subject. This chapter, a revision of the treatment in the previous edition (Hodgson et aI., 1991), is narrower in scope but revised in most aspects covered. The more recent reviews and book chapters referred to throughout are recommended as sources of detailed recent information.
23.2 POSSIBILITY OF EXTERNAL TRANSFORMATION OR METABOLISM The finding of a derivative of a compound in the tissues or excreta of an animal is not necessarily proof that the compound is the result of biotransformation in that organism. Compounds, especially in thin films, may undergo chemical change when exposed to light or heat. As early as 1961 Mitchell reported the effects of ultraviolet light on 141 pesticides and Matsumura (1975, 1985) summarized the effects oflight and other physical factors on pesticides and their movement in the environment. The rate and extent of photochemical degradation of pesticides depends upon the chemical nature of the pesticide, the wavelength of the light, and the presence of other chemicals. The latter may act as photosensitizers, forming reactive light-energized intermediates that react with pesticides, or they may react with photoenergized pesticides. The four best known types of photochemical reactions of aromatic pesticides are ring Handbook of Pesticide Toxicology Volume 1. Principles
substitution, hydrolysis, oxidation, and polymerization. Examples sumarized by Matsumura (1975) include the following: substitution of a ring chlorine in 2,4-D by a hydroxyl group, hydrolysis of carbaryl, oxidation of parathion, and polymerization of pentachlorophenol. The enzymes of plants and microorganisms are responsible for a wide range of biotransformations (Matsumura, 1985) and, as a result, an animal eating plants may ingest one or more derivatives as well as the compound originally applied. Some toxic ants may be metabolized by bacteria in the intestine, still external to the cells of the animal itself. An example of the latter is reduction of parathion to aminoparathion by rumen bacteria (Ahmed et aI., 1958). Athough all of these possibilities must be kept in mind, they are usually of secondary importance. Many derivatives known to be formed by light, plants, or microorganisms are also formed by mammalian enzymes.
23.3 BIOTRANSFORMATION-MAJOR REACTIONS A number of books review the biotransformation of xenobiotics, either in general or of particular chemical or use classes (e.g., Hodgson and Levi, 1994, 1997; Hodgson and Smart, 2000; Jakoby, 1980; Jakoby et aI., 1982; Klaassen, 1996; Wilkinson, 1976; Williams, 1959). Many treatments of pesticides (e.g., Chambers and Carr, 1995; Ecobichon, 1996; Hodgson and Meyer, 1997; Hodgson et aI., 1995a; Kulkarni et aI., 1984 and Rose et al., 1999) include considerations, not only of pesticide metabolism, but also of the significance of metabolism in the toxicity of pesticides to target and nontarget species. Many of the chemical reactions involved in the biotransformation of pesticides have now been traced to particular enzymes, although some are only inferred from the appearance of derivatives of the parent compound in the tissues or excreta of the dosed animal. Chemical reactions reported to occur in the metabolism of pesticides are summarized in Table 23.1. It should be noted that biotransformation reactions of pesticides may be either detoxications or activations. Hollingworth
531
Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
532
CHAPTER 23
Metabolism of Pesticides
Table 23.1 Pesticide Metabolism-Chemical Reaction with Examples Reaction
Examples
Oxidation N-Dealkylation
R_N ..... G.1Hs 'G.1 H5 Phosphamidon
O-Dealkylation
Stable epoxide formation
0 G.!HsO -}I-OX G.! H50
--+
--+
(0
* =CI
Naphthalene 1 ,2-epoxide
~
--+
~ON:~::droXYI_1_naPhthYI
Oi l
~
....::
n-methyl carbamate
H
06
~ONHCH20H
CH
~
Acetaldehyde
H [05 J
Carbaryl
Side chain hydroxylation
Q-i3 CHO
:d:r!H
--+
~
05~ ~
+
Heptachlor epoxide
Naphthalene
Ring hydroxylation (usually via arene oxide formation)
~
HO G.!HsO::r -OX
Acetaldehyde
H
:ap ~
Q-i3 CHO
Desmethylchlorfenvinphos
Heptachlor
Arene oxide formation
+
Desethylphosphamidon
Chlorfenvinphos
Epoxidation
H R_N ..... 'G.1 H5
--+
"
~
--+
0:)1 ~
~
1·Naphthyl N-hydroxymethyl carbamate
Carbaryl
S
S
Sulfoxidation
(C2H50l2PSCH2S~Hs Phorate
--+
0
(C2Hs0l2PScH2t~H5 Phorate sulfoxide
(continues)
et al. (1995) provided a detailed review of the detection and significance of active metabolites of pesticides. The biotransformation of most pesticides involves a combination of several chemical reactions and in some instances breakdown products may become part of the general metabolic pool. For example, formaldehyde formed in demethylation reactions may be incorporated into the one-carbon metabolic pool. Williams (1959) first suggested that the metabolism ofxenobiotics generally occurs in two phases. Phase I involves predominantly oxidations, reductions, and hydrolyses and serves to introduce a polar group into the molecule. Phase 11, consisting primarily of conjugation reactions, involves the combination of
the products of phase I reactions with one of several endogenous molecules to form water soluble, and hence excretable, products. In the past most emphasis has been placed on microsomal cytochrome P450 (P450)-dependent oxidations and reductions of pesticides. These are discussed in detail in subsequent sections. However, much has been learned of the roles of other phase I enzymes such as flavin-dependent monooxygenases (FMO), hydrolases, and epoxide hydrolases, and of cooxidation during prostaglandin synthesis. Considerable emphasis has also been placed on the phase 11 conjugation reactions as they apply to pesticide metabolism.
23.3 Biotransformation-Major Reactions
533
Table 23.1 (continued) Examples
Reaction
o N-Oxidation
G-NH8OCH(CH3 )2 IPC
N-Oxide Formation
--
o G - 6 :8OCH(CH3 )2 Hydroxy IPC
--
er:? N
CH 3
0-90 N
CH3
Nicotine-1 '-N-oxide
Nicotine
Methylenedioxy ring cleavage
R~OH
R~OH+
_
C~O'C~. . / R~O
I
'-...
HCHO
Catechol Complexes with Fe+ 2 of cytochrome P450
Desulfuration/or dearylation
Reduction
--
Reduction of nitro group
Dechlorination DDT
DDD
(continues)
By definition, microsomal enzymes are those found in the microsomal fraction of a tissue following differential centrifugation. The terms microsomal fraction and microsomes refer to a biochemical preparation and do not correspond to any particular cell structure. However, the major component is derived from the endoplasmic reticulum and its constituent ribosomes. The microsomal fraction consists primarily of rough (with ribosomes) and smooth (without ribosomes) membranous vesicles that correspond to rough and smooth endoplasmic reticulum. The majority of studies focusing on pesticide metabolism and the regulation of pesticide-metabolizing enzymes have been performed in experimental animals, primarily rats, mice,
and monkeys. However, there has been an increase in information about human enzymes, especially the P450 isozymes. Much of this information has been gained through the use of specific substrates, antibodies, and cDNA probes. Studies with human P450s, although fragmentary, have demonstrated that xenobiotic metabolism and the regulation and expression of xenobiotic-metabolizing enzymes may be quite different in humans and in experimental animals. Such differences make the extrapolation of metabolism studies from experimental animals to humans difficult. It is only as we learn to understand these differences that we can make more accurate and realistic extrapolations to humans.
534
CHAPTER 23
Metabolism of Pesticides
Table 23.1 (continued) Reaction Reduction of a double bond
Examples
R-C::c~H R-
'Cl
+
2H
DDMU
Hydration of a double bond
R-C::c~H R-
'H
+
HOH
-
R H R-t-t-CI
~ ~ DDMS R H R-t-t-oH
~ ~ DOOH
DDNU
Hydrolysis Phosphate ester hydrolysis
o A:2~O(S)R' Most organic phosphorus esters
o Amide cleavage
R&HCH 3 Dimethoate (showing part of side chain)
o Thioester cleavage
CH3 (CH2 l2S8N(C2 Hs)2 Pebulate
-
0
A:2~OH
+
HO(S)R' Alcohol
Acid
-
0 R80H
NH 2CH 3
+
O-Q-Dimethyl-S-carboxylmethyl phosphorodithioate
-
Methylamine
0 CH 3 (CH2 )SH
+
H08N(C2 H5 J2 Acid
Propyl mercaptide
qa 0
11
Deamination
Rulene®
.
Desaturation
.'0'. 1,2,3.4,5,-Hexacyclorocyclohexane
CH3 O-P-OH
NH 2 CH 3
+
Methylamine
C(CH 3h Deaminomethyl Rulene®
.'0'. Hexachlorocyclohexene
(continues)
23.3.1 CYTOCHROME P450 MONOOXYGENASES
Although many enzymes acting in concert may be required for degradation, the initial reaction usually involves a microsomal phase I enzyme catalyzing an oxidation reaction. Reduction reactions, although they may also occur, are relatively uncommon. These enzymes include many of the isozymes of P450 active in the P450-dependent monooxygenase system, as well as FMO isozymes. Many different pesticide monooxygenation reactions are attributed to P450, including epoxidation (e.g., aldrin), N-de-
alkylation (e.g., alachlor, atrazine), O-dealkylation (e.g., chlorfenvinphos), S-oxidation (e.g., phorate), and oxidative desulfuration (e.g., parathion) (Kulkarni and Hodgson, 1980, 1984a, b). Currently the P450 superfamily comprises over 480 genes classified into 74 gene families, 14 of which exist in all mammals (Nelson et aI., 1996). The total number of functional P450 genes in any mammalian species is thought to range from 60 up to 200 (Gonzalez, 1990). In vertebrates, some P450 families encode proteins involved primarily in specific endogenous functions (i.e., steroid hormone biosynthesis and metabolism). Other families appear to have more to do with the oxidation
23.3 Biotransformation-Major Reactions
535
Table 23.1 (continued) Examples
Reaction With glutathione
RX +
HSCH2yHC(0)NHCH2COOH NHC(0)CH2CH2CH(NH2)COOH ,
glutathione S-transferase
RSCH 2qHC(0)NHCH2COOH NHC(0)CH2CH2CH(NHz)COOH
, r - glutamyltranspeptidase RSCH2~(0)NHCH2COOH
NH2
+ glutamate
,cysteinYI glycinase
+ glycine
RSCH2CH(NH,)COOH ,
N-acetyl transferase
RSCH2CHCOOH I
NHC(0)CH 3 Mercapturic acid
x
X
Rl0~~y
+
GSH
RIO ~~y HO'
_
RIO'
+ GSR
1
(I)
x
X
R,O~by +
GSH
RIO'
Rl0~by
./'
+ GSR 1
(11)
X Rl0~bH + GSY
(Ill)
HO'
~
RIO'
X Rl0~by +
R:?'
./' GSH
X Rl0~bH
+ GSY
(IV)
+ YOH
(V)
R:?' X
~ Rl0~SG
R:?'
RI = alkyl; R;F aryl; X= S or 0; Y= leaving group
With thiosulfate
C5NCyanide
+
-
eNSThiocyanate
(continues)
of exogenous compounds, such as pesticides, and often display a broad range of substrate specificity (Bogaards et al., 1995; Nebertetal.,1989). The details of the interaction of P450 with xenobiotics have been the subject of intense study for some time, although in these studies clinical drugs have been utilized to a much greater extent than pesticides. Reviews of pesticide studies include Hodgson and Kulkami (1974), Hodgson (1974), and Kulkarni and Hodgson (1980, 1984a, b). Studies of spectral interactions of pesticides with P450, interactions that may be indictive of the ability to act as substrate or inhibitor, have also been carried out (Mailman and Hodgson, 1972; Mailman et aI., 1974). Studies
using specific isozymes (Hodgson et aI., 1998; Levi and Hodgson, 1984, 1988) indicate that even in the same organ of the same species particular pesticides are metabolized at different rates by different P450 isozymes. The specificity of different isoforms for pesticide substrates is an area of current interest. Due to the availability of heterologous expressed human isoforms these studies can now be carried out on human enzymes as well as those from experimental animals. In an early study of fenitrothion metabolism by mouse liver utilizing four constitutive and two induced P450 isoforms, Levi et al. (1988) showed all isoforms produced both the cresol detoxication product and the oxon. However, there
536
CHAPTER 23
Metabolism of Pesticides
Table 23.1 (continued) Reaction
Examples eOOH
With cystine
+ Cystine _ _
C::N"
COOH
.CH S NI; .... S ,N CH2
·CH
CH 2
·c'
Cyanide
'e'
11
I
NH
NH2
2-lminothiocidine-4-carboxylic acid
OH
,-,-e ''3
With acetate
Acetate
__
~
w-He -CH3
17 I
-..;:
h
N02
2·Amino-6-methyl-4-nitrophenol (a metabolite 01 DNOC)
2-Aetamido-6-methyl-4-nitrophenol
3-Hydroxymethyl phenylsullone
Methylation
Fonolos
o
Fonolos oxon
~e-M-Q-oH o
4-Hydroxymethyl phenylsullone
were significant differences both in overall activity and in the ox on/cresol ratio. The most active isoform, induced by phenobarbital and now known as CYP2B 10, was also active in the metabolism of parathion and methyl parathion and in all cases produced more significantly more oxon and detoxication products. Human CYP3A4 was shown to be most active in the metabolism of parathion although CYP1A2 and 2B6 also showed activity (Butler and Murray, 1997). In ongoing studies on the metabolism of chlorpyrifos by human P450 (Tang and co-workers, unpublished) it has been shown that CYP2B6 and CYP2C19 are both active, CYP2B6 producing an excess of chlorpyrifos ox on and CYP2C 19 an excess of detoxication products. Studies of triazine herbicides in mice (Adams et aI., 1990) and rats (Hanioka et aI., 1999) as well as in rats and pigs (Lang et aI., 1996) suggested a broad lack of isoform specificity for these substrates. However, Lang et al. (1997) showed that in humans CYP1A2 appeared to be the principal, if not the only, isoform responsible for triazine herbicide oxidation. Inui et al. (2000) expressed human CYP1A1, CYP2B6, and CYP2C19
in potatoes and produced resistance to several herbicides, including atrazine, in the host plants, presumably by enabling the plants to metabolize the herbicides. In studies of chloroacetanilide herbicides (Coleman et aI., 1999) it was shown that human CYP3A4 was responsible for the initial O-dealkylation of alachlor. Subsequent studies (Coleman et al., 2000) extended these studies to acetochlor, butachlor, and metachlor. In all cases, CYP3A4 was the most active human isoform, although CYP2B6 also had some activity. One of the significant features of many of the microsomal enzymes is their inducibility by xenobiotics; thus, stimulation of the metabolism of a chemical by prior administration of the same or another chemical is often taken as presumptive evidence of its metabolism by microsomal enzymes. For example, in mice pretreated with phenobarbital there is an increase in phorate metabolism, suggesting that P450 isoforms, such as CYP2B or CYP3A forms, may be important in the metabolism of similar pesticide substrates (Kinsler et aI., 1990). Table 23.2 gives a list of reactions, using pesticides as examples, catalyzed by these enzymes. A less descriptive, but more
23.3 Biotransformation-Major Reactions
537
Table 23.2 Some Examples of Metabolic Pathways Catalyzed by Liver Microsomal Enzymes« Reaction
Example
References
4-Nitrophenyl N.N-dimethyl carbamate --f
Hodgson and Casida (196 I),
OXIDATION Cytochrome P450 N -Dealkylation
4-nitrophenyl N -methyl carbamate
Strother (1972)
Dimethoate --f des-N-methyl derivatives
Lucier and Menzer (1970)
Dicrotophos --f des-N -methyl derivatives
Tseng and Menzer (1974)
Nicotine --f nomicotine
Papadopoulos (1964)
Methoxychlor --f mono- and dihydroxy
Kapoor et al. (1970)
0- Dealky lation Ether cleavage
derivatives Propoxur --f 2-hydroxyphenyl N -methyl
Oonithan and Casida (1966, 1968)
carbamate Alachlor --f O-dealkylated derivatives
Hodgson et al. (1998),
Chlorfenvinphos --f desethyl derivatives
Hutson (198 I),
EPN oxon --f desmethyl EPN oxon
Nomeir and Dauterman (1979) Mazel et al. (1964)
Coleman et al. (1999, 2000) Ester cleavage
Donninger et al. (1967,1972) S-Dealkylation
6-Methylthiopurine --f 6-mercaptopurine
Dearnination
Amphetamine --f phenylacetone
Axelrod (1955,1956)
Epoxidation (stable
Heptachlor --f heptachlor epoxide
Khan (1969)
epoxides) Ring hydroxylation (via arene oxide)
Aldrin --f dieldrin
Kulkami and Hodgson (l984a)
I-Naphthyl N-methyl carbamate (carbaryl) --f
Dorough and Casida (1964),
4- and 5-hydroxy-I-naphthyl N -methyl
Dorough (1970), Hutson (1981)
carbamate Naphthalene --f naphthalene epoxide
Jerina et al. (1968, 1970)
Dieldrin --f 12-hydroxy dieldrin
Baldwin et al. (1972),
Methoxychlor
Dehal and Kupfer (1994)
Hutson (1976) Side chain hydroxylation
I-Naphthyl N-methyl carbamate (carbaryl)--f I-naphthyl N-hydroxymethyl carbamate
Heterocyclic ring
Dorough and Casida (1964) Hutson (1981)
Butacarb --f hydroxybutyl derivatives
Douch and Smith (1971a, b)
TOCP --f hydroxy methyl TOCP cyclic phosphate
Eto et al. (1962)
Pyrethrins --f hydroxymethyl derivatives
Casida etal. (1975-1976)
Nicotine --f hydroxynicotine
Hucker et al. (1960)
Parathion --f paraoxon
Davison (1955),
Diazinon --f dizoxon
Yangetal. (1969, 1971)
hydroxy lation Desulfuration or dearylation
Kamataki and Neal (1976) Other organophosphates
Kulkami and Hodgson (1984a)
Dehydrogenation
f3- and y-Chlordane --f dichlorochlordene
Street and B1au (1972),
Sulfoxidation
Phorate --f phorate sulfoxide and phorate sulfone
Levi and Hodgson (1988),
Aldicarb --f aldicarb sulfoxide and sulfone
Perkins et al. (1999)
Chadwick et al. (1975) Hodgson et al. (1998) Flavin-containing monooxygenase
N -Oxidation
Trimethylamine --f trimethylamine oxide
Baker and Chaykin (1961)
Nicotine --f nicotine N -oxide
Tynes and Hodgson (1985a, b)
Tetram --f tetram N -oxide
Hodgson (1982-1983) (continues)
538
CHAPTER 23
Metabolism of Pesticides
Table 23.2 (continued) Reaction Sulfoxidation
Example
References
Phorate ---> phorate sulfoxide
Hajjar and Hodgson (1980, 1982a),
Methiocarb ---> methiocarb sulfoxide
Tynes and Hodgson (1985a, b),
Metam sodium
Smyser et al. (1985)
Fonofos ---> fonofos oxon
Hajjar and Hodgson (1982b),
Cherrington et al. (1998a, b) Buronfosse et al. (1995) Oxidative desulfuration
Smyser et al. (1985) REDUCTION Nitro group
Parathion ---> amino parathion
Hitchcock and Murphy (1967)
Dechlorination
DDT--->TDE
Esaac and Matsumura (1984)
Azo compounds
Azobenzene ---> aniline
Fouts et al. (1957)
HYDROLYSIS Deesterification
Procaine
=}
2-diethylaminoethanol and
Brodie (1956)
P -aminobenzoic acid Acetanilide ---> aniline and acetic acid
Hollunger and Niklasson (1962)
O,O-DimethyI2,2-dichlorovinyl phosphate
Hodgson and Casida (1962)
(DVP) ---> desmethyl DDVP Deltamethrin ---> 3-(2,2-dibromovinyl)-2,2-cydo-
Akhtar (1984)
propane carboxylic acid and 3-phenoxybenzaldehyde Permethrin ---> hydrolysis products
Ghiasuddin and Soderlund (1984)
CONJUGATION Glucuronidation
O-Aminophenol ---> O-aminophenol glucuronide
Isselbacher and Axelrod (1955)
Dieldrin ---> dieldrin glucuronide
Baldwin et al. (1972), Hutson (1976), Matthews and Matsumura (1969)
Carbaryl ---> naphthyl glucuronide
Chin et al. (1979a, b, c)
Sulfation
Carbaryl ---> naphthyl sulfate
Chin et al. (1979a, b, c)
Acetylation
Fluoroacetamide ---> fluoroacetyl CoA
Peters, 1963
Glutathione conjugation
Methyl parathion ---> desmetyl methyl parathion
Hollingworth (1969)
a Pesticides
are cited as examples whenever possible.
mechanism-based, classification of P450-catalyzed xenobiotic oxidations is that of Guengerich and MacDonald (1984). They classified such reactions into six general categories:
1. Carbon hydroxylation-the formation of an alcohol at a methyl, methylene, or methine position 2. Heteroatom release-the oxidative cleavage of the heteroatom part of a molecule resulting from a hydroxylation adjacent to the heteroatom that generates a geminal hydroxy heteroatom-substituted intermediate such as a carbinol amine, halohydrin, hemiacetal, hemiketal, or hemithioketal (This intermediate then collapses to release the heteroatom and form a carbonyl compound.) 3. Heteroatom oxygenation-the conversion of a heteroatom-containing substrate to its corresponding heteroatom oxide as in the formation of N -oxides, sulfoxides, or phosphine oxides
4. Epoxidation-the formation of oxirane derivatives of olefins or aromatic compounds 5. Oxidative group transfer-a type of reaction that involves a 1,2-carbon shift of a group with the concurrent incorporation of oxygen to form a carbonyl at the Cl position 6. Olefinic suicide destruction-inactivation of the heme of P450 by an enzyme product
23.3.2 FLAVIN-CONTAINING MONOOXYGENASE The microsomal flavin-containing monooxygenase was known for a number of years as an amine oxidase but is now known to be also a sulfur oxidase and a phosphorous oxidase. Like P450, the FMO is a microsomal enzyme, a monooxygenase requiring
23.3 Biotransformation-Major Reactions
539
Table 23.3 In Vitro Tests for Microsomal Xenobiotic-MetaboIizing Enzymes Reaction or enzyme
Substrate
Reference
Ester hydrolysis
Methylthiobutyrate,
Heymann and Mentlein (198 I)
p-nitrophenyl acetate Cytochrome P450 Benzo[a ]pyrene
Gelboin and Conney (1968)
Ethoxyresorufin
Pohl and Fouts (1980)
lA2 (rat)
Acetanilide
Mitoma and Udenfriend (1962)
2BI12 (rat)
Pentoxyresorufin
Lubet et al. (1985, 1990)
Benzphetamine
Werringloer (1978)
Testosterone, 16a,16f1-
Wood et al. (1983), Sonderfan et al. (1987)
IAl (rat)
Denison et al. (1983) Aitio (1978), Burke and Mayer (1974) Lewandowski et al. (1990)
hydroxylation 2El (rat)
p- Nitrophenol
Koop (1986)
3Al (rat)
Testosterone,
Sonderfan et al. (1987), Li et al. (1995)
4AI (rat)
Lauric acid
6,B-hydroxylation Kinsler et al. (1988)
Flavin-containing monooxygenase Phorate
Levi and Hodgson (1988)
N ,N -Dimethy laniline
Tynes and Hodgson (l985a, b)
Methimazole
Dixit and Roche (1984)
Thiobenzamide
Cashman and Hanzlik (198 I)
NADPH and oxygen, and exists as multiple isozymes in various tissues. FMO, unlike P450, catalyzes only oxygenation reactions, has a more specific substrate requirement, and is not known to be subject to induction or inhibition by xenobiotics, apart from competitive inhibition by alternate substrates (Kulkarni and Hodgson, 1984a, b; Ziegler, 1980). The mechanism of catalysis is also distinct in that electrons are transferred directly from NADPH, and not via an NADPH-reductase. Also, because the formation of the hydroperoxyfiavin form of the enzyme precedes interaction with the substrate, maximum velocity (Vmax ) for a particular FMO isoform is constant for all substrates although the Michaelis constant (Km) can vary from one substrate to another. P450 isoforms, on the other hand, show variations from one substrate to another in both Vmax and Km. The FMO is found in highest levels in the liver, but is also found in significant levels in the lung and kidney. Recent studies have identified five forms of FMO (FM01FM05) which are differentially expressed with respect to species and tissue (Lawton and Philpot, 1995; Lawton et aI., 1994). Each species which has been examined by analysis of genomic DNA appears to contain the same set of FMO genes (Lawton et aI., 1994). Thus whereas the FMO family possesses multiple isozymes, the number of forms is small compared to that of the P450 family. Although the FMO isozymes are
catalytically similar, marked differences do exist in substrate specificity. The importance of the FMO in pesticide metabolism was established when it was discovered that the FMO oxidizes a variety of thioether-containing pesticides (Cherrington et aI., 1998a, b; Hajjar and Hodgson, 1980, 1982a, b; Levi and Hodgson, 1992; Smyser et aI., 1985; Tynes and Hodgson, 1985a). It has since been shown that the FMO is capable of oxidative desulfuration (oxon formation) of certain phosphonate insecticides such as fonofos through a mechanism distinct from that of oxon formation by P450 (Smyser and Hodgson, 1985; Smyser et aI., 1985) as well as the oxidation of pesticides from a number of different chemical classes (Tynes and Hodgson, 1985a). Reviews include Hodgson (1982-1983), Kulkarni and Hodgson (l984a, b), and Hodgson et al. (1998). FMO isoform specificity in pesticide metabolism is also being investigated. For example, Cherrington et al. (l998a, b) showed that, in the mouse, FMOl metabolizes phorate to phorate sulfoxide but FM05 is without activity. 23.3.3 OTHER PHASE I REACTIONS 23.3.3.1 Epoxide Hydrolases
These enzymes are also known to attack pesticide substrates, although the reactions are subsequent to the initial formation of
540
CHAPTER 23
Metabolism of Pesticides
epoxides. Examples include naphthalene 1,2-oxide and the 3,4and 5,6-epoxides of carbaryl (Dorough and Casida, 1964) and tridiphane (Magdalou and Hammock, 1987). 23.3.3.2 Prostaglandin Synthetase
Prostaglandins are synthesized in mammals via a reaction sequence starting with arachidonic acid as substrate. During the second, or peroxidase, step of prostaglandin synthetase action, xenobiotics can be cooxidized to yield products similar to those formed by various isozymes of P450 (Eling et aI., 1983; Marnett and Eling, 1983). A number of pesticides (e.g., aminocarb, parathion) have been shown to act as substrates. These reactions may be important in extrahepatic tissues low in P450 and high in prostaglandin synthetase, such as the seminal vesicle and the inner portion of the medulla of the kidney. 23.3.3.3 Hydrolases and Amidases
Hydrolase and amidase activity are known to be important in phase I metabolism of pesticides. For example, dimethoate is detoxified by amidase activity and the selective toxicity of malathion is due, in large part, to the presence in mammals of carboxylesterases not widely distributed in insects. These enzymes are known from both microsomes and the soluble cytoplasm but are more commonly found in the latter. It appears likely that in most cases amidase and esterase activity are different activities of the same enzymes (Satoh, 1987). 23.3.4 PHASE 11 REACTIONS: CONJUGATIONS
Conjugations may be simple, as in the case of phenol, but often they are more complicated processes in which the final product is derived by several steps. In spite of this possible complexity, it is useful to think of conjugation of xenobiotics taking place with glucuronic acid to form glucoronides, N -acetylcysteine to form mercapturic acids, glycine to form hippuric and related acids, sulfate to form ethereal sulfates, thiosulfate ions to form thiocyanate, and glutamine to form conjugates of the same name. In fact, the actual conjugations often occur with derivatives of the conjugating molecule, for example, with glutathione, uridine diphosphate glucuronic acid, or phosphoadenine phosphosulfate. Conjugates of foreign chemicals that are rare in mammals, or known only in other classes or phyla, include glucosides, ribosides, ornithines, sulfides, and conjugates with serine, metal complexes, and methylated or acetylated compounds. With the exception of glutathione conjugation, most conjugation reactions involving pesticides are secondary, involving, as substrates, the products of phase I reactions. They include glucoside formation, glucuronic acid formation, sulfate formation, and conjugation with amino acids. This area, as it applies to pesticides, was reviewed in detail by Dorough (1984). Conjugation with glutathione, mediated by one of the gluthathione S-transferases, is the first step in a sequence leading to a mercapturic acid (Table 23.1). Several pesticides are
metabolized in this way, particularly organophosphorus compounds, DDT, y-HCH, and organothiocyanates. These reactions and their relationship to pesticides have been reviewed by Motoyama and Dauterman (1980) and by Fukami (1984). The glutathione S-transferases (GSTs) are an abundant family of dimeric proteins that have the capacity to conjugate glutathione (GSH) with a variety of compounds containing electrophilic centers. The major hepatic cystolic GSTs in mammalian liver can be divided into three classes-alpha (a), mu (p.,), pi (n )-based on sequence similarity and catalytic activity (Mannervik et aI., 1985). Each class may contain one or more functional enzymes. Although all of these classes are capable of binding to a wide variety of pesticides, the mu class has somewhat higher affinity than the alpha or pi classes (Dillio et aI., 1995; Hayes and Wolf, 1980). Members of the mu class GSTs are responsible for conjugating benzo[a ]pyrene-7 ,8-diol9,1O-epoxide (BPDE) as well as a wide variety of pesticides such as the organophosphate insecticides, the halogenated hydrocarbon insecticides, and the S-triazine herbicides (Hayes and Wolf, 1980). Polymorphisms are known to occur in humans in regard to GST enzymes. About 50% of the Caucasian population in the United States is deficient in mu class GSTM 1. This polymorphism is due to a deletion in the GSTMl gene resulting in the lack of GSTMl protein formation. Epidemiological studies have implicated this deficiency in an increased risk of lung cancer in smokers, presumably due to the ability of GSTM 1 to detoxify chemical carcinogens such as BaP in tobacco smoke (Bell et aI., 1992; Nakachi et aI., 1993; Seidegard and Pero, 1985; Wormhoudt et aI., 1999). Because conjugation reactions other than those mediated by the GSTs are less well known in the metabolism of pesticides, the enzymatic basis of these conjugations is not discussed in detail. However, this matter has been reviewed in detail by Motoyama and Dauterman (1980), Dorough (1984), Matsumura (1985), and Hollingworth et al. (1995) and several types of conjugation are known to involve pesticides. Glucuronides are important in the metabolism of carbamates such as banol, carbaryl, and carbofuran (Mehendale and Dorough, 1972) as well as some organophosphate compounds (Hutson, 1981) and other chemcials. Ethereal sulfates, while less important in the metabolism of pesticides than glucuronides, nevertheless may be formed from carbofuran and other carbamates (Dorough, 1968). Glutathione conjugation is important in the metabolism of organophosphates (Motoyama and Dauterman, 1980) and the conjugated products of glutathione adducts may be further metabolized to mercapturic acids, the N -acety1cysteine derivative of the original xenobiotic substrate. Insects and plants are unusual in forming glucosides rather than glucuronides. 23.3.5 BIOTRANSFORMATION IN EXTRAHEPATIC TISSUES
The liver is more important than other organs in the biotransformation of foreign chemicals. However, other organs and tissues may be active to some degree. For example, it was shown
23.3 Biotransformation-Major Reactions early that DDT is degraded by rat diaphragm, kidney, and brain in vitro (Judah, 1949). Later study showed that these changes must be at a very slow rate in vivo. However, not all extrahepatic metabolism is inefficient. Organophosphorus compounds may be rapidly degraded. For example, slices of rabbit skin hydrolyze paraoxon (at a concentration of 7.7 x 10-3 M) to the extent of 20% in 1 hr per gram of tissue. Because absorption of paraoxon and related compounds is slow, this metabolism may be an important defense mechanism (Fredriksson et aI., 1961). Furthermore, parathion is metabolized to paraoxon and diethylphosphorothioic acid by rabbit lung at about 20% of the rate in liver (Neal, 1972). Some carbaryl is hydrolyzed and the resulting naphthol is conjugated with glucuronic acid by the intestine (Pekas and Paulson, 1970). Some enzymes outside the liver may be induced, but the matter has received little attention. Wattenberg (1971) demonstrated that the small intestines of rats fed a balanced purified diet or starved for 1 day possess virtually no benzo[a]pyrene hydroxylase activity, whereas the intestines of rats fed the same diet plus turnip greens, broccoli, cabbage, or brussel sprouts have marked activity of this enzyme. The same activity in human skin is induced by polycyclic hydrocarbons (Alvares et aI., 1973). Neal (1972) showed that monooxygenases of the lung active in the metabolism of parathion can be induced by phenobarbital. 23.3.5.1 Lung The lung is a primary site of exposure to airborne as well as blood-borne environmental pollutants, such as pesticides, and for this reason is a target organ for many chemically induced toxicities (Bond, 1983, 1993; Dahl and Lewis, 1993). Because the lung has a full complement of metabolic enzymes, it has the capacity to activate and deactivate pesticides and other xenobiotics. Several studies have demonstrated the importance of pulmonary P450 and FMO enzymes in pesticide oxidation (Feng et aI., 1990; Li et aI., 1992). In the lung FMO appears to play a more important role than P450 in the oxidation of certain pesticides and xenobiotics (Kinsler et aI., 1988; Tynes and Hodgson, 1983, 1985a, b). Other studies have shown the existence of an FMO form (now known as FM02) in the lung not present in the liver (Lawton et aI., 1990; Tynes and Hodgson, 1983; Tynes et aI., 1985; Venkatesh et aI., 1992a; Williams et aI., 1984).
541
epoxide hydrolases, glutathione s-transferases, and uridine 5'-diphosphate (UDP) glucuronyl transferases. It is of some interest that, despite the low concentrations of nasal CYP enzymes, these have been demonstrated to have greater specific activity toward several substrates than liver CYPs; perhaps as a result of higher rations of NADPH cytochrome P450 reductase to CYP, in the nasal tissues. It is also interesting that nasal P450s appear to be less inducible than liver isoforms, although they appear to be sensitive to a number of P450 inhibitors. Few pesticides are known to have negative interactions with nasal tissues. However, alachlor, a restricted-use chloroacetamide herbicide which at one time was widely used in agriculture, was demonstrated to cause rare nasal carcinomas in rats. The putative metabolic product thought to be responsible for its carcinogenicity was identified as diethy Ibenzoquinone imine (DEBQI), which is produced only after extensive metabolism of alachlor, involving CYPs as well as an aryl amidase. Human CYP isoforms 2B6 and 3A4 are among those which have been identified as being important in the production of metabolite precursors to DEBQI (Coleman et aI., 1999,2000). Gente + w-water (Deamer et aI., 1994; Genter et aI., 1995, 1998) have demonstrated the role of microsomal epoxide hydrolase and CYP2E1 in the nasal toxicity of dichlobenil in the mouse. It was subsequently shown that CYP2AlO and 2A11, isoforms that comprise some 25% of the total olfactory cytochrome P450 content, also play an important role in the nasal toxicity of dichlobenil (Ding et aI., 1994, 1996). 23.3.5.3 Skin Because the skin is the largest organ in the human body, is continuous over the surface area of the body, and is in direct contact with the environment, it is often the portal of entry for pesticides. The skin is known to contain many of the xenobiotic metabolizing enzymes found in the liver, and some of these have been shown to be inducible, primarily by polycyclic hydrocarbons (Goerz et aI., 1994; Jugert et aI., 1994). By use of in vitro methods, such as the isolated perfused porcine skin flap (Carver et aI., 1990) and mouse skin microsomes (Venkatesh et aI., 1992b), the skin has been shown to have the capacity to metabolize a variety of pesticides. For example, Chang et al. (1994), using the isolated perfused porcine skin flap, showed that both carbaryl and parathion were metabolized during uptake by the skin.
23.3.5.2 Nasal Tissues The nasal mucosa is the first tissue of contact for inhaled xenobiotics and compounds have been identified which cause nasal lesions or tumors in experimental animals. For humans, nasal toxicants which have been identified include industrial exposure to nickel, chromium, wood dust, and more recently the Mexico City urban environment (Calderon-Garciduenas et aI., 1992). The drug-metabolizing activity of nasal tissues has been reviewed by Reed (1993). Enzymes known to be present include a variety of cytochromeP450s (CYP1A1, 2Bl, 2E1, 3A1, 4A1, 2G 1), flavin-containing monooxygenases, carboxylesterases,
23.3.5.4 Kidney Because of the kidney's high blood flow, its ability to concentrate chemicals, and the presence of renal xenobiotic metabolizing enzymes, the kidney is often a site of toxicity from foreign chemicals. Many of these toxicities can be directly attributable to the presence and localization of specific forms of enzymes responsible for activation (Hu et aI., 1993; Speerschneider and Dekant, 1995). Several studies have highlighted the importance of renal oxidative enzymes, particularly FMO, in the metabolism of pesticides and other xenobiotics (Kinsler et aI., 1988; Tynes and Hodgson, 1983). As was the case with
542
CHAPTER 23
Metabolism of Pesticides
the lung, the renal FMO enzymes played a greater role in microsomal systems in the oxidation of several pesticides than renal P450, suggesting an important role for FMO in the extrahepatic metabolism of toxicants. Studies of kidney FMO have provided evidence for several isoforms in the kidney, including the forms found in liver and lung (Atta-Asafo-Adjei et aI., 1993; Burnett et aI., 1994; Venkatesh et aI., 1991). 23.3.5.5 Central Nervous System Very little is known about xenobiotic metabolizing enzymes in the central nervous system (CNS). Several studies have demonstrated P450 activity and constitutive expression of various P450 isozymes (Britto and Wedlund, 1992; Ghersi-Egea et aI., 1993; Hansson et aI., 1992; Hodgson et aI., 1993). The activation or detoxication of pesticides by the CNS is of particular interest for pesticides that exihibit their action and are metabolized in the brain. Studies by Chambers and Chambers (1989) demonstrated that the neurotoxicity of a series of organophosphorus compounds correlated better with activation in the brain than with activation in the liver. Several studies have reported activity of known FMO substrates by brain microsomes (Bhamre et aI., 1993; Duffel and Gillespie, 1984; Kawaji et aI., 1994), and one form of FMO has been demonstrated using polymetese chain reaction (PCR) amplification (Blake et aI., 1996). 23.3.6 TOXICITY OF METABOLITES In general, metabolites are less toxic than their parent compounds, if for no other reason than that they are usually more water soluble and, therefore, more rapidly excreted. There are notable exceptions for which biotransformation results in an inherently more toxic product. Such reactions are generally referred to as activation reactions. These reactive metabolites may combine covalently with cellular constituents such as DNA, RNA, or protein. Carcinogenesis, mutagenesis, and cellular necrosis are often attributable to such reactive metabolites (Parke, 1987; Guengerich, 1992, 1993; Anders et aI., 1992). The metabolic production of a more toxic compound is sometimes called lethal synthesis to emphasize that biotransformation in this instance is the source of danger. The term "lethal synthesis" was introduced in a lecture given on June 7, 1951, by Peters (1952) in connection with fluoroacetic acid. This compound is not itself toxic to enzymes, but is converted by enzymes into a highly toxic material. Peters (1963) later reviewed and extended the concept of lethal synthesis. The effects of metabolism may be complex, as illustrated by studies of bromobenzene. It has been known for some time that the liver necrosis associated with this compound is caused by one or more toxic metabolites. Stimulation of its biotransformation by phenobarbital potentiates the injury of toxic doses to the liver, and inhibition of its metabolism by SKF 525-A prevents this injury. However, although 3-methylcholanthrene causes a slight in vitro stimulation of the metabolism of bromobenzene and does not alter the overall rate in vivo, it does pro-
tect against the hepatotoxicity. Rats dosed with bromobenzene after induction with 3-methylcholanthrene excrete more bromophenyldihydrodiol, bromocatechol, and 2-bromophenol than do uninduced rats. Increase in the first two compounds suggests an increased capacity to detoxify the highly reactive epoxide. Increase in 2-bromophenol suggests that 3-methylcholanthrene diverts the metabolism of bromobenzene to a comparatively nontoxic pathway (Zampaglione et aI., 1973). Of particular concern has been the role of metabolic activation in the carcinogenic process, particularly in the formation of DNA adducts by reactive metabolites. For example, monooxygenase enzymes have been postulated to play a role in the metabolic activation of alachlor and metolachlor (Brown et aI., 1988; Feng and Wratten, 1989; Feng et al., 1990; Jacobsen et aI., 1991; Li et aI., 1992). Although studies suggest that a1achlor has a greater carcinogenic potential than metolachlor, the carcinogenic response to these compounds are species and tissue specific, alachlor being a nasal-specific carcinogen in rats but not in mice. Metolachlor, on the other hand is carcinogenic in liver but not nasal tissue (U.S. Environmental Protection Agency, 1986, 1987). Available evidence suggests that the species- and tissue-specific responses observed, particularly for alachlor, result from specific metabolic enzymes, including monooxygenases and arylamidases, and the generation of the putative carcinogenic metabolite, diethylbenzoquinone imine (see Coleman et aI., 1999,2000, for appropriate references). The most studied generation of reactive metabolites from pesticides is the generation of oxons from organophosphorus compounds containing the P=S moiety, by oxidative desulfuration. Not only does this reaction produce the oxons, cholinesterase inhibitors responsible for the neurotoxicity of these compounds, but it also releases reactive sulfur, a potent inhibitor of cytochrome P450 (see Section 23.4.4.4 for appropriate references). The mode of action of insecticide synergist, piperony I butoxide, and other methylenedioxyphenyl compounds is also due to a reactive metabolite, believed to be a carbene derivative, that combines with the heme iron of cytochrome P450 (Dahl and Hodgson, 1979; and see Sections 23.4.4.4 and 23.4.4.5 below).
23.4 CHEMICAL FACTORS AFFECTING METABOLISM 23.4.1 INTRODUCTION Although the study of the metabolism and toxicity of pesticides is simplified by considering single compounds, humans and other living organisms are not exposed in this way; rather, they are exposed to many xenobiotics simultaneously, involving different portals of entry, modes of action, and metabolic pathways. Because it bears directly on the problem of toxicityrelated interactions between different xenobiotics, the effect of chemicals on the metabolism of other exogenous compounds is important in the study of pesticide toxicity.
23.4 Chemical Factors Affecting Metabolism
Pesticides, and other xenobiotics, in addition to serving as substrates for a number of enzymes, may also serve as inhibitors or inducers of these or other enzymes. Many examples are known of compounds that first inhibit and subsequently induce enzymes such as the microsomal monooxygenases. The situation is even further complicated by the fact that, although some substances have an inherent toxicity and are detoxified in the body, others without inherent toxicity can be metabolically activated to potent toxicants. The following examples are illustrative of the situations that might occur involving two compounds:
1. Compound A, without inherent toxicity, is metabolized to a potent toxicant. In the presence of an inhibitor of its metabolism, there would be a reduction in toxic effect. 2. Compound A, given after exposure to an inducer of the activating enzymes, would appear more toxic. 3. Compound B, a toxicant, is metabolic ally detoxified. In the presence of an inhibitor of the detoxifying enzymes, there would be an increase in the toxic effect. 4. Compound B, given after exposure to an inducer of the detoxifying enzymes, would appear less toxic. In addition to these cases, the toxicity of the inhibitor or inducer, as well as the time dependence of the effect, must also be considered because, as mentioned previously, many xenobiotics that are initially enzyme inhibitors ultimately become inducers. In considering pesticides and metabolism, there are several ways to view the interactions-for example, other xenobiotics, such as clinical or other drugs, by causing enzyme induction or inhibition can affect the metabolism and thus the toxicity of pesticides. Conversely, pesticides, by acting as either enzyme inducers or inhibitors, can affect the metabolism of other xenobiotics, such as drugs, as well as the metabolism of endogenous compounds, such as steroid hormones. In the following sections, the discussion and examples will serve to illustrate these various interactions.
543
23.4.2 INDUCTION OF MICROSOMAL ENZYME ACTIVITY The stimulatory effect of xenobiotics on liver microsomal enzymes was reported first in the 1950s (Brown et al., 1954; Conney et al., 1957; Miller et al., 1954; Remmer, 1958) and, since then, has been entensively investigated. Numerous early experiments with laboratory rodents (Table 23.4) confirmed hepatic enzyme induction although until the end of the century methods were not available for identification of individual isozymes and the inducer was often classified as a phenobarbital-, a 3-methylcholanthrene-, or a mixed-type inducer. Reviews in this area include those of Conney (1967), Gelboin and Conney (1968), Sher (1971), Gillette et at. (1972), Nebert and Jensen (1979), Okey et at. (1986), Okey (1990), Batt et al. (1992), and Denison and Whitlock (1995). Reviews with emphasis on pesticides include those of Fouts (1963), Conney et al. (1967), Leibman (1968), Street et at. (1969), DuBois (1969), Hodgson (1974), Hodgson and Kulkarni (1974), Hodgson et at. (1980), Wilkinson and Denison (1982), Khan (1984), Kulkarni and Hodgson (1984a, b), Hodgson and Levi (1996), and Hodgson and Meyer (1997). It should be noted that induction is not restricted to xenobiotics, and enzymes may also be induced by hormones and other normal body constituents (Conney, 1967; Conney et aI., 1967, 1979; Kobliakov et al., 1991; Pantuck et al., 1979, 1984; Ronis and Cunny, 1994; Schenkman et al., 1989) and by dietary constituents (Anderson and Kappas, 1991; Donaldson, 1994; Wattenberg, 1971). A number of studies have provided evidence for induction of liver enzymes in animals and in humans who have been exposed occupationally or environmentally to pesticides (Table 23.5). For the most part these studies employed noninvasive in vivo techniques such as examination of the half-life of aminopyrene or phenylbutazone or excretion of 6,B-hydroxycortisol (Kolmodin et al., 1969; Kolmodin-Hedman, 1973; Kreiss et al., 1981; Poland et al., 1970).
Table 23.4 In Vivo Assessment of Altered Microsomal Activities in Humans and Animals Using Test Compounds Test compound
Test
Species
Reference
Aminopyrene
Breath test
Human
Jager et al. (1980)
Antipyrine
Plasma half-life, urinary excretion
Human
Mehta et al. (1982)
Rat
Butler and Dauterman (1989)
Caffeine
Plasma half-life, urinary excretion
Human
Kadlubar et al. (1992) ReIIing et al. (1992)
Chloramphenicol
Plasma half-life
Human
Mehta et al. (1975)
Hexobarbital
Sleep time
Rat
Butler and Dauterman (1988)
,B-Methy1digoxin
Plasma half-life, urinary excretion
Human
Hinderling and Garrett (1977)
Phenylbutazone
Plasma half-life
Human
Krishnaswamy et al. (1981)
Procaine
Paralysis
Rat
Butler and Dauterman (1988)
Salicylates
Plasma half-life, urinary excretion
Rat
Yu and Varma (1982)
Theophylline
Plasma half-life
Human
Mehta et al. (1982)
Rat
Butler and Dauterman (1988)
Table 23.5 Some Examples of Stimulation of Liver Microsomal Enzymes Involving Pesticides as Inducers or Substrates Dose
Species
Effect
Reference
Sesoxane
50 mg/kg, ip
Mouse
Hexabarbital sleeping time increased
Fine and MolIoy (1964)
Piperonyl butoxide
50 mg/kg, ip
Mouse
Hexobarbital sleeping time increased
Fine and Molloy (1964)
200 mglkg, ip
Mouse
Hexobarbital sleeping time increased up to 12 hr,
Kamienski and Murphy (1971)
Inducer
Synergists
decreased after 24-72 hr 400 mg/kg, ip
Mouse
Parathion toxicity increased after 1 hr, decreased
Kamienski and Murphy (1971)
after 48 hr 500 mglkg, ip
Mouse
P450 content decreased 2-12 hr, increased 12-36 hr
Philpot and Hodgson (1971-1972)
y-200mg/kg
Rat
In vitro metabolism of hexobarbital increased, hexobarbital
Koransky et al. (1964)
Cblorinated bydrocarbon insecticides BHC
sleeping time decreased, scillicocide toxicity decreased,
,8-200 mg/kg
all isomers similar
a-60mglkg Trichloro-237
25 mglkg, days 1-3,
Rat
Hexobarbital sleeping time decreased
Hart and Fouts (1963)
Rat
Hexobarbital sleeping time decreased
Hart and Fouts (1963)
Rat
Hexobarbital sleeping time decreased
Hart and Fouts (1963)
Rat
No effect on hexobarbital sleeping time
Hart and Fouts (1963)
Rat
Hexobarbital sleeping time decreased, hexobarbital
Hart and Fouts (1963)
6.25 mg/kg, day 4 Ut
y-Chlordane
t
25 mg/kg, days 1-3, 6.25 mg/kg, day 4
Endrin
25 mglkg, days 1-3, 6.25 mglkg, day 4
DDT
25 mg/kg, days 1-3, 6.25 mg/kg, day 4 500 ppm in diet,
metabolism in vitro increased, metabolism of aminopyrine
0.5-4 months
and p-nitrophenol increased, no effect on aniline metabolism 5-50 ppm in diet,
Increased detoxication of EPN, O-demethylation of
Rat
0.05-5.0 mg/kglday,
Kinoshita et al. (1966)
p-nitroanisole, and N -demethylation of aminopyrine
13 weeks Squirrel
2-6 months
monkey
In vitro metabolism of EPN and p-nitroanisole increased
Cranmer et al. (1972)
at higher dose levels
5 g in peanut oil, po
Human
CNS arousal following phenobarbital increased
Rappolt (1970)
50 mg/kg/day, 3 days
Rat
Increased 6,8-hydroxylation of testosterone, increased
Li et al. (1995)
ring hydroxylation of methoxychlor, increased CYP2B andCYP3A DDT and analogs
100 mg/kg/day, 3 days
Mouse
Increased P450 levels, aniline hydroxylase activity,
Abemathy et al. (1971 a, b)
zoxazolamine paralysis time, QSAR for 28 analogs dietary Increased CYP2B, hepatomegaly
Flodstrom et al. (1990) (continues)
Table 23.5 (continued) Inducer Methoxychlor
Dose
Species
Effect
Reference
200 mg/kg/day, 4 days
Rat
Increased 6,B-hydroxylation of testosterone, increased
Li et al. (1995)
ring hydroxylation of methoxychlor Diphenyl hydantoin
250 ppm, 3- 6 days
Rat
Storage of DDT and DDE decreased and in vivo
Cranmer (1970)
metabolism of DDT increased
Chlordane
o,p-DDD 100 mg/kg/day se
or Guinea Phenobarbital sleeping time decreased, in vitro
300 mg/kglday po, 2 days pig
metabolism of phenobarbital increased
10-100 mg/kg,
In vitro metabolism of hexobarbital, aminopyrine and
Rat
Straw et al. (1965) Hart and Fouts (1963)
chloropromazine not affected after 1 dose, an increased
1-3 days, ip
three doses 50 mg/kg, 14days
Dog
Decreased toxicity of dicoumarol
Welch and Harrison (1966)
1-50 mg/kg
Rat
Increased in vitro metabolism of estrone
Welch et al. (1971)
Lindane
15 mg/kg
Rat
Increased invitro metabolism of estradiol-l 713
Welch et al. (1971)
Heptachlor
lOmg/kg
Rat
Increased in vitro metabolism of estradiol-1713
Welch et al. (1971)
Toxaphene
5-50 ppm in diet,
Rat
Increased detoxication of EPN, O-demethylation of
Kinoshita et al. (1966)
p-nitroanisole and N -demethylation of aminopyrine
13 weeks Ut ~
Ut
Dieldrin
Mirex
Rat
Increased in vitro metabolism of estradiol-1713
p,p-DDD 25 mg/kg
Rat
Increased in vitro metabolism of estradiol-1713
Welch et al. (1971)
p,p-DDE 25 mg/kg
Rat
Increased in vitro metabolism of estradiol-1713
Welch et al. (1971)
0.07 mg/kg/day, 6 yrs
Rhesus Monkey
Increased in vitro metabolism of chlorfenvinphos
Wright et al. (1972)
2.0 mg/kg/day, 7 days
Dog
Increased in vitro metabolism of chlorfenvinphos
Wright et al. (1972)
8.0 mg/kg/day, 28 days
Rat
Increased in vitro metabolism of chlorfenvinphos
Wright et al. (1972)
1.6 mg/kglday, 350 days
Mouse
Increased in vitro metabolism of ch1orfenvinphos
Wright et al. (1972)
3 ng/kg
Rat
Increased in vitro metabolism of estradiol-1713
Welch et al. (1971)
5, 10, 25 mg/kg, 3 doses
Rat,
Increased in vitro p-nitroanisole O-demethylase
Baker et al. (1972)
Mouse 5, 10, 50 ppm diet Mirex and Chlordecone
Welch et al. (1971)
25 mg/kg
50 or 5 mg/kg, 5 doses
Rat Mongolian gerbil
30 mg/kg, ip
Increased biphenyl hydroxylation
Cianftone (1980)
Increased in vitro warfarin hydroxylase activity
Kaminsky et al. (1978)
and P450 content
10 or 1 mg/kg, 5 doses 10 mg/kg, 5 doses
activity and P450 content
Mouse
Increased in vitro benzo[a]pyrene hydroxylase
Crouch and Ebel (1987)
activity and P450 content Increased acute acetaminophen hepatotoxicity,
Fouse and Hodgson (1987)
mirex had greater effect 1, 10, 50 ppm in diet
Mouse
Increased P450, N- and O-demethylation
Fabacher and Hodgson (1976)
20 mg/kglday for 3 days
Mouse
Increased CYP2BlO, CYPIA2, CYP3A, CYP2BlO
Dai et at. (1998) (continues)
Table 23.5 (continued) Inducer
Dose
Species
Effect
Reference
20 mg/kg, 14 days
Rat
Increased in vitro metabolism of azinphosmethyI
Murphy and Dubois (1957)
Organophosphorus insecticides 3-Methylcholanthrene
to a cholinesterase inhibitor Phenobarbital
50 mg/kg/day, 5 days
Rat, mouse
In vivo toxicity of parathion, methyl parathion,
DuBois (1969)
demoton, disulfoton, azinphosmethyl, dioxathion, ethion, carbophenothion, mevinohos and EPN, all decreased S,S,S- Tri-n-butyl
phosphorotrithioare Carbamate insecticides Carbaryl
100-1000 mg/kg,
Hen
single dose 100--400 mg/kglday,
Increased P450 and p-chloro-N-methyl aniline
Lapadula et al. (1984)
demethylase activity Chicken
3-6 days, po
Pentobarbital sleeping time decreased, no effect
Puyear and Paulson (1972)
at 10-50 mg/kg/day, 3-6 days
Rodenticides
... tI'I
Phenobarbital
Q\
2 mg/kg, 2 weeks
Human
Pharmacological activity of warfarin decreased
Robinson and MacDonald (1966)
10 mg/kglday, 8 days
Dog
Dicoumarol activity in vivo decreased
Welch et al. (1967)
Acetylsalicylic acid
100 g/kg, single dose
Rat
Dicoumaral prothrombin time decreased
ColdwelI and Zawidzka (1968)
Heptobarbital
400 mg/day, 4 or 8 days
Human
Excretion of dicoumarol metabolites increased
Aggeler and O'Reilly (1969)
100-2000 ppm,
Rat
Detoxication of EPN, O-demethylation of p-nitro-
Kinoshita and DuBois (1967)
Herbicides Monuron
13 weeks, po Diuron
100-2000 ppm,
anisole and N -demethylation of aminopyrine increased for 1-3 weeks, then returned to normal Rat
13 weeks, po
Detoxication of EPN, O-demethylation of p-nitro-
Kinoshita and DuBois (1967)
anisole and N -demethylation of aminopyrine increased for 1-3 weeks, then returned to normal
Atrazine
430 mg/kg, single dose
Rat
Increase in in vivo hexobarbital elimination, increase
Ugazio et al. (1993)
in 7-pentoxyresorufin O-dealkylase activity Tridiphane Fungicides Griseofulvin
1-2 glday, ip, 3 days
Mouse
Increased lauric acid hydroxylase activity
Levi et al. (1992)
1-2 g/day
Human
Pharmacological action of warfarin decreased
Cullen and Catalano (1967)
Parnon
15-100 ppm in diet
Rat
In vitro metabolism increased
Hoffman et al. (1968, 1970)
Captan
7.5 and 15 mg/kg ip
Mouse
Induces CYP3A in kidney, CYP 1A2 in kidney and lung and 2a-testosterone hydroxylation in male liver
Paolini et al. (1999)
23.4 Chemical Factors Affecting Metabolism
Work by Abernathy et al. (1971a, b) demonstrated significant decreases in zoxazolamine paralysis time, hexobarbital sleeping time, and aniline hydroxylase activity in mice following treatment with DDT or DDE, a major metabolite of DDT and an important contaminant of animal fat. Different inducers may activate different enzymes and, therefore, different metabolic pathways. Thus, Chadwick et al. (1971) showed that repeated doses of lindane and DDT increased oxidative hydrolysis, O-demethylase, dehydrochlorinase, and glucuronyl transferase activity, but to different degrees. Pretreatment of rats with lindane caused them to metabolize a single dose of radioactive lindane 2.5 times more than controls, and pretreatment with DDT caused a 3.5-fold increase in metabolism of radioactive lindane. Furthermore, the DDT pretreatment was followed by proportionally more neutral and weakly polar, but less free-acid-type metabolites of the radioactive lindane. Thus, metabolism was qualitatively different following administration of the two inducers. Subsequent studies (Chadwick and Freal, 1972) confirmed these findings, including the increased excretion of metabolites following pretreatment with DDT. In addition, it was shown that rats pretreated with DDT plus lindane excreted more 2,4,5-trichlorophenol and 2,3,4,6- and 2,3,4,5-tetrachlorophenols by the second day of treatment that did rats receiving lindane alone. The results suggested that DDT treatment stimulates the metabolism of lindane through a selective effect on certain metabolic pathways involved in the oxidative degeneration of lindane, notably those leading to the formation of tetrachlorophenols, particularly 2,3,4,5-tetrachlorophenol. Incidentally, when two inducers are involved, the resulting induction may either be additive or slightly antagonistic. Thus, Gielen and Nebert (1971) found an additive effect when either phenobarbital or p,p'-DDT was present with a polycyclic hydrocarbon, but not when combinations of phenobarbital plus DDT, or one polycyclic hydrocarbon plus another were involved. Although it had been known for many years that various pesticides could induce P450, neither the specific isozymes induced nor the implications of the induction were well characterized. Later studies using enzymatic and immunochemical techniques examined isozyme specificity of specific pesticides. For example, mirex and chlordecone were shown to induce CYP2B lOin mouse liver, a pattern of induction similar to that of phenobarbital (Lewandowski et aI., 1989). Enzymatic activities suggested that, in addition to 2B 10, other P450s were induced and later studies demonstrated induction of lA2 and 3A (Dai et aI., 1998; Hodgson and Levi, 1996). Another group of pesticides, the phenoxyacetic acid herbicides (e.g., 2,4-D) and the herbicide synergist tridiphane were found to induce the CYP4A isozymes in rodents (Levi et aI., 1992; Moody et aI., 1991b, 1992). These P450 forms are known to be involved in the oxidation of fatty acids and the maintenance of lipid homeostasis. Moreover, in rodents, compounds that are CYP4A inducers also cause peroxisome proliferation, an event associated with nongenotoxic induction of liver tumors in rodents. Another peroxisome prolifer-
547
ate, the fenvalerate metabolite fenvaleric acid has been shown to induce several P450-dependent enzyme activities including 7-ethoxyresorufin-deethylation, catalyzed by CYP1As,7 -pentoxyresorufin O-dealkylation, catalyzed by CYP2Bs, and testosterone hydroxylation, catalyzed by CYP3A and CYP2B 11 (Morisseau et aI., 1991). The exact relationship of these interactions and the relevance to humans has not yet been defined. Methylenedioxyphenyl (MDP) compounds, such as piperonyl butoxide (PBO) and sesamex (SES), have been used as synergists with pyrethroid and carbamate pesticides. Other well-known MDP compounds such as safrole and isosafrole are found in many common foods of plant origin. These chemical compounds affect multiple enzyme pathways, including the P450 system. Their effect on P450 enzymes is biphasic, that is, inhibition followed by induction, and is discussed more fully in Section 23.4.5. Reviews include those of Philpot and Hodgson (1971-1972), Hodgson and Levi (1998), and Hodgson (1999). However, MDP compounds are known to induce both enzyme and mRNA for several CYP isoforms in the mouse, including CYP1A1, CYPIA2, and CYP2BlO. CYPIA2 is induced by both Ah-receptor-dependent and Ah-receptor-independent mechanisms. The fungicide captan, although inhibiting many hepatic P450-dependent activities in mouse liver (Paolini et aI., 1999), induces both CYP3A and CYPIA2 in the kidney and CYPIA2 in the lung. The ergosterol biosynthesis inhibiting fungicides (EBIFs), for example, clotrimazole and propioconazole, have been shown to have multiple effects on the rodent P450 system (Ronis et aI., 1994). The EBIFs induced P450s 3A, 2B, and lA, while suppressing the activity of 2Cll. These alterations were found to have significant changes in testosterone metabolism in male rats. Cellular techniques are becoming more available for the study of induction by pesticides. DuBois et al. (1966) used hepatocytes from rat and quail as well as human hepatoma (Hep G2) cells to study induction of P450 isoforms by pesticides. The pesticides fell into four groups: first, CYP3A inducers such as pentachlorophenol; second, 3-methy1cholanthrene-type inducers, such as lindane, an inducer of CYP1A isoforms; third, phenobarbital-type inducers, such as dieldrin, an inducer of CYP2B isoforms; fourth, pesticides with little or no capacity to induce P450 isoforms. Pentachlorophenol and lindane were the strongest inducers in these cell lines and lindane appeared to be a member of both the second and third groups because it induced both CYPIA and 2B activities. Of particular concern is the ability of many pesticides to disrupt the normal functioning of the endocrine system, and this effect has become an important environmental concern (Birnbaum, 1994; Colborn et aI., 1993; Guillette et aI., 1994). It is now quite clear that pesticides may have complex effects on the hepatic monooxygenase system and, in addition to affecting xenobiotic metabolism, pesticides, by disturbing endogenous metabolism, have the potential to result in profound changes in both the physiological and reproductive capacities ofthe organism (Birnbaum, 1994; Tyler et aI., 1998).
548
CHAPTER 23
Metabolism of Pesticides
23.4.3 INDUCTION OF OTHER ENZYMES Microsomal enzymes are not the only ones in the liver subject to induction. 8-Aminolevulinic acid (ALA) synthetase is located in the mitochondria and may increase 40-100 times in those structures on induction (Granick, 1965). Interaction between induction of mitochondrial and microsomal enzymes is illustrated by the action of the pesticide m-dichlorobenzene in rats. Following daily doses at the rather high rate of 800 mglkg, there is a biphasic stimulation of ALA-synthetase activity and of the excretion of urinary coproporphyrin, both of which peak by 3 days and then decline. The decrease in ALA-synthetase and in excretion of coproporphyrin at 5 days corresponds with the maximal stimulation of drug metabolism and with a decrease in the concentration of m-dichlorobenzene in the serum and liver at the time (Poland et aI., 1971). In the supernatant fraction of homogenized rat liver, the activity of NAD-dependent aldehyde dehydrogenase (EC 1.2.1.3) is increased up to lO-fold after administration of phenobarbital for 3 days. The effect is genetically controlled and is inherited as an autosomal dominant characteristic. The mechanism is apparently unrelated to other drug-induced increases in enzyme activity such as those that occur in the hepatic microsomal systems for drug metabolism (Deitrich, 1971). Glutathione S-transferases as well as cytochrome P450 were shown to be induced by pesticides but the levels of induction of the former were much lower (Fabacher et aI., 1980; Hodgson et aI., 1980; Robacker et aI., 1981). 23.4.4 INHIBITION OF ENZYME ACTIVITY As previously indicated, inhibition of xenobiotic-metabolizing enzymes can cause either an increase or a decrease in toxicity. Several well-known inhibitors of such enzymes are shown in Fig. 23.1 and are discussed in this section. Inhibitory effects can be demonstrated in a number of ways at different organizational levels. 23.4.4.1 Types oflnhibition: Experimental Demonstration
In vivo Symptoms The measurement of the effect of an inhibitor (or inducer) on the duration of action of a drug in vivo is a common method of demonstrating its action, and previously the effects on the hexobarbital sleeping time or the zoxazolamine paralysis time were often used. Both of these drugs are fairly rapidly deactivated by the hepatic microsomal cytochrome P450-dependent monooxygenase system; thus, inhibitors of the P450 isozyme(s) involved prolong their action whereas inducers have the opposite effect. The well-known inhibitor of P450 metabolism, SKF-525A, increases both hexobarbital sleeping time and zoxazolamine paralysis time in rats and mice, as do the insecticide synergists piperonyl butoxide and tropital, the optimum pretreatment time being about 0.5 hr
o (CeHs)2djO(CH2bN(CeHsb
~H-,
SKF-525A [P450]
O~C:JH-,
<~ 0;:'" OH2 (OC3 H4 )2 OC4Hg Piperonyl Butoxide [P450]
~~'N ~N' I
NH2 1-Amlnobenzotriazole
Allylisopropylacetamide [P450]
[P450]
oN~
o &iCH3)2-(jI
N
CeHs", ~
~HsOFO~N02
Metyrapone
EPN
[P450]
[esterases]
o
0
G.2HsOCcH :=CHfuc2 Hs Diethyl maleate [glutathione transferases]
S S (C2HsbN&S~N(C2Hs)2 Disulluram (Antabuse) [aldehyde dehydrogenase]
Figure 23.1
before the narcotic is given. As a consequence of the availability of single, expressed, isoforms for direct studies of inhibitory mechanisms, these methods are now used much less often. In the case of activation reactions, such as the activation of the insecticide azinphosmethyl to its potent anticholinesterase ox on derivative, a decrease in toxicity is apparent when rats are pretreated with the P450 inhibitor SKF-525A. Distribution and Blood Levels Treatment of an animal with an inhibitor of xenobiotic metabolism may cause changes in the blood levels of an unmetabolized toxic ant and/or its metabolites. This procedure may be used in the investigation of the inhibition of detoxication pathways; it has the advantage over in vitro methods of yielding results of direct physiological or toxicological interest because it is carried out in the intact animal. Moreover, the time sequence of the effects can be followed in individual animals, a factor of importance when inhibition is followed by induction. Effects on Metabolism in vivo A further refinement of the previous technique is to determine the effect of an inhibitor on the overall metabolism ofaxenobiotic in vivo, usually by following the appearance of metabolites in the urine and feces and/or in blood or other tissue. Again, the use of the intact animal has practical advantages over in vitro methods, although little is revealed about the mechanisms involved. Effects on in vitro Metabolism Following in vivo Treatment This method of demonstrating inhibition is of variable utility. The preparation of enzymes from animal tissues usually involves considerable dilution with the preparative medium during homogenization, centrifugation, and resuspension. As a result, inhibitors not tightly bound to the enzyme in
23.4 Chemical Factors Affecting Metabolism
question are lost, either in whole or in part, during the preparative processes. Therefore, negative results can have little utility because failure to inhibit and loss of the inhibitor give identical results. Positive results, on the other hand, not only indicate that the compound administered is an inhibitor but also provide a clear indication of excellent binding to the enzyme, most probably due to the formation of a covalent or slowly reversible inhibitory complex. The inhibition of acetylcholinesterase following treatment of the animal with organophosphorus compounds, such as paraoxon, is a good example, because the phosphorylated enzyme is stable and is still inhibited after the preparative procedures. Inhibition by carbamates, however, is greatly reduced by the same procedures, because the carbamylated enzyme is unstable and, in addition, the residual carbamate is highly diluted. Microsomal monooxygenase inhibitors that form stable inhibitory complexes with P450, such as SKF-525A, piperonyl butoxide and other methy lenedioxyphenyl compounds, and amphetamine and its derivatives, can be readily investigated in this way because the microsomes isolated from pretreated animals have a reduced capacity to oxidize many xenobiotics.
In vitro Effects In vitro measurement of the effect of one xenobiotic on the metabolism of another is by far the most common type of investigation of interactions involving inhibition. Although it is the most useful method for the study of inhibitory mechanisms, particularly when purified enzymes are used, it is of more limited utility in assessing the toxicological implications for the intact animal. The principal reason for this is that in vitro measurement does not assess the effects of factors that affect absorption, distribution, and prior metabolism, all of which occur before the inhibitory event under consideration. The primary considerations in studies of inhibition mechanisms are reversibility and selectivity. The inhibition kinetics of reversible inhibition give considerable insight into the reaction mechanisms of enzymes and, for that reason, have been well studied. In general, reversible inhibition involves no covalent binding, occurs rapidly, and can be reversed by dialysis or by dilution. Reversible inhibition is usually divided into competitive inhibition, uncompetitive inhibition, and noncompetitive inhibition. Because these types are not rigidly separated, many intermediate classes have been described. Competitive inhibition is usually caused by two substrates competing for the same active site. Following classical enzyme kinetics, there should be a change in the apparent Km but not in Vmax . In microsomal monooxygenase reactions, type I ligands, which often appear to bind as substrates but do not bind to the heme iron, might be expected to be competitive inhibitors, and this frequently appears to be the case. Uncompetitive inhibition has seldom been reported in studies of xenobiotic metabolism. It occurs when an inhibitor interacts with an enzyme-substrate complex but cannot interact with free enzyme. Both Km and Vmax change by the same ratio, giving rise to a family of parallel lines in a Lineweaver-Burke plot.
549
Noncompetitive inhibitors can bind to both the enzyme and enzyme-substrate complex to form either an enzymeinhibitor complex or an enzyme-inhibitor-substrate complex. The net result is a decrease in Vmax but no change in Km. Metyrapone (Fig. 23.1), a well-known inhibitor of monooxygenase reactions, can also, under some circumstances, stimulate metabolism in vitro. In either case, the effect is noncompetitive, in that the Km does not change whereas Vmax does, decreasing in the case of inhibition and increasing in the case of stimulation. Irreversible inhibition, which is much more important toxicologically, can arise from various causes. In most cases, the formation of covalent or other stable bonds or the disruption of the enzyme structure is involved. In these cases, the effect cannot be readily reversed in vitro by either dialysis or dilution. The formation of stable inhibitory complexes may involve the prior formation of a reactive intermediate that then interacts with the enzyme. An excellent example of this type of inhibition is the effect of the insecticide synergist piperonyl butoxide on hepatic microsomal monooxygenase activity, reviewed by Hodgson and Levi (1998) and Hodgson (1999). This methylenedioxyphenyl compound can form a stable inhibitory complex that blocks CO binding to P450 and also prevents substrate oxidation. This complex results from the formation of a reactive intermediate, and the type of inhibition changes from competitive to irreversible as metabolism, in the presence of NADPH and oxygen, proceeds. It appears probable that the metabolite in question is a carbene formed spontaneously by elimination of water following hydroxylation of the methylene carbon by the cytochrome (see Table 23.1 for metabolism of methylenedioxyphenyl compounds). Piperonyl butoxide inhibits the in vitro metabolism of many substrates of the monooxygenase system, including aldrin, ethylmorphine, aniline, aminopyrene, carbaryl, biphenyl, hexobarbital, and p-nitroanisole. The inhibition, by organophosphorus compounds such as (EPN), of the carboxylesterase that hydrolyzes malathion is a further example of xenobiotic interaction resulting from irreversible inhibition, because in this case the enzyme is phosphorylated by the inhibitor. Oxidative desulfuration of phosphorothioate pesticides such as parathion and fenitrothion by P450s is known to release atomic sulfur, which covalently binds to and inactivates P450s (Halpert et aI., 1980; Kamataki and Neal, 1976; Levi et aI., 1988; Neal and Halpert, 1982). In a recent study, administration of fenitrothion at a dose as low as 7 mg/kg inhibited the hydroxylation of 17,B-estradiol by hepatic microsomes (Berger and Sultatos, 1996). These results suggest that, in low concentrations, organophosphates have the potential to inhibit enzymes important in normal sexual development. Another class of irreversible inhibitors of toxicological significance consists of those compounds that bring about the destruction of the xenobiotic-metabolizing enzymes. The drug allylisopropylacetamide (Fig. 23.1), as well as other allyl compounds, has long been known to cause the breakdown of P450 and the resultant release of heme, the hepatocarcinogen vinyl
550
CHAPTER 23
Metabolism of Pesticides
chloride has also been shown to have a similar effect, probably also mediated through the generation of a highly reactive intermediate.
23.4.4.2 Synergism and Potentiation The terms synergism and potentiation have been variously used and defined but, in any case, involve a toxicity that is greater when two compounds are given simultaneously or sequentially than would be expected from a consideration of the toxicities of the compounds given alone. An example of synergism has already been mentioned. Piperonyl butoxide, sesamex, and related compounds increase the toxicity of insecticides to insects by inhibiting insect P450. Other insecticide synergists that interact with P450 include aryloxyalkylamines such as SKF-525A, Lilly 18947, and their derivatives, compounds containing acetylenic bonds such as aryl-2-propynyl phosphate esters containing propynyl functions, phosphorothionates, benzothiadiazoles, and some imidazole derivatives. The best known example of potentiation involving insecticides and an enzyme other than the monooxygenase system is the increase in the toxicity of malathion to mammals that is brought about by certain other organophosphates. Malathion has a low mammalian toxicity due primarily to its rapid hydrolysis by a carboxylesterase. EPN (Fig. 23.1), another organophosphate insecticide, causes a dramatic increase in malathion toxicity to mammals at dose levels that, given alone, cause essentially no inhibition of cholinesterase. In vitro studies have shown that the oxygen analog of EPN, as well as oxons of many other organophosphate compounds, increases the toxicity of malathion by inhibiting the carboxylesterase responsible for its degradation.
23.4.4.3 Antagonism In toxicology, antagonism may be defined as that situation in which the toxicity of two or more compounds administered together, or sequentially, is less than would be expected from a consideration of their toxicities when administered individually. Apart from the effects mediated through induction of xenobiotic-metabolizing enzymes (discussed previously), antagonism does not appear to be important in pesticide interactions.
23.4.4.4 Pesticides as Inhibitors Pesticides may act to inhibit P450 or other enzymes by any of the mechanisms discussed previously-competitive inhibition, noncompetitive inhibition, and suicide inhibition. As discussed more fully in the following section methylenedioxyphenyl compounds have very complex interactions with the P450 system, being both inhibitors and inducers. Because MDP compounds are substrates for P450 enzymes, they may act initially as competitive inhibitors. As the MDP compound is metabolized, it becomes a suicide inhibitor with its reactive metabolite bound to the heme iron ofP450 (Goldstein et aI., 1973; Hodgson et aI., 1998; Hodgson and Philpot, 1974).
The herbicide synergist tridiphane, a postemergent herbicide, owes its activity to its ability to inhibit glutathione S-transferases. It has also been shown to induce epoxide hydrolase and P450, specifically CYP4A, and peroxisomal enzymes (Levi et aI., 1992; Moody and Hammock, 1987). In addition to induction of 4A, tridiphane functions as a selective P450 inhibitor, inhibiting CYP2B 10 while having little or no effect on other P450 isozymes (Moreland et aI., 1989). As assessed by in vitro studies, tridiphane appears to be a competitive inhibitor of P450; its effect in vivo, however, is not yet known. Organophosphorus insecticides such as parathion that contain the P=S moiety are metabolized by the P450 system to the corresponding oxon, P=O, by oxidative desulfuration. This activation reaction, which converts the relatively inactive compound to a potent cholinesterase inhibitor, is thought to involve the formation of a P-S-O (phosphooxythirane) ring intermediate (Table 23.1). Studies with both microsomes and purified enzymes (Halpert et aI., 1980; Kamataki and Neal, 1976; Morelli and Nakatsugawa, 1978) have demonstrated that, during oxidative desulfuration, the released sulfur exists as a highly reactive molecule which then binds to P450, inactivating the enzyme. This binding of reactive sulfur to P450 is accompanied by loss of P450 as detected by measurement of the dithionitereduced CO complex as well as loss of monooxygenase activity (Berger and Sultatos, 1996; Butler and Murray, 1993; Neal, 1985; Neal and Halpert, 1982; Neal et aI., 1983). Cohen (1984) showed that acetaminophen toxicity was reduced by the organophosphorus insecticide fenitrothion, as a result of inhibition of the P450-dependent activation of acetaminophen. Studies with purified P450s and fenitrothion demonstrated that the amount of inhibition varied with the P450 isozyme (Levi et aI., 1988). In human liver microsomes, metabolism of parathion resulted in a concurrent loss of total P450 as well as the loss of several P450-mediated enzyme activities (Butler and Murray, 1997). The activities inhibited included testosterone 6,B-hydroxylation, catalyzed by CYP3A4, 7-ethoxyresorufinodeethylation, catalyzed by CYP1A2, and tolbutamide methyl hydroxylation, catalyzed by CYP2C9110. Aniline 4-hydroxylation, catalyzed by CYP2E1, was not inhibited. Organophosphorus compounds may also inhibit enzymes other than P450, particularly esterases (e.g., Cohen, 1984; Gaughan et aI., 1980). Gaughan et al. (1980) showed that profenofos, EPN, and (DEF), when administered in vivo to mice, all inhibited the liver microsomal esterases hydrolyzing trans-permethrin as well as the carboxylesterase hydrolyzing malathion. The fungicide captan, apparently through reactive metabolites, inhibits several P450-dependent enzyme activities in mouse liver (Paolini et aI., 1999) although it induces the 2ahydroxy1ation of testosterone. Methoxychlor, again through a reactive intermediate, inhibits the oxidation of both testosterone and estradiol, the pattern of metabolites indicating inhibition of CYP2C11 in rats and CYP3A in humans (Li et aI., 1993). It may also be noted that nonpesticidal inhibitors of P450 isoforms may affect the metabolism and toxicity of pesticides.
23.5 Physiological Factors Affecting Enzyme Activity
For example, Agyeman and Sultatos (1998) showed that the H2-blocker cimetidine caused a moderate increase in the toxicity of parathion but did not affect the toxicity of paraoxon, an effect brought about by the inhibition of P450 isoforms. 23.4.5 BIPHASIC EFFECTS: INHIBITION AND INDUCTION
Many inhibitors of mammalian monooxygenase activity can act also as inducers. Generally, inhibition of microsomal monooxygenase activity is fairly rapid and involves a direct interaction with the cytochrome, whereas induction is a slower process. Therefore, following a single injection an initial decrease due to inhibition is followed by an inductive phase. As the compound and its metabolites are eliminated, the levels of activity return to control values. Some of the best examples of compounds showing such biphasic effects are the methylenedioxyphenyl compounds, such as the pesticide synergists piperonyl butoxide and sesamex, and the secondary plant compounds safrole and isosafrole. The effect of MDP compounds on P450 is an initial inhibition of activity followed by an increase above control levels (Kamienski and Murphy, 1971; Kinsler et al., 1990; Philpot and Hodgson, 1971-1972). The inhibitory effect of MDP compounds has been attributed to the formation of a stable inhibitory metabolite complex between the heme iron of the P450 and the carbene species formed when water is cleaved from the hydroxylated methylene carbon of the MDP compound (Dahl and Hodgson, 1979). Because P450 combined with MDP compounds in an inhibitory complex cannot interact with CO, the cytochrome P450 titer, as determined by the method of Omura and Sato (1964) (dependent upon CO-binding to reduced cytochrome), reflects the biphasic effect. MDP exposure induces several hepatic P450 isozymes not normally found in detectable quantities in the livers of unexposed animals, including 1A1 and 2BlO in mice as well as the constitutive isozyme 1A2 (Adams et aI., 1993a, b; Lewandowski et aI., 1990; Ryu et aI., 1995, 1996, 1997; Ryu and Hodgson, 1999). A number of studies have been published regarding the effects of MDP compounds on mammalian liver enzymes (for review see Adams et aI., 1995; Hodgson et aI., 1995a, b). It is apparent from extensive reviews of the induction of monooxygenase activity by xenobiotics that many compounds other than methylenedioxyphenyl compounds have the same biphasic effect. It may be that any synergist that functions by inhibiting microsomal monooxygenase activity could also induce this activity on longer exposure, resulting in a biphasic curve as described previously for methylenedioxyphenyl compounds. This curve has been demonstrated for NIA 16824 (2-methylpropyl-2-propynyl phenylphosphonate) and WL 19255 (5,6-dichloro-1 ,2,3-benzothiadiazole), although the results were less marked with R05-8019 [2,(2,4,5-trichlorophenyl)-propynyl ether] and MGK 264 [N-(2-ethylhexyl)-5norbornene-2,3-dicarboxirnide].
:':'1
23.5 PHYSIOLOGICAL FACTORS AFFECTING ENZYME ACTIVITY Species, strains, and individuals may all vary in their susceptibility to toxicants, including pesticides. In some cases it has been possible to explain these differences by one of several causes, including differences in metabolism. In this section some examples are presented in which it is known that the activity of microsomal enzymes are influenced by age, gender, and species. 23.5.1 DEVELOPMENTAL EFFECTS
Microsomal enzyme activity is low or absent in the fetus and the newborn, but increases rapidly during the early days or weeks of life (Fouts and Adamson, 1959; Ronis and Cunny, 1994, 2000). For this reason fetuses and newborns are often more susceptible to certain drugs and xenobiotics than adults. With ageing, there is generaly a decrease in enzymatic activity, although increases in some activities have been observed (Kitahara et aI., 1982; Van Bezooijen, 1984; Van Bezooijen et aI., 1986). In the mouse, FM01 and 5 are expressed as early as gestation days 15 and 17 and equally between genders until puberty. FM03 is not expressed until 2 weeks postpartum and is found equally in male and female until 6 weeks postpartum, when it becomes undetectable in the male. This developmental pattern, as seen in the female mouse, is similar to humans of either gender (Cherrington et al., 1998a). 23.5.2 SPECIES DIFFERENCES
Many species differences in the metabolism of xenobiotics can be explained in terms of differences in activity of liver microsomal enzymes (Brodie and Maickel, 1962; Quinn et aI., 1958; Walker, 1994) although broad similarities often exist across large systematic groups. For example, Barron et al. (1993) observed that the detoxication of chlorpyrifos in channel catfish was similar to that in other vertebrates in both phase I and phase 11 metabolism. An example of a comparative study of a specific pesticide is that of Chin et al. (1979b) on carbaryl. The gender-dependent expression of FMO isoforms also varies between species, as outlined in Section 23.5.4. 23.5.3 INDIVIDUAL AND STRAIN DIFFERENCES
Strain and individual differences often are discussed conveniently in terms of "tolerance" and "resistance," both implying reduced susceptibility to a toxicant. The word tolerance is used when the observed decrease in susceptibility occurs in an individual organism as a result of its own previous or continuing exposure to the particular toxicant or to some other conditioning stimulus (see Section 23.6).
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Not only may differences in basic levels of enzyme activity be detected in different species, but this has also been accomplished in different strains of mice (Jay, 1955), rats (Quinn et aI., 1958), rabbits (Cram et aI., 1965), and birds (Ronis and Walker, 1989; Walker, 1983). Comparative aspects of xenobiotic metabolism, particularly as they relate to P450, have been reviewed by Hodgson (1979) and Kulkarni et al. (1975). 23.5.4 GENDER DIFFERENCES Metabolism of xenobiotics may vary with the gender of the organism and in some cases differences in overall toxicity between males and females of various species are known (Bonate, 1991). In the absence of induction, microsomal enzyme activity is often higher in the adult male rat than in females or immature males. However, the stimulatory effect of foreign chemicals on microsomal enzymes is usually greater in females and immature males than in the adult male rat (Conney and Bums, 1962). Gender differences become apparent at puberty and are usually maintained throughout adult life. The differences in microsomal monooxygenase activity between males and females have been shown to be under the control of sex hormones, at least in some species. Sexually dimorphic P450s appear to arise, in the rat, by programming, or imprinting, that occurs in neonatal development. This imprinting is brought about by a surge of testosterone that occurs in the male neonate and appears to imprint the developing hypothalamus so that in later development growth hormone is secreted in a gender-specific manner. This pattern of growth hormone production (pulsatile) and the higher level of circulating testosterone in the male maintain the expression of male-specific P450 isozymes such as 2C1l. On the other hand, a more continuous pattern of growth hormone secretion and the lack of circulating testosterone appear to be responsible for female-specific P450s such as 2C12 (Gonzalez, 1989; Hosteter et aI., 1987; Kobliakov et aI., 1991; Schenkman et aI., 1989). Gender-specific expression is seen also with the FMO enzymes. It has been known for some time that hepatic FMO activity is higher in female mice than males and that the lower levels in males result from testosterone repression (Dannan et aI., 1986; Duffel et aI., 1981; Falls et aI., 1997; Lemoine et aI., 1991; Wirth and Thorgeirsson, 1978). In addition, hormonal changes during pregnancy have been reported to increase FMO levels (Williams et aI., 1985). With regard to pesticide oxidation, gender differences have also been observed, with higher activity in female mouse liver than in male (Kinsler et aI., 1988). Recent studies have identified FMO isozymes involved in these gender differences and some of the hormonal factors involved in regulation. In several strains of mouse liver, FM01 expression was found to be two to three times higher in female mice compared to males, and FM03, expressed in females at levels comparable to FM01, was not detectable in male liver (Falls et aI., 1995, 1997; Cherrington et aI., 1998a). In rat liver, however, FM01 is higher in the male whereas FM03 is
gender-independent in both rat and human. FM05 is genderindependent for mouse, rat, and human (Cherrington et aI., 1998a). 23.5.5 GENETIC FACTORS The existence of discontinuous or biphasic variation is a strong indication of the possibility of genetic involvement. Some examples are given here in which such variation in enzyme activity within populations has been proved to be genetic in origin. The fungicide ziram causes hemolytic anemia with Heinz body formation in a man later shown to be deficient in erythrocyte G-6-PD. Ziram also caused one of the typical in vitro reactions (formation of Heinz bodies) in the blood of another person known to be deficient in this enzyme (Pinkhas et aI., 1963). 23.5.5.1 Humau Genetic Polymorphism Many of the phenotypic variations in drug responses observed in human populations have been shown to result from polymorphisms in the expression of the xenobiotic metabolizing enzymes. For reviews see Smith et al. (1994a, b), Kalow (1991), Coutts and Urichuk (1999), and Wormhaudt et al. (1999). Several of the human P450s have been shown to be polymorphic (Daly et aI., 1998; Goldstein and De Morais, 1994; Smith et aI., 1994a, b). Pesticides metabolized by these pathways would have higher or lower risk factors in some proportion of the exposed population.
23.6 TOLERANCE AND RESISTANCE The terms resistance and tolerance refer to a relative insusceptibility of a population of organisms to the effects of a toxicant. If the genetic trait preexists in a population so that it is obvious when the population is first exposed to the toxic ant, this should be regarded as tolerance rather than resistance, the latter term being reserved for those cases in which the trait is brought to an observable level only through selection by exposure to the toxic ant. It is clear that many instances of resistance are based on differences in toxicant metabolism that distinguish the resistant from the nonresistant population. Certainly some features of metabolism are known to be determined genetically and the origin of resistance through selection always indicates a genetic mechanism. 23.6.1 TOLERANCE Tolerance to a compound is often the result of an organism's increased ability to metabolize the chemical subsequent to an initial exposure. This is true, for example, in connection with the pesticides nicotine (Werle and Uschold, 1948) and dieldrin (Wright et aI., 1972). In a few instances, it has been shown clearly that the increased metabolism responsible for tolerance was mediated by
23.6 Tolerance and Resistance
increased activity of the microsomal enzymes of the liver. It seems likely that the same explanation will hold in connection with some other instances of tolerance. As recorded in Table 23.5, pesticides frequently act as inducers of microsomal enzymes. Because activity of these enzymes usually leads to detoxication, it seems likely that many of the compounds listed in Table 23.5 are capable of producing tolerance under suitable conditions. Tolerance also may exist in situations in which it has been impossible to demonstrate any increased ability to metabolize the toxicant; finally, there are instances of tolerance for which the mechanism is not only unknown but unexplored. For example, rodents may develop true tolerance (as distinguished from bait shyness) to a number of rodenticides including arsenic oxide, zinc phosphide, strychnine, sodium fluoroacetate, ANTU, and norbormide (Lund, 1967). Certain populations of pine mice subjected to control with endrin lost susceptibility to the compound, but sublethal exposure conferred a degree of tolerance regardless of the past history of the population (Webb and Horsfall, 1967). 23.6.2 RESISTANCE Resistance in the toxicological sense is better known in insects and a variety of other pest species than in vertebrates. Hundreds of species with some public health importance are now resistant to one or more pesticides (Brown and Pal, 1971; Georghiou and Saito, 1983) and a much larger number of agricultural pests are resistant to some degree. Resistance to a particular compound does not involve an entire species but only the toxicant-stressed population of a limited area; nevertheless, resistance constitutes a serious public health and economic problem. Although many species are or have become resistant to one or more pesticides, it does not follow that every species is genetically capable of developing resistance to every poison. On the contrary, there is currently no way to predict that resistance to a particular compound is impossible for any given species. Some species are effectively tolerant when they first encounter a particular compound. In a few instances, it has been recognized that tolerance existed in one strain of a species before that strain was exposed, in spite of the fact that other strains were fully susceptible. More commonly, resistant strains are first recognized only after the parent population has been selected by killing off many of its susceptible members. In all three situations, a genotype of the organism exists that can cope successfully with the toxic ant. The rate at which selection progresses depends not only on the intensity of the selection pressure, but also on the duration of each generation. Therefore, it is not surprising that resistance has most often been observed among organisms such as bacteria or houseflies, characterized by vast numbers of individuals and a rapid rate of reproduction. However, resistance has also been observed in species with relatively small populations and relatively slow multiplication. Resistance of a vertebrate species to a pesticide apparently first was recognized in the 1960s and 1970s in connection with Norway rats exposed to warfarin. This phenomenon
553
was first reported from Scotland (Boyle, 1960) and subsequently from Denmark (Lund, 1964, 1967), England and Wales (Bentley, 1969; Drummond, 1966), the Netherlands (Ophof and Langeveld, 1969), Germany (Telle, 1971)), and the United States, specifically North Carolina and Idaho (Brothers, 1972; lackson and Kaukeinen, 1972). The early literature on the resistance of mammals to warfarin was reviewed by Lund (1967). Only a few points need to be recorded here. So far, resistance is known to occur in four species, the Norway and roof rats, the house mouse, and humans. In addition, in their original studies of coumarin compounds, Link and his students reported marked variation in susceptibility in rabbits as a Mendelian characteristic (Campbell et al., 1941). The exact mechanism of inheritance of resistance to warfarin is not clear. In humans, the facts are consistent with transmission by a single autosomal dominant gene (O'Reilly et aI., 1963), but only one kindred has been studied. In rats and mice, it seems that more than one gene is involved. The physiological basis of the resistance also is not clear and may be different in different instances. There is no evidence that resistant rats are more efficient in their use of vitamin K, but people resistant to warfarin were extremely sensitive to the antidotal action of the vitamin (O'Reilly et aI., 1963). In every instance studied, including humans, resistance extended to other coumarin anticoagulants and those based on indandione. Susceptibility to heparin is normal. Another early report of resistance among vertebrates involved mosquito fish collected from insecticide-contaminated waters near cotton fields (Vinson et aI., 1963). Further study revealed 2- to 500-fold levels of resistance to a variety of pesticides in mosquito fish and five other species of fish (Boyd and Ferguson, 1964a, b; Ferguson and Boyd, 1964; Ferguson et aI., 1964, 1965; Ferguson and Bingham, 1966a, b). Resistance to chlorinated hydrocarbon pesticides was found in three species of frog (Boyd et aI., 1963). The degree of resistance may be so great in some instances that resistant species can withstand enough poison to kill their predators. Ozbum and Morrison (1962) were the first to produce resistance to a pesticide in a mammal by selection under laboratory conditions. In mice selected by a single interperitoneal dose of DDT administered at 4 weeks of age, resistance in the ninth generation had increased by a factor of 1.7 as measured by the LD 50. Although the factor of 1.7 is small, about half of the susceptible mice withstood a dose that was uniformly fatal to control mice. Further study (Ozbum and Morrison, 1965) revealed that the selected and control colonies differed in their rates of oxygen consumption. The resistant mice were fatter than the susceptible ones and considerable evidence indicated that resistance depended on preferential deposit of DDT in the fat and consequently the avoidance of peak levels in sensitive tissues. The resistance was not specific for DDT but extended to lindane and dieldrin (Barker and Morrison, 1966). Success in the development of resistance in mammals in the laboratory has not been uniform; apparently some strains are not sufficiently heterozygous to respond to selection (Guthrie et aI., 1971).
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Metabolism of Pesticides
Thus many instances are known in which species or strains differ in their susceptibility to pesticides, the resistance arising through selection in the field. In other instances, it has been possible to produce resistance in the laboratory through selection. In many instances it has been possible to define the genetic mechanisms responsible for observed differences in the metabolism of the pesticides in question. Except in the case of insects, a genetic mechanism has been defined only rarely in connection with metabolism of pesticides. The following references are suggested on this subject: Dauterman (1994); Evered and Collins (1984); Georghiou and Saito (1983); Hayes et al. (1990).
23.7 CONCLUSIONS Knowledge of the metabolism of pesticides is essential for several reasons, including the development of more selective insecticides, and provides, in part, the fundamental basis for science-based risk assessments for human and environmental health. Until the turn of the century, and as a matter of necessity, this research was carried out almost exclusively on experimental animals and the results, particularly in the case of human health risk assessments, extrapolated to humans. Although much essential background will continue to be obtained from experimental animals, due to the new techniques of molecular biology, it is now possible to work directly on human enzymes. These same techniques through the study of genetic polymorphisms should enable us to identify populations at increased risk and should enable comparative studies to be carried out at the level of specific isoforms of the xenobioticmetabolizing enzymes involved. Thus the study of pesticide metabolism is entering a new, more molecular, era that will be fascinating as well as useful.
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CHAPTER
24 Absorption, Distribution, and Pharmacokinetics Kelly J. Dix Research Triangle Institute
Pesticides and other chemicals interact with biological systems, including people, in a many ways. Pesticides are inherently toxic. To understand how nontarget species (e.g., humans) exposure to a pesticide may result in toxicity, one must first understand how the pesticide enters, moves about in, and is eliminated from the body. More specifically, one must characterize the processes of absorption of the chemical after various routes of exposure (e.g., dermal, inhalation, oral), distribution of the chemical to different organs and tissues in the body, metabolism of the chemical, and elimination of the chemical and/or its metabolites from the body. Pharmacokinetics, the study of the absorption, distribution, metabolism, and elimination of chemicals, utilizes mathematical models to describe the time course of the chemical in the body. This chapter includes a general discussion of absorption, distribution, and pharmacokinetics.
24.1 INTRODUCTION For a pesticide to elicit toxicity, it must be transferred from the external site of exposure to the target site (e.g., organ, nucleic acid, receptor) and achieve a sufficiently high concentration in the target organ (Fig. 24.1). Absorption is the translocation of the pesticide from an external source of exposure to the bloodstream. Once in the blood, the chemical is distributed through the body and delivered to tissues, where it may leave the blood and enter the cells of the tissue or it may remain in the blood and simply pass through the tissue. In certain tissues such as the liver, the chemical may be effectively removed from the body by metabolism. Other tissues, such as kidney and lung, serve to eliminate xenobiotics from the body by excretion. Absorption, distribution, metabolism, and excretion, which are collectively termed disposition, are all factors that affect the concentration of a chemical in target tissues. Pharmacokinetics refers to the mathematical description of the time course of chemical disposition in the body. Metabolism and excretion are discussed in detail in other chapters of this work. This chapter focuses on absorption, distribution, and pharmacokinetics. Handbook of Pesticide Toxicology Volume I. Principles
24.2 FACTORS THAT INFLUENCE THE TRANSFER AND AVAILABILITY OF CHEMICALS IN THE BODY For the routes of pesticide exposure relevant to humans, the pesticide must cross one or more cell membranes to reach the bloodstream, then one or more additional cell membranes to leave the blood and enter tissues. The following discussion concerns the factors that influence the transfer of chemicals across biological membranes.
24.2.1 PROPERTIES OF CELL MEMBRANES Cell membranes (i.e., plasma membranes) consist of phospholipids and proteins (Fig. 24.2). The fluid and dynamic phospholipid bilayer, with polar head groups on the intracellular and extracellular surfaces and fatty acid chains filling the inner space, acts as a permeability barrier to water-soluble molecules. Proteins interspersed throughout the phospholipid bilayer mediate the transport of small water-soluble molecules into and out of the cell by forming pores or by acting as carriers. Molecules cross membranes by passive transport, which requires the expenditure of no energy, or by specialized transport systems. The ability of a chemical to cross various membrane barriers is determined by its physicochemical properties, which include lipophilicity, molecular size, and ionization.
24.2.2 TRANSPORT MECHANISMS 24.2.2.1 Passive Transport Passive transport occurs by simple diffusion or via pores in the plasma membrane (Fig. 24.2). Most lipophilic molecules cross membranes by simple diffusion in accord with Fick's first law of diffusion, which states that the rate at which a molecule diffuses across the plasma membrane is proportional to the concentration gradient, the membrane surface area, and the permeability coefficient of the molecule.
563
Copyright © 200 1 by Academic Press. All rights of reproduction in any fonn reserved.
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Water readily traverses the plasma membrane through pores, and may carry with it small hydrophilic solutes. The pores in most cells are approximately 4 A in diameter. In the kidney glomeruli, however, the pores may be as large as 70-80 A in diameter, which permits more efficient renal elimination of potentially toxic compounds. Weak organic acids and bases may cross plasma membranes by simple diffusion when they are
Dermal
iv, ip. se, im
nonionized. Ionized weak organic acids and bases, however, slowly permeate the plasma membrane through pores. According to the Br\'lnsted-Lowry theory, an acid is a proton donor and a base is a proton acceptor. The ratio of nonionized to ionized molecules of a weak organic acid or base depends on the dissociation constant (Ka) and the pH of the media (Table 24.l). The dissociation constant is usually expressed in terms of its negative logarithm, and the relationship between pKa and pH is derived from the Henderson-Hasselbalch equation as shown in Table 24.1. The pKa is the pH at which 50% of the acid or base is ionized. The concept of pKa is particularly important for oral absorption (see Section 24.3.1). 24.2.2.2 Specialized Transport
..Metabolites Kidney
..Excretion
Figure 24.1 Representation of the absorption, distribution, metabolism, and excretion of toxicants.
•
Extracellular Space
Active transport systems are characterized by (1) movement of solutes against a concentration or electrochemical gradient, (2) saturation at high solute concentration, (3) specificity for structural and/or chemical features of the solute, (4) competitive inhibition by molecules transported by the same transporter, and (5) inhibition of transport by compounds and/or processes that interfere with cellular metabolism. Facilitated diffusion is similar to active transport, except that the solute moves only in the direction of a concentration or electrochemical gradient and
•
•
Pho. pholipid Bilaycr Intracellular Space
Simple Diffu ion Figure 24.2
Diffu ion via Porc '
•
Facilitated Diffu ion
ctive Tran port
Schematic of the plasma membrane and mechanisms of transport across the membrane. Table 24.1 Acids and Bases According to the Brpnsted-LowlY Theory Acid
Base
Representation
AH .... A- + H+
B + H+ .... BH+
Definition
Proton donor (AH)
Proton acceptor (B)
Dissociation constant (Ka)
Ka =
pKa = log
f,; = -log Ka
[A-][H+] [AH]
pKa =PH+log~
_
Ka -
[B][H+ ] [BH+]
pKa = pH + log
[BUr] [
24.3 Absorption
the expenditure of energy is not required (Fig. 24.2). Additional types of specialized transport are exocytosis and endocytosis, processes by which cells secrete and ingest large molecules, respectively. There are two types of endocytosis; pinocytosis ("cell drinking"), which is the ingestion of fluids and solutes, and phagocytosis ("cell eating"), which is the ingestion of large particles. Phagocytosis is especially important in the removal of particulate matter in the respiratory tract.
24.2.3 PROTEIN (MACROMOLECULAR) BINDING Blood consists of red blood cells, white blood cells, and platelets suspended in plasma. Plasma, which comprises ca. 55% of the blood volume in humans, also contains a number of proteins, ions, and inorganic molecules. Many xenobiotics in blood are reversibly bound to plasma proteins, including albumin, oq -acid glycoprotein, lipoproteins, and globulins. Reversible binding to plasma proteins enhances the solubility of lipophilic compounds in blood and influences the rate of distribution to tissues. Proteins are amphoteric in nature and therefore possess cationic and anionic regions. Many acidic chemicals bind to albumin, whereas basic chemicals tend to bind to aI-acid glycoprotein and lipoproteins. The high molecular weight of proteins prevents them, and any toxicants they bind, from crossing cell membranes. Only the free (or unbound) chemical is available to cross plasma membranes (Fig. 24.3). The interaction of chemicals and plasma proteins, however, is rapid and reversible. Equilibrium is quickly established between the bound and the unbound forms of the chemical. As unbound chemical crosses a plasma membrane in a microenvironment, bound chemical dissociates to reestablish equilibrium with the unbound fraction. Gomez-Catalan et al. (1991) investigated the distribution of various organochlorines in rat and human blood. In rat blood, 87% of hexachlorobenzene was associated with red Blood
Tissue
Bound
Bound
o
cl
11 o
11
Free
~(==!==z)
0
565
blood cells, ca. 84% of DDE was bound to plasma proteins, lindane was nearly equally distributed between red blood cells and plasma, and 97% of pentachlorophenol was associated with plasma. In plasma, lindane (64%) and DDE (92%) were mainly associated with lipoproteins, pentachlorophenol was mainly associated with "other" plasma proteins (81 %), and hexachlorobenzene was nearly equally distributed. A very different pattern of distribution was observed in human blood. Hexachlorobenzene and lindane in plasma were nearly equally distributed between lipoproteins and "other" plasma proteins, whereas 60% of DDE was associated with "other" proteins. Other investigators have shown that dieldrin is greater than 99% bound to human serum proteins (Garrettson and Curley, 1969), and diflubenzuron is 40-50% bound to plasma proteins in chickens (Opdycke and Menzer, 1984). The organophosphate diazinon is 89% bound to proteins in rat plasma (Wu et aI., 1996).
24.3 ABSORPTION Human exposure to pesticides is typically by oral, dermal, and inhalation routes. Whereas occupational exposure to pesticides is likely to occur by inhalation or dermal contact, exposure of small children is more likely to be by oral and/or dermal routes. Absorption from the gastrointestinal and respiratory tracts is discussed in Sections 24.3.1 and 24.3.2, respectively. Dermal absorption is reviewed in Chapter 22. Other routes of exposure that are used primarily in the laboratory (subcutaneous, intravenous, intraperitoneal, and intramuscular) are discussed only briefly in Section 24.3.3. A chemical is considered to be absorbed when it reaches the bloodstream. For routes other than intravenous administration, which bypasses the process of absorption, a chemical is absorbed when it crosses a plasma membrane and enters the bloodstream from an external site of exposure. Continuous blood flow removes the xenobiotic from the site of absorption, thus maintaining a concentration gradient and enhancing continued absorption. For rapidly absorbed chemicals, equilibrium may be established between the blood and the site of absorption, and the rate of entry into the blood is limited by blood flow rather than by diffusion across the membrane. In this case, an increase in blood flow will increase the rate of absorption of the chemical and absorption is said to be perfusion- (or blood-flow) limited. For poorly absorbed chemicals, however, absorption is not sensitive to blood flow and is said to be diffusion-ratelimited.
24.3.1 ABSORPTION FROM THE GASTROINTESTINAL TRACT
Free
Figure 24.3 Equilibrium is established between free (unbound) and bound xenobiotic in blood and between free xenobiotic in blood and tissues. Only the free xenobiotic crosses the plasma membrane, which is represented by the dashed line, that separates the blood and tissue compartments.
People are potentially exposed to pesticides by oral exposure from pesticide residues in foods such as meat, milk, fruits, and vegetables. Children may also be orally exposed to pesticides when they place contaminated objects in their mouths. The rate and extent of absorption after oral exposure depends on the
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CHAPTER 24
Absorption, Distribution, and Pharmacokinetics
ability of the chemical to cross the plasma membranes of the gastrointestinal tract. As discussed in Section 24.2, diffusion across a plasma membrane depends to a large extent on the lipid solubility and degree of ionization of the chemical. The degree of ionization of weak acids and weak bases, and hence absorption, depends on pH. The pH range in the gastrointestinal tract varies from approximately 1-3 in the stomach to 6-8 in the intestines. Thus, the rate and extent of absorption of weak organic acids and bases varies with location in the gastrointestinal tract; weak acids are nonionized and are absorbed in the stomach, whereas weak bases are nonionized and are absorbed in the intestine (Fig. 24.4). Removal from the site of absorption by blood flow maintains a concentration gradient, thus enhancing absorption of chemicals. Residence time in the region of the gastrointestinal tract where the chemical is absorbed also affects absorption. The presence of food in the gut can alter the pH of the gut contents and the intestinal motility, which in turn can affect the rate of absorption from the gastrointestinal tract. The presence of stomach acid, gastric enzymes, and intestinal flora may decompose the chemical before absorption can occur, which may also decrease the potential for toxicity. Pekas (1972) demonstrated the hydrolysis of naphthyl N -methyl carbamate in the intestine (pH 6.4). The large surface area of the intestinal tract aids in absorption from this site; even chemicals that do not readily cross the plasma membrane (e.g., weak acids) can be absorbed to a high degree in the intestine because of the increased surface area. Particles may be absorbed in the intestines by endocytosis. Chemicals that are absorbed into the bloodstream from the gastrointestinal tract enter the portal circulation and are delivered directly to the liver, where they may be metabolized before reaching the systemic circulation. This "first-pass" metabolism decreases the systemic availability (bioavailability) of the parent compound. Some highly lipophilic compounds such as organochlorine pesticides are absorbed into the lymphatic sys-
Stomach contents (pH 1.0) Nonionized/ionized ratio favors absorption Blood (pH 7.4)
Nonionized [l]
Ionized [0.43]
Nonionized [l]
Ionized [460]
Intestinal contents (pH 5.3) Nonionizedlionized ratio does not favor absorption
Figure 24.4 Proportion of nonionized and ionized fonns of 2,4-D (pKa 2.6) in the stomach and intestinal contents (adapted from Hodgson et aI., 1991). Only the nonionized form crosses cell membranes.
tern in a manner similar to the absorption of nutritional fats (Turner and Shanks, 1980). Absorption into the lymphatic system bypasses delivery to the liver and the potential for first-pass metabolism.
24.3.2 ABSORPTION FROM THE RESPIRATORY TRACT A chemical must be in the form of a gas, vapor, or particulate (e.g., aerosol) to be absorbed in the respiratory tract. Although the anatomy of the respiratory tract varies widely within mammalian species, the respiratory system can be generally compartmentalized into the nasopharyngeal, tracheobronchial, and pulmonary regions (Kennedy and Valentine, 1994). The function of the nasopharyngeal region is to condition inspired air and to remove large inspired particles before they reach the tracheobronchial and pulmonary regions. The tracheobronchial region is lined by mucus-secreting and ciliated cells, which together make up the mucociliary escalator. The pulmonary region of the respiratory tract is the gas-exchange region, which consists of the respiratory bronchioles, alveolar ducts, and al veoli. Inspired gases and vapors may be absorbed throughout the respiratory tract, depending on their physicochemical properties, and the anatomy and physiology of the region. Inhaled gases and vapors diffuse across cell membranes in the direction of the concentration gradient until equilibrium is established (see also Section 24.2.2). The ratio of gas or vapor equilibrium concentrations in blood and air is termed the blood: air partition coefficient. Highly water-soluble and reactive gases and vapors tend to be absorbed in the mucus layer of the upper respiratory tract, whereas more lipophilic and nonreactive gases and vapors are absorbed from the deeper regions of the respiratory tract. The geometry, blood flow, and capacity for metabolism of the respiratory tract may also influence the rate and site of absorption of inhaled gases and vapors. Deposition of aerosols in the respiratory tract depends on a number of factors, including the physicochemical and aerodynamic properties of the aerosol and the geometry and airflow of the respiratory tract. Airflow (velocity) and turbulence decrease from the nasopharyngeal to the pulmonary region, and different mechanisms of deposition operate in these regions. Impaction is an important mechanism for deposition of particles larger than I I-lm in aerodynamic diameter in regions of the respiratory tract where air velocity and turbulence are high, such as airway bifurcations. Interception is an important mechanism of deposition for fibers and is dependent on fiber length rather than diameter. Sedimentation is settling due to gravity and is important for particles larger than 1 I-lm in aerodynamic diameter in regions of the respiratory tract where airways are small in diameter and airflow is low. Diffusion is an important mechanism of deposition for small particles throughout the respiratory tract, and particularly for particles <0.5 I-lm in the alveolar region where airflow is low. Particles are cleared from the respiratory tract in a number of ways, depending on the region of the respiratory tract.
24.4 Distribution In the nasopharyngeal region, particles are removed by nose wiping, nose blowing, sneezing, and swallowing. In the tracheobronchial region, particles are cleared by the mucociliary escalator and are ultimately swallowed or expectorated. In the pulmonary region, particles are removed by (1) dissolution and removal in the bloodstream or lymphatics, (2) alveolar macrophage phagocytosis and removal by the mucociliary escalator or lymphatics, and (3) direct penetration of epithelial membranes and absorption into tissue or blood. 24.3.3 ABSORPTION AFTER EXPOSURE BY OTHER ROUTES
Absorption after exposure via the dermal route is discussed elsewhere in this work. Intravenous administration, in which the chemical is introduced directly into the blood, by definition, bypasses the process of absorption. Other routes of exposure that are used in the laboratory include intraperitoneal injection, where chemicals are absorbed primarily through the portal circulation, and intramuscular or subcutaneous injection. Chemicals administered intramuscularly or subcutaneously are absorbed more slowly.
24.4 DISTRIBUTION Once in the bloodstream, the chemical is available for distribution throughout and elimination from the body. Metabolism and excretion, which are components of elimination, are discussed in other chapters. This section will focus on distribution, the reversible translocation of chemicals from one location to another in the body. Distribution of a toxic ant to and accumulation in the target organ may result in toxicity. Accumulation at nontarget sites, on the other hand, results in storage of the pesticide away from the site of action and ultimately protection from toxicity. The physiology of the organism and the physicochemical characteristics of the pesticide are important factors in the distribution of absorbed pesticides. 24.4.1 TOTAL BODY WATER
Chemicals in the body move throughout the water compartments of the body. As already discussed, the ability of chemicals to move between the various water compartments is limited by the physicochemical properties of the chemical. Total body water consists of plasma water, interstitial water, and intracellular water. In humans, approximately 60% of body weight is water, with plasma, intracellular, and interstitial water accounting for 5, 15, and 40% of body weight, respectively. Plasma water, which represents approximately 53-58% of blood volume in humans, plays an essential role in the distribution of absorbed chemicals. For a chemical to move from blood (plasma water) into tissues, it must cross the endothelial cell layer lining the capillaries (i.e., capillary wall) to enter the interstitial water, then
567
cross the plasma membrane to enter the intracellular water. Chemicals exist in blood as free circulating chemicals or are noncovalently bound to plasma proteins. The rates of association and dissociation with plasma proteins are very rapid (on the order of milliseconds) and it is assumed that the bound and free forms of the chemical are in equilibrium. The capillary wall is permeable to small molecules, but not readily permeable to high molecular weight molecules such as plasma proteins. Only free chemicals that are small enough to pass through the capillaries, then, are available to move from plasma water to interstitial water. The processes for crossing plasma membranes described in Section 24.2 govern passage from the interstitial water to intracellular water. 24.4.2 RATE AND EXTENT OF DISTRIBUTION
Factors that influence the rate and extent of distribution of a chemical to a particular tissue include blood flow to the tissue (rate of delivery), the mass of the tissue, the ability of the chemical to cross membranes, and the affinity of the chemical for the tissue relative to blood. The rate of distribution of a chemical from blood to tissues can be perfusion- or diffusionrate-limited. For lipophilic chemicals that rapidly cross membranes, the rate of delivery to tissues is limited by blood flow (perfusion-rate-limited). For polar and ionized chemicals that do not readily cross the plasma membrane, the rate of delivery to tissues is limited by diffusion (diffusion-rate-limited). Plasma protein binding increases the rate of distribution to tissues for toxicants that are not diffusion-rate-limited. The free toxicant may readily cross the capillary wall, effectively decreasing its free concentration in blood. Bound toxicant then dissociates from plasma proteins to maintain the equilibrium between the bound and free forms, yet the new free molecules rapidly leave the blood, which further increases dissociation of bound toxic ant, and so on. In contrast, distribution of more polar compounds that are diffusion-rate-limited is dependent on the extent of protein binding. Initial distribution is influenced primarily by blood flow to tissues, whereas final distribution is influenced primarily by the relative affinity of the chemical for various tissues relative to blood (i.e., the tissue partition coefficient). In the early phase of distribution, tissues that receive a high blood flow (e.g., liver, kidney, and brain) may achieve high concentrations of the chemical even though the tissue partition coefficient for that chemical is low. Likewise, tissues that are slowly perfused (e.g., adipose) may achieve a low concentration of the chemical in the early phase of distribution even though the tissue partition coefficient for that chemical is very high. Later in the distribution phase, however, the chemical redistributes to tissues based on tissue partition coefficients, and the chemical is more concentrated in tissues with relatively high partition coefficients. Pesticides and other xenobiotics do not have the same tissue partition coefficient for all tissues. For example, dimethoate has a relatively high affinity for liver, muscle, and brain (GarciaRepetto et aI., 1995), whereas the chlorinated insecticides DDT,
568
CHAPTER 24
Absorption, Distribution, and Pharmacokinetics
aldrin, and dieldrin are lipophilic and have high affinities for adipose tissue (Lehman, 1956; Robinson et aI., 1969). 24.4.3 VOLUME OF DISTRIBUTION
When a chemical is absorbed and distribution is complete, its concentration in blood depends on the amount absorbed and the extent of tissue distribution. The apparent volume of distribution (Vd) is a proportionality constant that relates the amount of chemical in the body to its concentration in plasma, amount in body Vd= ------------~-concentration in plasma where Vd is the theoretical volume of fluid the chemical would occupy to achieve the observed concentration in plasma and does not necessarily correspond to the volume of a particular body fluid compartment. For example, a chemical that is sequestered in a particular tissue will have a low concentration in plasma and a corresponding high volume of distribution, which may in fact be greater than the total body water. 24.4.4 BLOOD-BRAIN BARRIER
The blood-brain barrier, which protects the central nervous system, is not an absolute barrier. 2,4-Dichlorophenoxyacetic acid (2,4-D), for example, has been measured in the rabbit brain after an intravenous (IV) dose (Kim et aI., 1996). The tight junctions of the capillary endothelial cells and the surrounding glial cell processes are the main structural features that contribute to the low permeability of the blood-brain barrier. Chemicals that circulate in blood must pass through the capillary endothelial cell membrane and the glial cell membrane to access the interstitial fluid of the brain. The low protein content of the brain's interstitial fluid limits lipophilic chemicals that are highly bound to plasma proteins. Chemical access to the brain, then, is limited to those that are free (unbound), lipophilic, nonionized, and those that are transported by specialized carrier systems, whereas ionized and highly plasma protein bound chemicals are excluded by the blood-brain barrier. Another barrier to brain access is the presence of an adenosine 5' -triphosphate (ATP)-dependent multidrug-resistance (MDR) protein, which transports intracellular chemicals back into the extracellular space. Whereas the blood-brain barrier is not fully developed at birth, the risk of toxicity from exposure to some chemicals is higher for newborns and young children than for adults. 24.4.5 PLACENTAL TRANSFER
Functions of the placenta include delivery of nutrients to the fetus, removal of fetal waste, and maternal/fetal blood gas exchange, which suggests that many chemicals move freely across the placental membrane. In the framework of distribution, the placenta is not a barrier to protect the fetus from exposure to toxicants. Rather the placenta is a typical plasma membrane
barrier that is permeable to lipophilic and nonionized molecules that readily cross plasma membranes, thereby exposing the fetus to toxicants. DDE and 2,4-D have been demonstrated to cross the placenta in rats, rabbits, and bats (Kim et aI., 1996; Sandberg et aI., 1996; Thies and McBee, 1994), and prenatal human exposure to 2,4-D has been associated with mental retardation in offspring (Casey and Collie, 1984). 24.4.6 STORAGE AND REDISTRIBUTION
Chemicals may accumulate in body compartments due to protein binding, active transport processes, or high solubility in (i.e., affinity for) a particular tissue. These sites of accumulation can be considered storage depots. Whereas a chemical in any tissue compartment is in equilibrium with its free concentration in blood, storage is dynamic. Removal of free chemical from the body by metabolism or excretion shifts the equilibrium such that stored chemical is released. 24.4.6.1 Plasma Proteins
Plasma protein binding plays a very important role in chemicalinduced toxicity. Displacement of one chemical from plasma proteins by another chemical can have severe consequences. If the bound chemical is very toxic, its displacement results in a higher free concentration in plasma, which results in greater availability for distribution to its site of toxic action. 24.4.6.2 Fat
Adipose is a storage depot for a number of highly lipophilic chemicals, including pesticides. Storage in adipose tissue may be considered a protective mechanism in that the pesticide is stored in a nontarget tissue, thereby lowering its concentration at the site of toxic action. For example, the chlorinated insecticides DDT (Dale et aI., 1962; Hayes et aI., 1958), chlordane (Ambrose et aI., 1953), hexachlorobenzene isomers (Davidow and Frawley, 1951), lindane (Ludwig et aI., 1964), aldrin, and dieldrin (Robinson et aI., 1969) are lipophilic and accumulate in fat. Upon dieting and starvation, fat is mobilized and the stored chemical is released, which results in a sudden increase in the blood concentration of the pesticide and availability of the pesticide for redistribution. As was described for plasma proteins, chemicals stored in adipose tissue may be displaced by other chemicals. Street (1964) demonstrated that DDT displaces dieldrin from its storage sites in rat adipose, yet methoxychlor does not affect dieldrin storage. Toxicity may be observed if the released chemical is redistributed to the target organ. 24.4.6.3 Other Tissues and Tissue Components
Sequestration in tissues (e.g., kidney and liver) may be due to interaction of chemicals with tissue macromolecules such as proteins and nucleic acids, which influence the affinity of a tissue for a given chemical (tissue partition coefficient). Bone tissue, for example, is a potential storage depot for heavy metals.
24.5 Pharmacokinetics
24.4.7 STORAGE WITH REPEATED EXPOSURE
Body burden is the term for the concentration (or amount) of chemical in the body at any given time, and the biological half-life of a chemical is the time required to reduce the concentration of the chemical in the body by one-half, in the absence of further intake. Many pesticides are water-soluble and easily excreted or readily metabolized to more water-soluble compounds that are easily excreted. Lipophilic pesticides such as the organochlorines, however, are stored in fat and are not easily removed from the body, and most people around the world carry a low body burden of organochlorine pesticides (Burgaz et aI., 1994; Durham, 1969; Zatz, 1972). Repeated exposure to a chemical may result in cumulative storage and an increased body burden of the chemical. If the interval between exposures is long relative to the biological half-life of the chemical, all or most of the chemical will be removed from the body prior to subsequent exposures and it is unlikely that the chemical will accumulate in the body. If the interval between exposures is short relative to the biological half-life, on the other hand, there will be a residual body burden from the first exposure when the second exposure occurs, and so on, such that the chemical accumulates in the body. Cumulative storage of a chemical upon repeated exposure continues until a steady state of storage is reached. Factors that influence storage include exposure level (dosage), time interval between exposures, duration of repeated exposures, interaction with other chemicals, age, sex, species, disease status, and nutritional status. See Section 24.5.2.3 for a mathematical discussion of storage.
24.5 PHARMACOKINETICS Pharmacokinetics is the modeling and mathematical description of the time course of chemical disposition (absorption, distribution, metabolism, and excretion). Although urine and exhaled breath may be obtained from humans, blood is the only tissue that can be readily and repeatedly sampled in humans. Pharmacokinetic models typically describe the change in blood (or plasma) concentration of the chemical with time. There are two basic approaches to characterizing the pharmacokinetics of a chemical in the body: compartmental and noncompartmental. Compartmental pharmacokinetic models represent the body as discrete compartments with mathematical descriptions of the movement of chemical between compartments, including the processes of absorption and elimination. Compartmental models may be subdivided into classical and physiologically based models. In contrast to compartmental models, the noncompartmental approach assumes no compartmentalization of the body and applies the trapezoidal rule for calculating the area under the plasma concentration-time curve to characterize a chemical's pharmacokinetics.
569
24.5.1 NONCOMPARTMENTAL MODELS
Noncompartmental models use statistical moment theory for analysis of plasma concentration-time data. Area under the plasma concentration vs. time curve (AUC) is a measure of the total systemic exposure to the chemical. AUC is the integral of the rate of change of concentration in plasma as a function of time: AUC =
10
00
(1)
C dt
The first moment of the plasma concentration vs. time curve is the plasma concentration multiplied by time vs. time curve, and the area under the first moment curve (AUMC) is AUMC =
10
00
tC dt
(2)
In pharmacokinetic studies, plasma samples are not collected through infinite time, but rather the collection period ends at some time T. AUC and AUMC may be approximated from zero time to T using the trapezoidal rule (Fig. 24.5) and then be extrapolated from time T to 00 as
fOO Cdt iT
f
T
oo
= CT
(3)
f3 tCdt
TCT
CT
f3
f32
= --+-
(4)
In Eqs. 3 and 4, C T is the observed concentration at the last time point T, and f3 is the slope of the terminal elimination phase of the log plasma concentration vs. time curve. The mean residence time (MRT), which represents the time that is required for 63.2% of the chemical to be eliminated, is AUMC MRT=-AUC
(5)
Bioavailability (F), which is the fraction of chemical that is absorbed after extravascular administration, may also be calculated using noncompartmental models. Because the process of absorption is bypassed with intravenous administration, bioavailability after intravenous administration is assumed to be unity. Bioavailability is determined for extravascular administration (e.g., oral, dermal) with reference to an intravenous dose as DivAUCex DexAUCiv
F=----
(6)
where Div and Dex are the intravenous and extravascular doses, and AUCiv and AUC ex are the areas under the plasma concentration vs. time curve after intravenous and extravascular doses. When the intravenous and extravascular doses are the same, F is simply the proportion of AUC after extravascular and intravenous doses. Bioavailability is often referred to as a percentage. For example, the bioavailability of orally administered permethrin is 0.61 or 61 % (Anadon et aI., 1991). A number of factors, including route of administration and species, may affect bioavailability. For example, the bioavailability of
570
CHAPTER 24
Absorption, Distribution, and Phannacokinetics
AI proxirnaled AUC = Areal + Are
~
0
• ...-1 ~
cd ~C2 ~
~
Ij)
U
~
0
U
C, Ex trapolated Area =Cr/P
Co = Cl' Cu to
tl
1.j= T
12
Tilne Figure 24.5 Determination of area under the plasma concentration vs. time curve using the trapezoidal rule (refer to Section 24.5.1).
paraquat was 45, 12, or 3.8% after an intratracheal, oral, or dermal dose, respectively (Chui et aI., 1988). Species-dependent bioavailability has been shown for orally administered metosulam, which was only 20% bioavailable in mice and dogs, but greater than 70% in rats (Timchalk et al., 1996). Plasma clearance (Cl), a measure of the inherent ability to remove a chemical from the body, is the volume of plasma that is cleared of the chemical per unit time. Cl after an intravenous dose is calculated as Cl
=
Div AUC
(7)
Cl can be calculated after an extravascular dose only if the bioavailability is 100% (i.e., F = 1). The apparent volume of distribution at steady state (Vss ) can be calculated after a single IV dose as Vss = Cl
X
MRTiv
(8)
The first-order elimination rate constant (k e ) and elimination half-life (tl/2) can also be calculated after a single IV dose for chemicals that appear to be characterized by a onecompartment model (see Section 24.5.3) from the relationships 1 MRTiv = -
(9)
tl/2 = 0.693MRTiv
(10)
ke
For chemicals that cannot be described by a one-compartment model, MRTiv = k'e
k~
Cl
(11) (12)
and the effective elimination half-life is the product of 0.693 andMRTiv. Repeated exposure to a chemical at constant time intervals may lead to accumulation of the chemical in the body until a steady state is achieved. During any exposure interval T at steady state, the rate of chemical entry into the body is equal to the rate of its elimination (i.e., amount absorbed equals amount eliminated). Wagner (1967) proposed the concept of a concentration index (Rc), which provides information with regard to the increased accumulation with multiple exposures. Rc is defined as the ratio of the average concentration of a chemical in blood during an exposure interval of length T at steady state (Coo) and the average concentration in blood during the same time interval after a single exposure (C'), Coo
=-=-
Rc
(13)
C
where Coo and C are defined as
- 1/12 Coo = -
T
Coodt
110' Cdt
C= -
T
(14)
11
(15)
0
In Eq. 15, Coo is the concentration of the chemical in blood or plasma at time t after dosing during steady state, and T = t2 - tl . When C and Coo are measured, Coo and C can be calculated from the respective concentration vs. time curves using the trapezoidal rule.
24.5 Pharmacokinetics
24.5.2 OVERVIEW OF CLASSICAL COMPARTMENTAL MODELS
24.5.3 ONE-COMPARTMENT MODEL A one-compartment model (Fig. 24.7), which represents the body as a single homogeneous compartment, adequately describes the pharmacokinetics of chemicals that rapidly equilibrate between blood and tissues. Therefore, it is reasonable to assume that the concentration of the chemical in blood (or plasma) is proportional to its concentration at the site of toxicity.
Classical compartmental models typically divide the body into one or more compartments that have no physiological or anatomical reality (Fig. 24.6). It is assumed that the rate of transfer between compartments, as well as the rate of elimination from the compartments, are linear or first-order processes. Each model has an associated series of mathematical equations that describe the absorption, elimination, and transfer of chemicals between compartments. These equations are dependent only on the model structure and are independent of the chemical under study. Classical compartmental models can provide important parameters that describe chemical disposition, including the volume of distribution, absorption and elimination rate constants, elimination half-life, and plasma clearance. In the following discussion, elimination is assumed to occur only from compartment I, which is referred to as the central compartment.
24.5.3.1 Intravenous Bolus Dose In the simplest one-compartment model, the chemical is introduced directly into the single compartment, and elimination occurs by a first-order process (Fig. 24.7). The single compartment has a volume Vd, which in this case is the apparent volume of distribution. A typical plasma concentration-time curve for a one-compartment system is shown in Fig. 24.7. This system is mathematically described by a first-order equation in which the
Compartment 1
Compartment 1
•
..
Compartment
Compartment 2
..
1
~
Compartment 2
1
,Ir
Compartment 3
..
~
Compartment 1
..
•
, Figure 24.6
571
Schematic representations of one-, two-, and three-compartment models.
Compartment 2
572
CHAPTER 24
Absorption, Distribution, and Pharmacokinetics iv Dose,
Ao Central Compartment, Volume=Vd
100
10
:::
8
~
6
:::
.S Cil ..... ....... ::: (!)
0 .p
b
::: (!) u :::
= ArfVd slope
:::
2
U
0
0
= -kJ2.303
CCt) 1I2C(t)
u
4
0
U
y-intercept = Co
10
~tl),
0.1 0.01
Time
Time
Figure 24.7 Representation of a one-compartment model with IV administration and first-order elimination, including a typical plasma concentration vs. time profile (linear and logarithmic scales). The volume of distribution, elimination rate constant, and elimination half-life are estimated by graphical methods.
rate of removal of the chemical (mass per time) is proportional to the body load of the chemical (mass), dA = -keA dt
At = Ao exp (-ket)
(17)
where At is the amount of chemical in the body at time t, and Ao is the amount of chemical in the body at time zero. More frequently, concentration rather than the amount of chemical is measured in plasma, and Eq. 17 is rewritten as
= Co exp (-ket)
ket
= log Co - 2.303
(19)
The graph of log Ct vs. t has a y intercept of Co and a slope of -ke/2.303; hence ke can be determined from the slope of the log Ct vs. t graph (Fig. 24.7). The apparent volume of distribution, Vd can be determined from the known amount of chemical introduced into the body by intravenous injection at time zero
(20)
The apparent volume of distribution is the volume into which the initial dose (Ao) would have to be dissolved to achieve the initial concentration of the chemical in plasma, Co. The elimination half-life (t1/2) of a chemical is the time required for the amount or concentration of chemical in plasma to decrease by one-half in the absence of additional exposure. Therefore, Ct is equal to one-half of Co after one half-life has passed since the dose was administered, and Co
2
= Co exp (-ke t l/2)
(21)
Equation 21 can be solved for tl/2,
(18)
where Ct and Co are the concentrations (units of mass/volume) of the chemical in plasma at time t and time zero, respectively. Taking the logarithm of both sides of Eq. 18 yields log Ct
Ao
Vd=Co
(16)
where A is the amount of chemical in the body (units of mass) and ke is the first-order elimination rate constant (units of reciprocal time), which represents the fractional elimination of chemical per unit time. A solution to Eq. 16 is
Ct
and the intercept of the log Ct vs. t graph as
0.693
tl/2= ~
(22)
which may also be estimated by inspection of the graph of log Ct vs. t (Fig. 24.7). Plasma clearance (Cl) is a measure of the inherent ability to remove a chemical from the body. Cl represents the volume of plasma that is cleared of the chemical per unit time and is the ratio of the rate of elimination (mass/time) and concentration (mass/volume ):
1_
_ldA/dt keAt _ keCt Vd _ CI- - - - - - - - - - k e V d Ct Ct Ct
(23)
24.5 Pharmacokinetics
removal of the chemical from the site of absorption is
Integration of Eq. 23 yields dose Cl=-AUC
(25) At = kaF D
Substitution of Eq. 24 into Eq. 25 and rearrangement leads to the equation for Vd: dose Vd=---AUC x ke
Humans are not typically exposed to pesticides by the intravenous route, but by extravascular routes (oral, dermal, inhalation), and the pesticide must be absorbed to enter the blood. Absorption is assumed to occur by a first-order process with an absorption rate constant ka as shown in Fig. 24.8. For extravascular exposure, then, the rate of removal of the chemical from the body is the net difference in the rates of introduction (by absorption) and elimination (by metabolism and excretion):
= kaAa
Ct
(28)
(29)
kaF D exp (-ket) - exp (-kat)
= - - ---------Vd
ka - ke
(30)
A typical plasma concentration-time curve for a compound that is absorbed by a first-order process rapidly equilibrates between blood and tissues, and is eliminated by a first-order process as shown in Fig. 24.8. After oral administration to rats, the plasma concentration vs. time profiles for tric10pyr (Timchalk et aI., 1996), diazinon (Wu et aI., 1996), and paraquat (Chui et aI., 1988) are all described by the model in Fig. 24.8. Some time after administration, absorption is essentially complete and the Eq. 30 is reduced to ka F D exp ( -ket) Vd
ka - ke
(31)
Taking the logarithm of both sides ofEq. 31 yields
ka
log Ct = log
kaF D ket - -Vd(ka - k e) 2.303
Central Compartment, Volume = Vd
100
8
::: 0
.
10
'p
~
6
::: u :::
4
::: ~ u :::
U
2
U
C':l
......
0
0
0
0.1 0.01
Time
(32)
The postabsorption phase of the graph of log Ct vs. t has a slope of -ke /2.303, and as in the case of a one-compartment model with an intravenous dose, ke can be determined from the
10
~
exp (-ket) - exp ( -kat) ka - ke
(27)
-keA
Extravascular Dose
b
-kaAa
Ct = - -
In Eq. 27, Aa is the mass of chemical at the site of absorption and A is the mass of chemical in the body. As was noted in Section 24.5.1, extravascular exposure to chemicals is different from intravenous exposure in that it cannot be assumed that 100% of the dose is absorbed. Some fraction F of the dose (D) is absorbed, or only the product F D is bioavailable. The rate of
::: .8
=
which can be rewritten in terms of concentration to yield
(26)
24.5.3.2 Extravascular Dose
dt
dt
Solving for A as a function of time in the preceding equations yields
Cl ke= Vd
-
dA
-
(24)
Equation 23 can also be rearranged to solve for
dA
573
Time
Figure 24.8 Representation of a one-compartment model with first-order absorption (i.e .• extravascular administration) and first-order elimination, including a typical plasma concentration vs. time profile (linear and logarithmic scales).
574
CHAPTER 24
Absorption, Distribution, and Pharmacokinetics
ka
10
,-.
Dose
~
bJJ
S
~
Volume = Vd
ke
y-intercept = kaFDoseN d(ka-~)
'-" ~
0
.~
~
\
~ ~
~
~
0.1
slope = -kJ2.303
, \ slope = -k/2.303
Cl)
U
~
\
0
U
0.01 0
6
12
18
Time (h) Figure 24.9 Estimation of volume of distribution, and absorption and elimination rate constants for the onecompartment model in Fig. 24.8 by graphical methods (i.e., curve stripping). Data are shown in Table 24.2 (refer to Section 24.5.3.2).
terminal slope of the log C t vs. t graph (Fig. 24.9). The absorption rate constant may be obtained from the y intercept of the plasma concentration vs. time graph, where the intercept is ka F D / (Vd (k a - k e )), or by the method of residuals as shown in the example in Fig. 24.9 and Table 24.2. Integration ofEq. 32 from zero time to infinity yields kaFD
AUC = Vd(k a - k e ) -
( 1 1) ke - ka
Wagner, 1967) derived the following equation for the average plasma concentration during any interval T at steady state: -
FD
FD
Coo = - - = VdkeT Ch
(37)
The average plasma concentration during the first dose interval (from time zero to T) is
(33)
which reduces to FD
(34)
(38)
Cl and Vd are derived from Eqs. 25 and 26, where dose is adjusted for bioavailability:
Substituting Eqs. 37 and 38 into Eq. 13 and rearranging, yields the concentration index
AUC=Vdke
Cl= Vd
FD
AUC FD
= ---AUC x ke
(35)
Rc =
1/(1- (~eXp(-keT) ka ke
(36) _ _k_e-eXP(-kaT))) ka - ke
24.5.3.3 Storage with Repeated Extravascular Exposure
(39)
For chemicals with ka » ke, which is the case for many of the organoch1orine pesticides, for instance (Zatz, 1972),
As discussed in Section 24.5.1, repeated administration of a given dose (D) of a chemical at fixed time intervals (T) eventually leads to a steady state (equilibrium) of storage (Fig. 24.10). The following discussion continues to assume that first-order processes govern absorption and elimination. Accumulation of a chemical in the body is described by the concentration index defined in Eq. 13. Wagner and colleagues (Wagner et aI.,
1965;
(40) and 1
Rc=-----1 - exp (-keT)
(41)
24.5 Phannacokinetics
The accumulation ratio (RA) for multiple exposures at fixed time intervals was defined by Wagner (1967) as the ratio of the average mass of chemical in the body during any exposure interval at equilibrium and the average mass of chemical absorbed after a single exposure. The average mass of chemical absorbed after a single exposure is simply F D, the percent of dose absorbed. Use of the relationship between concentration, mass, and volume of distribution shown in Eq. 20, allows Eq. 37 to be rearranged to solve for average mass of chemical during an exposure interval at steady state: FD Aoo = - ke r
Substitution of Eq. 22 into Eq. 42 yields 1.44F Dtl/2 Aoo = -------'-
Hence, 1.44FDtl/2 FDr
RA = ------'--
Extrapolated Residual
0.25
0.218
1.083
0.865
0.5
0.382
1.057
0.675
1
0.597
1.005
0.408
2
0.759
0.910
0.151
4
0.724
8
0.499
12
0.334
16
0.224
20
0.150
24
0.101
r
(44)
Many chemicals do not rapidly equilibrate between blood and tissues, and their plasma concentration-time profiles do not conform to the one-compartment model already described. Instead, elimination from plasma is multiphasic, the simplest case being biphasic elimination. The early phase is referred to as the distribution phase, and the later phase is the postdistribution or elimination phase. The plasma concentration of the chemical declines more rapidly during the distribution phase compared to the elimination phase. Two schematics of two-compartment models are shown in Fig. 24.6. The central compartment includes blood and tissues in which the chemical rapidly equilibrates (e.g., tissues that receive a high blood flow), and is considered a homogeneous compartment. This is analogous to the single compartment of a one-compartment model. The peripheral compartment consists of tissues for which equilibrium is not instantaneous. In classical compartmental models, the chemical moves between the central and peripheral compartments with associated transfer rate constants, but elimination is assumed to occur only from the central compartment with an associated elimination rate constant (Fig. 24.11). Absorption,
concentration (mglL) Plasma
1.44tl/2
24.5.4 MULTICOMPARTMENT MODELS
(42)
Plasma
(43)
r
Table 24.2 Data Used for the Method of Residuals Example Shown in Fig. 24.9 (onecompartment model with IV administration and first-order elimination)
Time (h)
575
95% of steady state concentration
--- / ............. .
...- _ _ Time to 95% steady state
Time Figure 24.10 Approach to steady state plasma concentration with repeated administration at constant dose intervals (r). See Section 24.5.3.3 for calculation of pharmacokinetic parameters.
576
CHAPTER 24 Absorption, Distribution, and Pharmacokinetics iv Dose
'"
~ k!2
Central Compartment
,.
~
Peripheral Compartment
k2!
klO
10 \:I 0
.~
~~
()
\:I 0
U
10
8
\:I 0
.~
6
.b \:I ~
4
()
\:I 0
U
2
0.1
0
Time
Time
Figure 24.11 Representation of a two-compartment model with IV administration and first-order elimination, including a typical plasma concentration vs. time profile (linear and logarithmic scales).
distribution, and elimination are assumed to be first-order processes.
Similar to the one-compartment model, the volume of the central compartment is C -
A schematic of a two-compartment model with first -order elimination is shown in Fig. 24.11. The diazinon plasma concentration vs. time after an IV dose is represented by this twocompartment model (Wu et aI., 1996). The concentration of chemical in plasma after an intravenous bolus dose as a function of time can be expressed as the sum of two monoexponential terms, (45) At some time after dosing, the distribution phase is complete and the only process that contributes to removal of the chemical from plasma is elimination. During this time, Eq. 45 reduces to Ct
= Be-j3t
(46)
(47)
The method of residuals is used to estimate A and a (Fig. 24.12 and Table 24.3). The initial concentration, Co, is determined by substituting t = 0 into Eq. 45: Co = A
+B
(48)
(49)
The rate constants a and f3 are composites of k12, k2l, and k10 with the relationships
a
+ f3
= k12
+ k2l + k10
(50)
(51)
af3 = klOk21
The rate constants k12, k2l, and k10 are determined from the relationships below (see Gibaldi and Perrier, 1982, for derivations). k2l =
Aa + Bf3 A+B af3
k10 = k2l k12
where B is the intercept and f3 is the slope of the terminal phase of the log Ct vs. t curve (Fig. 24.12). The rate constant f3 is analogous to ke in the one-compartmental model described in Section 24.5.3. The elimination half-life is estimated from f3 according to the equation 0.693 tl/2 = -f3-
dose A+B
V. - - -
24.5.4.1 Intravenous Bolus Dose
= a + f3 -
k2l - k10
(52) (53) (54)
24.5.4.2 Extravascular Dose A schematic of a two-compartment model with first-order absorption and elimination is shown in Fig. 24.l3. Absorption and elimination are assumed to occur via the central compartment only. The plasma concentration as a function of time can be expressed as Ct = A exp (-at)
+ B exp (-f3t) + C exp (-kOlt)
(55)
It is often difficult to distinguish the absorption phase from the distribution phase of the log Ct vs. t curve because kOl is similar
24.5 Pharmacokinetics
C(t) = Ae-at + Be-~t
10 ,,
~
0 ....... ~
~
Gl
\
~
C)
~
,,
-1---":"
~
0
y-intercept = A slope = -a/2.303
,,
1
~
U
~II
t"--- _ _
y-intercept = B slope = -~/2.303
0.1
'0,
,
U
e,
0.01
0
6
12
18
24
Time Figure24.12 Estimation of volume of A, B, et, and f3 for the two-compartment model in Fig. 24.13 by graphical methods (i.e., curve stripping). Data are shown in Table 24.2 (refer to Section 24.5.4.1).
Table 24.3 Data Used for the Method of Residuals Example Shown in Fig. 24.12 (twocompartment model with IV administration and first-order elimination) Extrapolated concentration (mg/L) Time (h)
Plasma
Plasma
Residual
1
6.534
0.488
6.046
2
4.318
0.468
3.850
4
1.988
0.431
1.558
8
0.614
0.364
0.250
12
0.345
0.308
0.037
16
0.264
20
0.220
24
0.187
in magnitude to a. Estimation of rate constants after an extravascular dose often requires data after an intravenous dose to distinguish between kOl and a (Gibaldi and Perrier, 1982). 24.5.5 PHYSIOLOGICALLY BASED MODELS
Unlike classical compartmental models, physiologically based pharmacokinetic (PBPK) models represent physiological and anatomical reality (Fig. 24.14). The compartments are connected by blood flow, and chemicals may enter the body by any route. The model in Fig. 24.14 incorporates exposure by the oral, dermal, and inhalation routes and elimination by urinary excretion, exhalation, and metabolism. PBPK models use mathematical descriptions of chemical disposition that are based on the physiological, physicochemical, and biochemical
determinants of disposition, which include biochemical reaction rates and tissue partition coefficients for the chemical, and the physiology (e.g., organ volumes, blood flows, respiration rates) of the animal. PBPK models do not assume that all processes governing disposition are linear, and saturable metabolism (Michaelis-Menten kinetics), for example, is easily incorporated into PBPK models. PBPK models exist for pesticides from a variety of chemical classes (Table 24.4). Like classical compartmental models, the compartments in Fig. 24.14 represent organs or tissue groups in which a chemical is uniformly distributed, and arrows represent the pathways that govern chemical disposition. Each compartment has an associated volume. Absorption is represented by arrows to the portals of entry for various routes of exposure, blood flow is represented by arrows that interconnect the compartments of the model, and metabolism and excretion are represented by arrows from the compartments in which these processes occur. PBPK models require three types of parameters as inputs: physiological, physicochemical, and biochemical. Physiological parameters (e.g., pulmonary ventilation rate, cardiac output, blood flow to tissues, and tissue volumes) for humans and several laboratory animal species are available in the literature (Arms and Travis, 1988). Physiological parameters, which are not dependent on the chemical under study, are assumed to be constant for a given species. However, if it is known that physiological parameters change with time or a particular exposure scenario, those changes can be easily incorporated into PBPK models. Physicochemical parameters used in PBPK models (i.e., tissue partition coefficients) describe the relative solubility of the chemical in various media (e.g., air, blood, and tissues). Biochemical parameters include the rates of absorption, metabolism, macromolecular binding, and excretion.
578
CHAPTER 24
Absorption, Distribution, and Pharmacokinetics
k12
kOl
Extravascular Dose
Central Compartment
~
~
"4
Peripheral Compartment
k21
,
kw
10
= 0
8
ro
6
= = 0
4
10
= 0
.~
!: ~ ()
U
.~
~
~ ()
= 0
U
2 0
0.1
Time
Time
Figure 24.13 Representation of a two-compartment model with first-order absorption (i.e., extravascular administration) and first-order elimination, including a typical plasma concentration vs. time profile (linear and logarithmic scales).
Inhaled Dose
E xhale d .....
Alveolar Space
...
Lunlklood
I",
_ ,~
~
Kidney
....
L6.
J
~
Poorly Perfused Tissues
Excreted
~
........
I~ ~
1
I.... I.....
Rapidly Perfused Tissues
I~
1
I
...
Fat
....
....
~----+
.. ./"
1
Skin
~
....
Liver
~
Dennal Dose
~
.....
.... ....
I~
Oral Dose
7 /
¥ Metabolized Figure 24.14 Schematic representation of a physiologically based pharmacokinetic model. This model contains descriptions of exposure by the inhalation, oral, and dermal routes, and elimination by exhalation, excretion, and metabolism (refer to Section 24.5.5).
To develop a PBPK model, one must first consider which compartments of the organism to include. The compartments may be specific organs, anatomical regions, or lumped tissue
groups, and their inclusion depends on which animal is being studied and whether or not the compartment contributes to the uptake, disposition, and/or toxicity of the chemical being
24.5 Pharmacokinetics
579
Table 24.4 Some Existing PBPK Models for Pesticides from Various Chemical Classes References
Pesticide Dieldrin
Leung and Paustenbach (1988)
Kepone (chlordecone)
el-Masri et al. (1995, 1996); Yang et al. (1995a, b)
Lindane (hexachlorocyclohexane)
DeJongh and Blaauboer (1997)
Hexachlorobenzene
Freeman et al. (1989); Roth et al. (1993)
Diisopropylfiuorophosphate
Gearhart et al. (1994)
Dichlorobenzene
Hissink et al. (1997)
2,4-Dichlorophenoxyacetic acid
Kim et al. (1994, 1995, 1996)
Captan
Fisher et al. (1992); Woollen (1993)
modeled. For example, the lung, gastrointestinal tract, and skin may be included because of their ability to serve as sites of absorption, and the kidneys may be included because of their ability to serve as portals of excretion. Tissue partition coefficients and the metabolic capacity of a particular tissue also contribute to chemical disposition. For example, fat is often included in PBPK models due to the high partition coefficient of lipophilic chemicals (e.g., organochlorine pesticides) in adipose tissue, and the liver is often included as a separate compartment because of its involvement in the metabolism of a wide variety of chemicals. Tissues that are target sites for toxicity are often included in PBPK models. The next step in PBPK model development is to write a mass balance equation for each compartment to describe the rate of change of chemical concentration in that compartment as a function of time. In the most general case, the mass balance equation for each tissue compartment is rate of change
= (rate of uptake) -
(rate of removal)
The rate of removal is the summation of removal by efflux back into the bloodstream, metabolism within the tissue, and excretion (e.g., biliary excretion in the liver or urinary excretion in the kidney). For simplicity, tissue compartments that have no capacity for metabolism or excretion will be considered. In this case, the rate of change of chemical concentration in the tissue is the difference in the rates of uptake and efflux. For uptake to occur from blood into a particular tissue compartment, the free chemical must diffuse out of the capillary space into the interstitial fluid, then diffuse across the plasma membrane to enter the intracellular space (Fig. 24.l5A). It is assumed that diffusion from the capillary membrane into the interstitial space is very rapid relative to diffusion from the interstitial space into the intracellular space, and the vascular and interstitial subcompartments are represented as one homogeneous subcompartment referred to as the extracellular space (Fig. 24.l5B). Free chemical in the blood enters the extracellular space of the tissue compartment at a rate (mass/time) that is the product of blood flow to the tissue (Q{, units of volume/time) and the concentration of the free chemical in the arterial blood (Ca). Diffusion of the chemical from the extracellular space across the plasma membrane and into the intracellular space is governed by Fick's
law of diffusion, which states that the rate of transfer of chemical across a membrane (flux) is proportional to its concentration gradient across the membrane flux =
PAt~C
(56)
where PAt is the tissue membrane permeation area crossproduct and ~C is the concentration gradient (units of mass/ volume) of free chemical across the membrane. For chemicals that have a perfusion-rate-limited distribution, diffusion of the chemical across the membrane is very rapid relative to its rate delivery to the tissue (i.e., PAt» Qt), and the rate of uptake by tissues is limited by blood flow rather than the rate of diffusion across the membrane. For such chemicals, the free chemical concentration in the intracellular and extracellular spaces is in equilibrium, and the tissue compartment can be represented as a single homogeneous compartment as shown in Figs. 24.14 and 24.15C. The mass balance equation for such a tissue compartment is dCt Vt = Qt(Ca - Cvt ) dt
(57)
where Vt is the volume of the tissue compartment, C t is the concentration of free chemical in the tissue compartment, Ca is the concentration of free chemical in the entering arterial blood, and Cvt is the concentration of free chemical in the venous blood exiting the tissue. The concentrations of free chemical in tissue and venous blood leaving the tissue are related by the tissue-to-blood partition coefficient (Pt) as
Ct
Cvt = Pt
(58)
The overall mixed venous blood concentration is Cv = LCvtQt Qc
(59)
where Qc is cardiac output (i.e., total blood flow or L Qt). For chemicals that have a distribution that is limited by the rate of diffusion across the cell membrane rather than blood flow (i.e., diffusion-rate-limited), separate mass balance equations must be written for the intracellular and extracellular compartments of the tissue (Fig. 24.l5B). The mass balance
580
CHAPTER 24
Absorption, Distribution, and Pharmacokinetics
QtCa
QtCvt
0
~
---------~---------
0
QtCa
QtCvt
Extracellular space
Intracellular space
0
c
Extracellular space
Intracellular space
~ ---------~---------
B
}
Interstitial space
0 ----------~---------0
A
Vascular space
o
Figure 24.15 Uptake from the vascular space of a tissue compartment into the intracellular space where the tissue is represented as three distinct compartments (A), two compartments (B), or a single homogeneous compartment (C). See discussion in Section 24.5.5.
equation for the extracellular space is
described by the equation dA
dmet t = VtKfCvt
(60) where Yes and Ces are the volume of and the concentration in the extracellular space, respectively. The mass balance equation for the intracellular space (i.e., tissue matrix) is des = PAt ( Cvt - -Ct) Vis-dt Pt
(61)
where V;s and Cis are the volume of and the concentration in the intracellular space, respectively. Chemicals may be effectively eliminated from a tissue by metabolism, macromolecular binding, and/or excretion. For these tissues, the mass balance equation is more complex than those shown in Eqs. 57, 60, and 6l. The mass balance equation for a tissue that metabolizes the chemical (e.g., the liver) is
where K f is the first-order rate constant. Both saturable and first-order metabolism may occur simultaneously. For example, if metabolism by both saturable and first-order processes is occurring in the liver, the mass balance equation for the liver would be dCz Vzdt
= QZ(Ca -
Cvl) -
dC t
tTt =
Qt(Ca
-
Cvt )
dA
-
dmet t
(62)
where dAmetldt is the rate of metabolism. Many enzyme systems are saturable, and dA met dt
(63)
where Vmax and Km are the maximal velocity and Michaelis constant of the enzymatic reaction. First-order metabolism is
VmaxCvl
Km
+ Cvl
- KfCvIVZ
(65)
For inhalation exposure to volatile chemicals, equilibrium is established between the chemical in the alveolar air space of the lung and the chemical in arterial blood. The concentration of chemical in arterial blood (Ca) is described by the equation Ca
QcCv + QpCi = -----'---Qc
V
(64)
+ (Qp/ Ph)
(66)
where Qp is alveolar ventilation rate, Ci is the concentration in inhaled air, and Ph is the blood: air partition coefficient. If the chemical can be inhaled, it can also be exhaled. The concentration of the chemical in exhaled alveolar air is the ratio of its concentration in arterial blood and the blood: air partition coefficient, Ca / Ph. Upon oral ingestion of a chemical (e.g., food, drinking water), the chemical may be absorbed into the portal circulation or into the lymphatic system. For simplicity, we assume
References
first-order absorption into the portal circulation only. In this situation, the chemical is delivered directly to the liver prior to distribution throughout the body. This requires an input term in the mass balance equation for the liver and a mass balance equation that describes the rate of loss of chemical from the site of absorption (stomach):
dA st dt
(67)
dCI VI dt
VrnaxC vt Km + Cvt
-
K C v: f vi I
where Ka is the oral absorption rate constant and Ast is the amount (mass) of chemical in the stomach (i.e., the site of absorption). Incorporation of dermal absorption is more complex in that a skin compartment must be added to the model (see Krishnan and Andersen, 1994). In addition to metabolism, chemicals may be eliminated from the body by excretion in urine, exhaled air, and other routes that are not be discussed here (e.g., sweat, bile, and milk). A typical equation for the concentration of chemical in exhaled air is Ca
Cex = 0.7 Ph
+ 0.3Ci
Durham, W. F. (1969). Body burden of pesticides in man. Ann. NY Acad. Sci. 160, 183-195. el-Masri, H. A., Thomas, R. S., Benjamin, S. A., and Yang, R. S. H. (1995). Physiologically based pharmacokinetic/pharmacodynamic modeling of chemical mixtures and possible applications in risk assessment. Toxicology 105, 275-282. el-Masri, H. A., Thomas, R. S., Sabados, G. R., Phillips, J. K., Constan, A. A., Benjamin, S. A., Andersen, M. E., Mehendale, H. M., and Yang, R. S. H. (1996). Physiologically based pharmacokinetic/pharmacodynamic modeling of the toxicologic interaction between carbon tetrachloride and Kepone. Arch. Toxicol. 70,704-713. Fisher, H. L., Hall, L. L., Sumler, M. R., and Shah, P. V. (1992). Dermal penetration of 4 C]captan in young and adult rats. J. Toxicol. Environ. Health 36,251-271. Freeman, R. A., Rozman, K. K., and Wilson, A. G. E. (1989). Physiological pharmacokinetic model of hexachlorobenzene in the rat. Health Phys. 57 (Suppl. 1), 139-147. Garcia-Repetto, R., Martinez, D., and Repetto, M. (1995). Coefficient of distribution of some organophosphorous pesticides in rat tissue. Vet. Human Toxico!. 37,226-229. Garrettson, L. K., and Curley, A. (1969). Dieldrin: Studies in a poisoned child. Arch. Environ. Health 19, 814-822. Gearhart, J. M., Jepson, G. W., Clewell, H. J., Andersen, M. E., and Conolly, R. B. (1994). Physiologically based pharmacokinetic model for the inhibition of acetylcholinesterase by organophosphate esters. Environ. Health Perspect. 102 (Suppl. 11),51-60. GibaIdi, M., and Perrier, D. (1982). "Pharmacokinetics," 2nd ed. Dekker, New York. Gomez-Catalan, J., To-Figueras, J., Rodamilans, M., and Corbella, J. (1991). Transport of organochlorine residues in the rat and human blood. Arch. Environ. Contam. Toxico!. 20,61-66. Hayes, W. J., Jr., Quinby, G. E., Walker, K. C., Elliott, J. w., and UphoJt, W. M. (1958). Storage of DDT and DDE in people with different degrees of exposure to DDT. Arch. Ind. Health 18, 398-406. Hissink, A. M., Van Ommen, B., Kruse, J., and Van Bladeren, P. J. (1997). A physiological based pharmacokinetic (PB-PK) model for 1,2dichlorobenzene linked to two possible parameters of toxicology. Toxico!. App!. Pharmaco!. 145,301-310. Hodgson, E., Silver, I. S., Butler, L. E., Lawton, M. P., and Levi, P. E. (1991). Metabolism. In "Handbook of Pesticide Toxicology" (W. J. Hayes, Jr. and E. R. Laws, Jr., eds.), pp. 107-167. Academic Press, San Diego. Kennedy, G. L., Jr., and Valentine, R. (1994). Inhalation toxicology. In "Principles and Methods of Toxicology," 3rd ed. (A. W. Hayes, ed.), pp. 805-838. Raven Press, New York. Kim, C. S., Binienda, Z., and Sandberg, J. A. (1996). Construction of a physiologically based pharmacokinetic model for 2,4-dichlorophenoxyacetic acid dosimetry in the developing rabbit brain. Toxico!. App!. Pharmacol. 136, 250-259. Kim, C. S., Gargas, M. L., and Andersen, M. E. (1994). Pharmacokinetic modeling of 2,4-dichlorophenoxyacetic acid (2,4-D) in rat and in rabbit brain following single dose administration. Toxico!. Left. 74, 189-201. Kim, C. S., Slikker, W., Jr., Binienda, Z., Gargas, M. L., and Andersen, M. E. (1995). Development of a physiologically based pharmacokinetic model for 2,4-dichlorophenoxyacetic acid dosimetry in discrete areas of the rabbit brain. Neurotoxico!. Terato!' 17, 111-120. Krishnan, K., and Andersen, M. E. (1994). Physiologically based pharmacokinetic modeling in toxicology. In "Principles and Methods of Toxicology," 3rd ed. (A. W. Hayes, ed.), pp. 149-188. Raven Press, New York. Lehman, A. J. (1956). The minute residue problem. Q. Bull. Assoc. Food Drug Off. 20, 95-99. Leung, H. W., and Paustenbach, D. J. (1988). Application of pharmacokinetics to derive biological exposure indexes from threshold limit values. Am. Ind. Hyg. Assoc. J. 49,445-450. Ludwig, G., Weis, J., and Korte, F. (1964). Excretion and distribution of aldrin14C and its metabolites after oral administration for a long period of time. Life Sci. 3, 123-130.
e
(68)
(69)
Equation 69 uses the assumption that exhaled air is a mixture of inhaled air (30%) and expired alveolar air (70%). Urinary excretion may be described in a number of ways, including excretion by first-order and saturable processes similar to the descriptions of metabolism in Eqs. 63 and 64. The same is true for elimination by other routes.
REFERENCES Ambrose, A. M., Christensen, H. E., Robbins, D. J., and Rather, L. J. (1953). Toxicological and pharmacological studies on chlordane. Arch. Ind. Hyg. Occup. Med. 7, 197-210. Anadon, A., Martinez-Larranaga, M. R., Diaz, M. J., and Bringas, P. (1991). Toxicokinetics of permethrin in the rat. Toxico!. App!. Pharmaco!' 110, 1-8. Arms, A. D., and Travis, C. C. (1988). "Reference Physiological Parameters in Pharmacokinetic Modeling." Rep. NTIS PB 88-196019, Office of Health and Environmental Assessment, U.S. Environmental Protection Agency, Washington, DC. Burgaz, S., Afkham, B. L., and Karakaya, A. E. (1994). Organochlorine pesticide contaminants in human adipose tissue collected in Ankara (Turkey) 1991-1992. Bull. Environ. Contam. Toxicol. 53,501-508. Casey, P. H., and Collie, W. R. (1984). Severe mental retardation and multiple congenital anomalies of uncertain cause after extreme parental exposure to 2,4-D. J. Pediatr. 104,313-315. Chui, Y. c., Poon, G., and Law, F. (1988). Toxicokinetics and bioavailability of paraquat in rats following different routes of administration. Toxico!. Ind. Health 4, 203-219. Dale, W. E., Gaines, T. B., and Hayes, W. J., Jr. (1962). Storage and excretion of DDT in starved rats. Toxico!. App!. Pharmaco!. 4, 89-106. Davidow, B., and Frawley, J. P. (1951). Tissue distribution, accumulation and elimination of the isomers of benzene hexachloride. Proc. Soc. Exp. Bio!. Med. 76,780-783. DeJongh, J., and Blaauboer, B. J. (1997). Simulation of lindane kinetics in rats. Toxicology 122, 1-9.
581
582
CHAPTER 24
Absorption, Distribution, and Pharmacokinetics
Opdycke, J. C., and Menzer, R. E. (1984). Phannacokinetics of diflubenzuron in two types of chickens. J. Toxieol. Environ. Health 13,721-733. Pekas, J. C. (1972). Intestinal hydrolysis, metabolism and transport of a pesticidal carbamate in pH 6.5 medium. Toxieol. Appl. Pharmaeol. 23, 62-70. Robinson, J., Roberts, M., Baldwin, M., and Walker, A. I. T. (1969). The pharmacokinetics of HEOD (dieldrin) in the rat. Food Cosmet. Toxieol. 7, 317-332. Roth, W. L., Freeman, R. A., and Wilson, A. G. E. (1993). A physiologically based model for gastrointestinal absorption and excretion of chemicals carried by lipids. Risk Anal. 13,531-543. Sandberg, J. A., Duhart, H. M., Lipe, G., Binienda, Z., Slikker, W., Jr., and Kim, C. S. (1996). Distribution of 2,4-dichlorophenoxyacetic acid (2,4-D) in maternal and fetal rabbits. J. Toxieol. Environ. Health 49, 497-509. Street, J. C. (1964). DDT antagonism to dieldrin storage in adipose tissue of rats. Science 146, 1580-1581. Thies, M. L., and McBee, K. (1994). Cross-placental transfer of organochlorine pesticides in Mexican free-tailed bats from Oklahoma and New Mexico. Arch. Environ. Contam. Toxieol. 27, 239-242. Timchalk, C., Dryzga, M. D., Johnson, K. A., Eddy, S. L., Freshour, N. L., Kropscott, B. E., and Nolan, R. J. (1996). Comparative pharmacokinetics of 4 C]metosulam (N -[2,6-dichloro-3-methylphenyI1-5,7-dimethoxy1,2,4-triazolo[I,5al-pyrimidine-2-sulfonamide) in rats, mice and dogs. J. Appl. Toxieol. 17,9-21.
e
Turner, J. c., and Shanks, V. (1980). Absorption of some organochlorine compounds by the rat small intestine-in vivo. Bull. Environ. Contam. Toxieol. 24,652--655. Wagner, J. G. (1967). Drug accumulation. J. Clin. Pharmaeol. 7, 84-88. Wagner, J. G., Northam, J. I., Alway, C. D., and Carpenter, O. S. (1965). Blood levels of drug at the equilibrium state after multiple dosing. Nature 207, 1301-1302. Woollen, B. H. (1993). Biological monitoring for pesticide absorption. Ann. Occup. Hyg. 37, 525-540. Wu, H. X., Evreux-Gros, C., and Descotes, J. (1996). Diazinon toxicokinetics, tissue distribution and anticholinesterase activity in the rat. Biomed. Environ. Sci. 9,359-369. Yang, R. S., el-Masri, H. A., Thomas, R. S., and Constan, A. A. (1995a). The use of physiologically-based pharmacokinetic/pharmacodynamic dosimetry models for chemical mixtures. Toxieol. Lett. 82-83, 497-504. Yang, R. S., el-Masri, H. A., Thomas, R. S., Constan, A. A., and Tessari, J. D. (1995b). The application of physiologically based pharmacokinetic!pharmacodynamic (PBPKlPD) modeling for exploring risk assessment approaches of chemical mixtures. Toxieol. Lett. 79, 193-200. Zatz, J. L. (1972). Accumulation of organochlorine pesticides in man. J. Pharm. Sci. 61,948-949.
CHAPTER
25 Pesticide Excretion Emest Hodgson North Carolina State University
25.1 INTRODUCTION Except in simple life forms, elimination of toxicants, including pesticides and their metabolites, is part of a specialized system that, in addition to elimination, maintains the balance of water, minerals, and other substances necessary for terrestrial life. Pesticides, again typical of toxic ants in general, are taken up by the body in most cases because of their lipophilicity. Before elimination is possible, they must first be metabolized into a form simulating that used by the body for the elimination of endogenous compounds. In general, they are metabolized by Phase I and Phase II enzymes to conjugation products that are more polar and hence more hydrophilic than the parent compound and then excreted by either the renal or the hepatic route. Although similar anion and cation transport systems are found in both kidney and liver, they differ in the type of excretory products eliminated. The renal system eliminates molecules of molecular mass smaller than 400-500, whereas the liver handles larger molecules. The molecular mass threshold between renal and biliary excretion varies with species (Hirom et aI., 1972), although in most species there are excretory products that are excreted by both systems. In addition, highly lipophilic chemicals that are recalcitrant to metabolism may be excreted as the parent chemical by a number of alternate routes, although these are generally of minor importance compared to urine and bile. General aspects of excretion of toxicants and their metabolites may be found in Levi et al. (1997), Matthews (1994), Pritchard and James (1982), and Tarloff (2000). Excretion of pesticides and their metabolites has not been extensively investigated perhaps because the rate of excretion seldom appears to be a rate-limiting step in the ultimate expression of toxicity. However, urinary metabolites have been utilized as biomarkers of exposure.
metabolism. They are also the primary organs for excretion of polar xenobiotics and polar metabolites of lipophilic xenobiotics. A useful description of kidney structure and function has recently been published by Tarloff (2000). The functional unit of the kidney, the nephron, is shown in Fig. 25.1.
25.2.2 GLOMERULAR FILTRATION Passive filtration of the blood plasma in the glomerulus, under the influence of the blood pressure generated by the heart, is the initial step in urine formation. All molecules small enough to pass through the glomerular pores (70-100 A) appear in the ultrafiltrate; any molecule larger than these pores or bound to molecules larger than these pores will not appear in the ultrafiltrate.
25.2.3 TUBULAR REABSORPTION Tubular reabsorption is the second major step in urine formation. Most of the reabsorption of solutes necessary for normal body function such as amino acids, glucose, and salts takes place in the proximal part of the tubule. This reabsorption may be active, as in the case of glucose, amino acids, and peptides, whereas water, chloride, and other ions are passively reabsorbed. Reabsorption of water and ions also occurs in the distal tubule and in the collecting duct. Reabsorption of xenobiotics is usually passive and controlled by the same principles that regulate their passage across any membrane. That is, lipophilic compounds cross cell membranes more rapidly than polar ones; hence, lipophilic toxicants will tend to be passively re absorbed more than polar ones and, overall, elimination of polar toxic ants and their polar metabolites will be facilitated.
25.2 RENAL FUNCTION 25.2.4 TUBULAR SECRETION 25.2.1 OVERALL ASPECTS The kidneys are primarily organs of excretion, and elimination by the kidney accounts for most by-products of normal body Handbook of Pesticide Toxicology Volume 1. Principles
Tubular secretion is another important mechanism for excretion of solutes by the kidney. Secretion across the wall of the tubule is generally active, with two systems, one for the secre-
583
Copyright © 2001 by Academic Press. All rights of reproduction in any fonn reserved,
584
CHAPTER 25
Pesticide Excretion PROXIMAL TUBULE
25.3 BILIARY EXCRETION
GLOMERULUS
ARTERY
(~~~I~r-4~- OISTAL TUBULE
~
COLLECTING DUCT
THIN SEGMENT Figure 25.1
Nephron and vasculature of mammalian kidney.
tion of organic acids, including conjugates, the other for the secretion of organic bases. Passive secretion may occur as a result of a process known as diffusion trapping. Un-ionized weak acids and bases pass across the membrane into the lumen of the tubule and, depending on the pH of the urine, one or the other may become ionized and unable to diffuse back across the lumen wall. Diffusion trapping is, of course, extremely sensitive to variations in urine pH, a factor that may be utilized to speed elimination of toxicans. For example, alkalinization of the urine by ingestion of bicarbonate speeds up the elimination of salicylate. Tubular secretion, and hence excretion, of organic anions has been known to be of importance in the excretion of certain pesticides for some time (Pritchard and James, 1982) 2,4-Dichlorophenoxyacetic acid (2,4-D) and 2,4,5-trichlorophenoxy acetic acid (2,4,5-T) are usually applied as salts or esters, the latter being readily hydrolyzed in the body, and studies of their excretion have emphasized the patient acids although various conjugates are also transported by the organic anion transport system (Erne, 1966; Pritchard and James, 1982). Active tubular secretion of 2,4-D has been demonstrated in a number of species, including rabbit (Dybing and Kolberg, 1967), rat (Fang et al., 1973), chicken (Erne and Sperber, 1974), dog (Hook et al., 1976), goat (Orberg, 1980), and flounder (Pritchard and James, 1979). (DDT) and its principal metabolite, (DDE) are highly lipophilic and the latter is recalcitrant further metabolism. Thus, DDT and, to a greater extent, DDE are sequestered in body lipids and have an extremely long half-life in the body. Some portion of DDT, however, is metabolized a organic acid, DDA, by dechlorination and oxidatioh at the I-position (Pinto et al., 1965), (DDA) is a substrate for the organic acid transport system (Pritchard, 1976, 1978) and, as a consequence, is excreted considerably more rapidly than DDT or DDE.
Excretion by the liver, through the biliary system, has been known for a considerable time but, due to the difficulty in obtaining uncontaminated bile, has been less intensively investigated than renal excretion. A brief review of hepatic excretion may be found in Levi et al. (1997). Bile is secreted by the liver cells into the bile canaliculi. It then flows into the terminal branches of the bile duct, the hepatic duct, and the gall bladder. The contents of the gall bladder are discharged into the gut under the influence of hormones whose release is triggered by food ingestion. In species that lack a gall bladder, such as the rat, bile flows continuously into the duodenum. Secretion of xenobiotics or their metabolites into the bile is largely a function of molecular mass and may be by passive diffusion or by active transport. Enterohepatic circulation is an important aspect of biliary excretion. Nonpolar xenobiotics are normally oxidized and then conjugated. If the molecular mass of the conjugate is appropriate for biliary excretion, it enters the gut where hydrolysis by intestinal microflora or gut conditions may occur. The compound, then being again in a less polar form, can be reabsorbed by the intestine and returned to the liver through portal circulation and the process repeated. Enterohepatic circulation thus increases the biologic half-life and possibly adverse effects of toxicants, particularly to the liver. For therapeutic purposes, the cycle can be interrupted by feeding an agent that binds the hydrolysis product and prevents its reabsorption, as in the use of cholestyramine in chlordecone poisoning.
25.4 RESPIRATORY EXCRETION Volatile toxic ants such as ethanol or pesticidal fumigants may be eliminated via the lungs as may volatile metabolites, including acetone, carbon dioxide, etc. Respiratory excretion is not known to be an important route for excretion of pesticides, in general, or their metabolites.
25.5 OTHER ROUTES OF EXCRETION There are a number of other, less important routes of excretion, including sex-linked routes and alimentary elimination, and several routes based on natural secretory or growth processes are known.
25.5.1 SEX-LINKED ROUTES OF EXCRETION Certain routes of xenobiotic elimination are restricted to females, including excretion through milk, eggs, and fetus. Although such excretion is probably of minimal benefit to the mother, it may have serious consequences to the offspring.
25.6 Cellular Elimination
Milk Because milk is an emulsion of lipids in an aqueous protein solution, it may contain xenobiotics of many different physicochemical properties ranging from polar compounds such as alcohol and caffeine to less polar drugs to highly lipophilic chemicals such as DDT and DDE. Elimination of toxicants in milk is highly dependent on the biological halflife of the toxicant. Milk normally plays a minor role in the excretion of chemicals with short half-lives but may be important for some chemicals with long half-lives. In experimental studies with chlorinated insecticides, up to 25% of the dose administered to cows was eliminated in the milk. In some South American countries, the DDT content of human mother's milk is close to the acceptable daily intake recommended by the World Health Organization (WHO). Although adverse effects on infants were not seen in these cases, when nursing mothers were accidentally exposed to hexachlorophene (Turkey) or polychlorinal biphenyls (PCB) (Japan), signs of intoxication were seen in a number of infants. Eggs Polar toxicants and metabolites may be eliminated in egg white and lipophilic compounds in the yolk. The effects of this on developing birds is controversial but may be significant, particularly if bioconcentration has occurred in the food chain. Effects of toxicants excreted into avian eggs should not be mistaken for the well-documented eggshell thinning, which is an effect on the female reproductive system. Fetus The elimination of maternally derived toxicants in the fetus is of little or no benefit to the mother and, due to the generally small amounts involved, is usually of little or no harm to the fetus. However, as shown by the toxic effects of mercury, thalidomide, and diethylstilbestrol, this is not always the case.
25.5.2 ALIMENTARY ELIMINATION Passive elimination of lipophilic toxicants directly through the wall of the alimentary canal is probably, in most cases, unimportant, at least from a quantitative viewpoint. However, although slow, it may be an important route for excretion of chlordecone, particularly if reabsorption is prevented by administration of cholestyramine.
25.5.3 OBSCURE ROUTES OF EXCRETION Because passive diffusion of lipophilic toxicants may occur across any cell membrane, it might be expected that such chemicals will appear in many body secretions, such as sweat, or growth products, such as hair, nails, and skin. The sebaceous glands secrete an oily secretion and, probably for this reason, insecticides and PCBs have been found in human hair. Arsenic, mercury, and selenium have also been associated with hair. Although such routes of excretion are probably only a small proportion of the total excretion of any particular xenobiotic, they may provide a noninvasive method of estimating exposure or total body burden. Analysis of bird feathers is useful for the
585
assessment of heavy-metal exposure and the amount of cotinine, a major metabolite of nicotine, in saliva has been used extensively as a biomarker for nicotine uptake. The excretion of atrazine in saliva has also been tested in rats as a potential biomarker of exposure in exposed workers and shows promise for this application (Lu et aI., 1997).
25.6 CELLULAR ELIMINATION To prevent concentration at toxic levels, hepatocytes and other cells have active transport processes to eliminate xenobiotics. Because the metabolism of xenobiotics generally yields products that are more polar and, consequently, have reduced capacity for passive diffusion than the parent compound, such transport processes are essential for cell viability. Two active transport proteins known to contribute to xenobiotic elimination are the multi drug resistance-associated protein (MRP) and p-glycoprotein, both adenosine 5' -triphosphate dependent active transport agents. Both were identified initially because of their overexpression in drug-resistant cells. MRP has the capacity to transport glutathione, sulfate, or glucuronide conjugates, whereas p-glycoprotein is known to transport a wide array of xenobiotics. The importance of their role is illustrated by the fact that p-glycoprotein knockout mice died when treated with the miticide ivermectin, subsequently shown to be due to the accumulation of ivermectin in the brain due to the absence of p-glycoprotein in the blood-brain barrier (Schinkel et aI., 1994). On the basis of their ability to inhibit the p-glycoproteinmediated efflux of doxorubicin, several pesticides of different chemical classes were shown to bind to human p-glycoprotein (Bain and LeBlanc, 1996). The most effective were the organochlorines, chlordecone, endosulfan, heptachlor, and heptachlor epoxide; the organophosphorus insecticides, chlorpyrifos, chlorthiophos, dicapthon, leptophos, parathion, and phenamiphos, as well as clotrimazole and ivermectin. None of the carbamates or pyrethroids tested was effective. Lipophilicity and molecular weight were major determinations of pesticide binding with log Kow values of 3.6-4.5 and molecular weights of 391-490 being optimal. The authors point out that the ability to inhibit p-glycoprotein function does not necessarily mean that the chemical will be transported. Only endosulfan, the compound with the best binding characteristics, could be shown to be transported by p-glycogen.
25.7 EXCRETION OF PESTICIDES AND THEIR METABOLITES AS BIOMARKERS OF EXPOSURE There have been a number of studies using urinary pesticides or their metabolites as biomarkers of exposures. The early studies in this area were summarized in a 1989 American Chemical Society monograph (Wang et aI., 1989). Some of these studies involve single compounds, primarily but
586
CHAPTER 25
Pesticide Excretion
Table 25.1 Some Examples of the Use of Urinary Metabolites of Pesticides as Biomarkers of Pesticide Exposure Pesticide
Urinary metabolite
Reference
Chlorpyrifos
Diethylphosphate
Griffin et aI., 1999
Diethylthiophosphate Chlorpyrifos/quinalphos
Diethylphosphate
Vasilic et aI., 1992
Diethylthiophosphate Chlorpyrifos-methyl
3,5,6-trichloropyridinol
Aprea et aI., 1997
Alkyl phosphates 2,4-Dichlorophenoxyacetic
2,4-D
Harris et aI., 1992
N -Acetyl-S-(cis-3-chloroprop-2-
Osterloh and Feldman, 1993
acid (2,4-D) 1,3-Dichloropropene
enyl) cysteine Dicofol
4, 4'-Dichlorobenzilic acid
Nigg et aI., 1991
Guthion
Dimethylphosphorothioic acid
Franklin et aI., 1981
Malathionlthiometon
Dilimethyl phosphate
Vasilic et aI., 1999
Dimethyl phosphorothioate Dimethy Iphosphorodithioate Organophosphorus insecticides
Alkyl phosphates
not exclusively organophosphorus compounds, and examples are given in Table 25.1. Other studies are surveys of populations either exposed or potentially exposed to multiple pesticides, For example, urine from a sample of 1000 residents of the United States was analyzed for 12 analytes potentially derived from pesticides and 6 were frequently found (Hill et al., 1995). These, with possible parent compounds, were 2,5-dichlorophenol (from lA-dichlorobenzene); 2A-dichlorophenol (from bifenox, clomethoxyfen, dichlofenthion, etc.); I-naphthol (from naphthalene, carbaryl, etc.); 2-naphthol (from naphthalene, etc.); 3,5,6-trichloro-2-pyridinol (from chlorpyrifos, chlorpyrifos-methyl); and pentachlorophenol (from pentachlorophenol, pentachloronitrobenzene). In another large study of multiple exposure, in tree nursery workers, only a small number, 42 out of 3134, of urine samples were positive, in this case for benomyl, bifenox, and carbaryl (Lavy et aI., 1993). A summary of all methods, including measurement of urinary metabolites, of estimating exposure by biomarkers has recently appeared (Maroni et aI., 2000). Recently, urinary mercapturic acids have been extensively explored for use as biomarkers of exposure and several detailed reviews are available (De Rooij et aI., 1998; Van We lie et aI., 1992). Although the emphasis has been on industrial and environmental chemicals, this potentially valuable technique has not been applied extensively to pesticides. However, the soil nematocide dichloropropene was included. The effect of pesticides on the excretion of metabolites of endogenous metabolism has been explored to some extent. For example, in rats treatment with dimethoate decreased the excretion of proline and lysine derivatives known to be collagen metabolites (Reddy et aI., 1991). The mechanism of this effect was not investigated and the magnitude of the effect did not seem to be large enough for practical application.
Azaroff, 1999
N -acetylglucosamidase was found to be slightly increased in the urine of applicators exposed to the soil nematocide, 1,3dichloropropene, along with its principal metabolite, N -acety 1S-(cis-3-chloroprop-2-enyl) cysteine (Osterloh and Feldman, 1993).
REFERENCES Aprea, c., Sciarra, G., Sartorelli, P., Sartorelli, E., Strambi, E, Farina, G. A., and Fattorini, A. (1997). Biological monitoring of exposure to chlorpyrifosmethyl by assay of urinary alkylphosphates and 3,5,6-trichloropyridinol. J. Toxicol. Environ. Health 50, 581-594. Azaroff, L. S. (1999). Biomarkers of exposure to organophosphorus insecticides among farmers families in rural El Salvador: Factors associated with exposure. Environ. Res. 80, 138-147. Bain, L. J., and LeBlanc, G. A. (1996). Interaction of structurally diverse pesticides with the human MDRI gene product p-glycoprotein. Toxicol. Appl. Pharmacol. 141, 288-298. De Rooij, B. M., Commandeur, J. N. M., and Vermeulen, N. P. E. (1998). Mercapturic acids as biomarkers of exposure to electrophilic chemicals: Applications to environmental and industrial chemicals. Biomarkers 3, 239-303. Dybing, E, and Kolberg, A. (1967). Inhibition of the renal tubular transport of p-aminohippurate (Tm-PAH) in the rabbit caused by subtoxic doses of dichlorophenoxyacetate (2,4-D). Acta Pharmacol. Toxicol. 25, 51-61. Erne, K (1966). Distribution and elimination of phenoxyacetic acids in animals. Acta Vet. Scand. 7, 240-256. Erne, K, and Sperber, 1. (1974). Renal tubular transfer of phenoxyacetic acids in the chicken. Acta Pharmacol. Toxieol. 35,233-241. Fang, S. c., Fallin, E., Montgomery, M. L., and Freed, V. H. (1973). The metabolism and distribution of 2,4,5-trichlorophenoxyacetic acid in female rats. Toxieo!. Appl. Pharmaeol. 24, 555-563. Franklin, C. A., Fenske, R. A., Greenhalgh, R., Mathieu, L., Denley, H. v., Leffingwell, J. T., and Spear, R. C. (1981). Correlation of urinary pesticide metabolite excretion with estimated dermal contact in the course of occupational exposure to guthion. 1. Toxieol. Environ. Health 7, 715-731. Griffin, P., Mason, H., Heywood, K, and Cocker, J. (1999). Oral and dermal absorption of chlorpyrifos: A human volunteer study. Oeeup. Environ. Med. 56,10-13.
References
Harris, S. A., Solomon, K. R., and Stephenson, G. R. (1992). Exposure of homeowners and bystanders to 2,4-dichlorophenoxyacetic acid (2,4-D). 1. Environ. Sci. Health, B 27, 23-38. Hill, R. H., Jr., Head, S. L., Baker, S., Gregg, M., Shealy, D. B., Bailey, S. L., WiIIiams, C. c., Sampson, E. J., and Needham, L. L. (1995). Pesticide residues in urine of adults living in the United States: Reference range concentrations. Environ. Res. 71, 99-108. Hirom, P. C., Milburn, P., Smith, R. L., and WiIIiams, R. T. (1972). Species variations in the threshold molecular-weight factor for the biliary excretion of organic acids. Biochem. 1.129, 1071-1077. Hook, J. B., Cardona, R., Osborn, J. L., Bailie, M. D., and Gehring, P. J. (1976). The renal handling of 2,4,5-trichlorophenoxyacetic acid (2,4,5-T) in the dog. Food Cosmet. Toxicol. 14, 19-23. Lavy, T, L., Mattice, J. D., Massey, J. H., and Skulman, B. W. (1993). Measurements of year-long exposure to tree nursery workers using multiple pesticides. Arch. Environ. Contam. Toxicol. 24, 123-144. Levi, P. E., Hodgson, E., and LeBlanc, G. A. (1997). Elimination of toxicants. In "A Textbook of Modern Toxicology" (E. Hodgson and P. E. Levi, eds.), 2nd ed. Appleton & Lange, East Norwalk, CT. Lu, c., Anderson, L. c., and Fenske, R. A. (1997). Determination of atrazine levels in whole saliva and plasma in rats: Potential of salivary monitoring for occupational exposure. 1. Toxicol. Environ. Health 50, 10 I -111. Maroni, M., Colosio, c., FerioIi, A., and Fait, A. (2000). Introduction. Toxicology 143, 5-8; Organophosphorus pesticides. Toxicology 143,9-37. Matthews, H. B. (1994). Excretion and elimination of toxicants and their metabolites. In "Introduction to Biochemical Toxicology" (E. Hodgson and P. E. Levi, eds.), 2nd ed., Chap. 8. Appleton & Lange, East Norwalk, CT. Nigg, H. N., Stamper, J. H., Deshmukh, S. N, and Queen, R. M. (1991). 4,4'Dichlorobenzilic acid urinary excretion by dicofol pesticide applicators. Chemosphere 22,365-373. Orberg, A. (1980). Observations on 2,4-dichlorophenoxyacetic (2,4-D) excretion in the goat. Acta Pharmacol. Toxicol. 46, 78-80. Osterloh, J. D., and Feldman, B. J. (1993). Urinary protein markers in pesticide applicators during a chlorinated hydrocarbon exposure. Environ. Res. 63, 171-181.
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Pinto et al. (1965). Pritchard (1976). Pritchard (1978). Pritchard, J. B., and James, M. O. (1979). Determinants of the renal handling of 2,4-dichlorophenoxyacetic by winter flounder. 1. Pharmacol. Exp. Ther. 208, 208-286. Pritchard, J. B., and James, M. O. (1982). Metabolism and urinary excretion. In "Metabolic Basis of Detoxication: Metabolism of Functional Groups" (w. B. Jakoby, J. R. Bend, and J. Caldwell, eds.). Academic Press, San Diego. Reddy, P. N., Raj, G. D., and Dhar, S. C. (1991). Toxicological effects of an organophosphorus pesticide (dimethoate) on urinary collagen metabolites in normal and high protein diets fed female albino rats. Life Sci. 49, 13091318. Schinkel, A. H., Smit, J. J. M., van TeIIingen, 0., Beijnen, J. H., Wagenaar, E., van Deemter, L., Mol, C. A. A. M., van der Valk, M. A., RunbanusMaandag, E. C., te Riele, H. P. J., Berns, A. J. M., and Borst, P. (1994). Disruption of the mouse mdrla p-glycoprotein gene leads to a deficiency in the blood-brain barrier and to increased sensitivity to drugs. Cell 77, 491-502. Tarloff, L. B. (2000). Biochemical mechanisms of renal toxicity. In "Introduction to Biochemical Toxicology" (E. Hodgson and R. C. Smart, eds.), 3rd ed. WiIey, New York. Van Welie, R. T. H., van Dijck, R. G. J. M., and Vermeulen, N. P. E. (1992). Mercapturic acids, protein adducts, and DNA adducts as biomarkers of electrophilic chemicals. Crit. Rev. Toxicol. 22, 271-306. Vasilic, Z., Drevenkar, v., Rumenjak, v., Stengl, B., and Frobe, Z. (1992). Urinary excretion of diethylphosphorus metabolites in persons by quinalphos or chlorpyrifos. Arch. Environ. Contam. Toxicol. 22, 351-357. Vasilic, Z., Stengl, B., and Drevenkar, V. (1999). Dimethylphosphorus metabolites in serum and urine of persons poisoned by malathion or thiometon. Chem.-Biol. Interact. 119-120,479-487. Wang, R. G. M., Franklin, C. A., Honeycutt, R. c., and Reinert, J. C. (eds.) (1989). "Biological Monitoring for Pesticide Exposure: Measurement, Estimation and Risk Reduction." Am. Chem. Soc., Washington, DC.
CHAPTER
26 Diagnosis and Treatment of Poisoning Due to Pesticides Wayne R. Snodgrass University of Texas
26.1 INTRODUCTION Pesticides have been used for many years to decrease adverse effects of a variety of pests. Pesticides may be categorized by major agricultural classes of insecticides, herbicides, and fungicides. Additional groupings are rodenticides, nematocides, molluscides, acaricides, larvacides, miticides, pediculicides, scabicides, attractants (pheromones), defoliants, desiccants, plant growth regulators, and repellants. Approximately 1000 chemical compounds, biological agents, and physical agents, sometimes marketed as various brand names and formulations, are utilized in many areas of our planet Earth. In the United States, estimates are that about two-thirds of all agricultural utilization involves herbicides and about one-sixth involves insecticides; less than 10% of usage is fungicides and other pesticides (U.S. EPA, 1992). The reader is referred to the chapter entitled Diagnosis and Treatment of Poisoning in the previous 1991 edition of this textbook for much useful and still relevant additional information. On a planetwide basis, one estimate of human toxicity due to pesticides ranges up to 3 million cases per year of acute severe poisoning with perhaps an equal number of unreported cases and over 200,000 deaths per year (Ferrer et aI., 1995; WHO, 1990). In the United States, one estimate is 80,000 cases per year of pesticide-related illness (Coye et aI., 1986). Greater risk rates of pesticide poisonings may be expected in countries with less regulatory controls where pesticides may be used extensively. For example, about 13,000 hospital admissions and 1000 deaths associated with pesticide poisoning occur annually in Sri Lanka, which has a population of less than 15 million (Jeyaratnam,1990).
26.2 TYPES OF PESTICIDES Organophosphates and N -methyl carbamates are the pesticides that most commonly cause systemic illness. Acute severe organophosphate poisoning is one of the most life-threatening Handbook of Pesticide Toxicology Volume 1. Principles
human poisonings, but it is also treatable (atropine plus pralidoxime), often with good outcome if treatment is begun promptly. The organophosphates that cause the most illness in persons who do not work in agriculture are the moderate toxicity compounds chlorpyrifos, dichlorvos, dimethoate, malathion, and propetamphos. The organophosphates that cause the most illness in agricultural workers are the high toxicity compounds mevinphos, methomyl, methamidophos, oxydemeton, and parathion, and also the moderate toxicity compounds dimethoate and phosalone. Carbamate pesticides typically cause less severe and shorter duration toxicity compared to organophosphates due to a more reversible complex formation between the carbamate and the cholinesterase enzyme (Fleming et aI., 1997). Pyrethrin and synthetic pyrethroid insecticides cause much less systemic toxicity in humans compared to many other pesticides. Infrequent occupational poisoning and the rare occurrence of seizures have been reported (He et aI., 1989). Extremely rare deaths (a total of two reported) are attributable to anaphylaxis from allergy. Organochlorine insecticides have low acute toxicity. DDT and chlordane are no longer used in the United States due to long-term environmental persistence. DDT also has a long-term persistence in body fat; most people have low parts per billion (ppb) concentrations of DDE, a metabolite of DDT, and low ppb levels of certain metabolites of chlordane (e.g., heptachlor epoxide) in their body fat. Endrin is more rapidly metabolized, but is more acutely toxic. Food contamination by endrin has resulted in human illness with symptoms similar to encephalitis (Rowley et aI., 1987). Chlorinated hydrocarbon insecticides are stable lipophilic chemicals and usually are contained in various organic solvents or as petroleum distillates. Often the petroleum distillates or organic solvents used as vehicles for the chemicals are as toxic as the pesticides themselves, and in the event of a significant ingestion, the vehicle toxicity should be considered as well (i.e., hydrocarbon pneumonitis). Many chlorinated hydrocarbon insecticides are rapidly absorbed and produce central nervous
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syetem (CNS) toxicity. Because of the halogenated nature of these organic compounds, hepatotoxicity, renal toxicity, and myocardial toxicity also may occur. Examples of chlorinated hydrocarbon pesticides include chlordane (no longer marketed in the United States), DDT (no longer marketed in the United States), dieldrin, chlordecone (Kepone), lindane, toxaphene, and paradichlorobenzene (moth balls). Clinical manifestations after ingestion include apprehension, agitation, vomiting, gastrointestinal (GI) upset, abdominal pain, and CNS depression. Convulsions may occur at higher doses and may be preceded by symptoms of ataxia, muscle spasms, and fasciculations. In cases of ingestion, activated charcoal often is indicated. Epinephrine may be contraindicated because it may induce ventricular fibrillation as a result of sensitization of adrenergic receptors of the myocardium by chlorinated hydrocarbons. Convulsions may be treated with lorazepam or diazepam in a dose of 0.1-0.3 mg/kg administered intravenously over about 5 minutes. Methods to enhance elimination have not been successful, yet, other than as a supportive measure for hepatic and renal failure. Cholestyramine, which has been shown to bind chlordecone in the intestinal tract, may offer a means to treat chronic chlordecone poisoning and, pending further study, may have application to other agents (Boylan et al., 1978). Herbicides typically pose a low risk of acute systemic toxicity. The phenoxy herbicides [e.g., 2,4-D (2,4-dichlorophenoxy acetic acid)] may cause mild to moderate acute skin and pulmonary irritation. Peripheral neuropathy may occur after large skin exposures over a few days. Some evidence exists for a cancer risk with long-term occupational exposure (e.g., in farmers) to certain herbicides (Hoar et aI., 1986; Zahm et aI., 1997). Fungicides are a diverse group of compounds when grouped by chemical structure. Many fungicides do not pose a large risk of acute systemic toxicity; however, many are mutagenic for fungi and thus concern exists for possible chronic toxicity due to lingering traces in certain foods (Edwards et aI., 1991).
26.3 GENERAL MANAGEMENT OF ACUTE POISONING
Table 26.1 General Management of Pesticide Poisonings
Evaluate adequacy of oxygenation Ensure an open airway and adequate ventilation Assess and support vital signs Decontaminate and limit absorption Gastric lavage if within 1 or a few hours of ingestion Induce emesis if soon after ingestion and if greater than 30 minutes from health care facility Remove contaminated clothing and shoes Wash contaminated skin, eyes, hair Enhance elimination Administer activated charcoal Supportive therapy Intravenous fluids and volume replacement Electrolyte balance Pulmonary support Blood pressure and cardiovascular support Antidote(s) if available Anticonvulsants Close observation for 24 to 48 hours after therapy stopped and recovery complete
Estimation of the severity of poisoning is an important initial step. A thorough history and details of exposure and/or poisoning should be obtained, and a physical examination should be performed. Clinical evaluation includes the use of one of several available clinical scoring systems for coma and hyperactivity. These scoring systems serve as useful monitoring parameters to follow to determine whether the patient's condition is improving or deteriorating. They also are useful to semiguantitate clinically the response to therapy. Table 26.2 identifies these scoring Table 26.2 Scoring Systems for Coma and Hyperactivity
Classification of Coma
Clinical toxicology is a rapidly expanding area that encompasses both acute and chronic exposure to drugs, chemicals (including pesticides), and naturally occurring toxins. Human toxicological exposures range from acute overdoses (accidental and deliberate) to chronic exposures (environmental and occupational). Treatment of a poisoned patient that is based on pharmacological principles promotes the use of rational methods that are beneficial to the patient's recovery. There is a great need for additional prospective, randomized, controlled, blind clinical trials of treatment modalities in clinical toxicology. These are difficult studies to carry out in many instances. In the absence of such information, recommended therapies and procedures must be evaluated with appropriate skepticism and in the context of the best benefit-risk ratio for the individual patient. General management of pesticide poisoning is described in Table 26.1.
o
Asleep but can be aroused and can answer questions
2
Comatose; does not withdraw from painful stimuli; most reflexes intact; no respiratory (ventilatory) or circulatory depression
3
Comatose; most or all reflexes are absent but without depression of ventilation or circulation
4
Comatose; reflexes absent; ventilatory depression with cyanosis, circulatory failure, or shock
1+
Restlessness, irritability, insomnia, tremor, hyperreflexia, sweating, mydriasis, flushing
2+
Confusion, hyperactivity, hypertension, tachypnea, tachycardia, extrasystoles, sweating, mydriasis, flushing, mild hyperpyrexia
3+
Delirium, mania, self-injury, marked hypertension, tachycardia, arrhythmias, hyperpyrexia
4+
Above plus convulsions, coma, circulatory collapse
Comatose; does withdraw from painful stimuli; reflexes intact
Classification of Hyperactivity
26.3 General Management of Acute Poisoning
methods and criteria. The Glasgow coma scale, a more complex scale that is not shown, also is used widely to describe coma. In the United States and most developed countries there are now regional poison centers with expertise available for consultative toxicology services. These poison centers are available by telephone to physicians and toxicologists who seek information from their large toxicology data bases (Olson et aI., 1991; Rumack, 1997). Lethal dose in 50% of animals tested (LD50) and median lethal dose (MLD) determined typically in laboratory rodents have less value for quantitatively evaluating clinical toxicity in humans. These parameters are not directly or at times even proportionally extrapolatable to humans. The clinical history obtained in an overdose and/or exposure in an attempt to estimate the dose is known to be inaccurate in one-half or more of all cases. Thus, LD50 and MLD often are not used; instead, careful clinical monitoring along with the clinical history is done. A clinically useful estimate of toxicity is margin of safety (MS), which may be defined as LDl divided by ED99, that is, the lethal dose (mg/kg) in 1% of a given human population divided by the therapeutically effective dose (for therapeutic drugs; mg/kg) in 99% of the population. This is a more conservative variation of the therapeutic index (TI = LD50 -:- ED50). The LDl often may be approximated from published case reports of overdoses or exposures. The ED99 in the case of drugs may be estimated from therapeutic clinical trial data on efficacy. A drug with a high MS ratio value generally requires a considerably higher dose relative to the therapeutic dose to cause toxicity in a patient. Overall, the adage "treat the patient, not the poison" represents the most basic and important principle in clinical toxicology. 26.3.1 SKIN DECONTAMINATION
A major route of exposure to pesticides is by skin exposure and absorption. For some pesticides (e.g., some organophosphates) skin absorption may be so extensive as to result in severe poisoning. Due to the lipophi1icity of many organophosphates, skin washing with an alcoholic (ethanol) solution is recommended in addition to washing with a detergent solution. Repeated skin washing (two, three or more separate washings) with detergent solution as well as at least one or more alcoholic solution washing is recommended. Do not harshly scrub the skin, because skin abrasion perhaps may increase skin absorption of the undesired pesticide. It is important to begin skin decontamination as soon as possible after the skin spill has occurred (Fredriksson, 1961; Wester and Maibach, 1985; Wolff et aI., 1992). 26.3.2 ACTIVATED CHARCOAL
Activated charcoal currently is the single most useful agent for prevention of absorption of orally ingested chemicals (including pesticides) and drugs. In addition, for other routes of exposure (e.g., skin and inhalation) for the many pesticides that likely undergo enterohepatic cycling or enteroenteric cycling,
591
oral activated charcoal may be an effective adjunctive treatment due to its ability to adsorb pesticide and trap it in the intraluminal space of the intestine followed by rectal excretion; in effect, this provides a "sink" to trap and accelerate the removal of previously absorbed and distributed pesticide. Again, there is great need for prospective, randomized, controlled, blind clinical trials to validate the efficacy of activated charcoal even for skin and inhalation routes of exposure. In many cases of acute oral overdose, charcoal may be administered (typically 1.0 g/kg per orogastric tube of plain activated charcoal not containing sorbitol) without prior ipecac-induced emesis or gastric lavage. When the ingested agent is not adsorbed to charcoal or the situation warrants a different approach based on patient presentation and clinical judgment, other means of gastrointestinal decontamination may be considered (e.g., lavage or whole-bowel irrigation). Prospective, randomized controlled clinical trials in orally overdosed patients have indicated uniformly that charcoal alone, compared with gastric emptying procedures with or without charcoal, results in at least similar patient outcomes with fewer complications or is superior (Albertson et aI., 1989; Kulig et aI., 1985; Merigian et aI., 1990). There is a glaring lack of published data that details the adsorption or non adsorption of various pesticides to activated charcoal. There are some substances known not to be significantly adsorbed to activated charcoal. Relevant to pesticides, such substances include inorganic borates, inorganic bromides, and mineral acids and alkalais. There are some substances that at times are cited as not being adsorbed to activated charcoal, but for which there is published evidence that clinically significant adsorption occurs (50 g of activated charcoal could adsorb an adult human toxic dose). Relevant to pesticides, such substances include carbamates, N -methyl carbamates, cyanide, DDT, diazinon, malathion, and mercuric chloride. There is a great need for additional published comprehensive and detailed in vivo laboratory animal data [area under the curve (AUC), repeat plasma drug and/or chemical concentration measurements, and degree of change/nonchange of LD50 or LD901 to determine the in vivo efficacy/nonefficacy of single-dose and multiple-dose activated charcoal for a variety of important clinically encountered pesticide inhalation, dermal, or oral exposures and/or overdoses. Pesticides for which there is a great need for in vivo laboratory animal data for activated charcoal efficacy/nonefficacy include, but are not limited to, paraquat, chlorpyrifos, diazinon, organic mercurials, organic arsenicals, strychnine, lindane, and diethyl-m-toluamide (DEET). Multiple-dose activated charcoal (MDAC) involves the repeated administration (more than two doses) by oral or orogastric tube routes of activated charcoal to enhance the elimination of chemicals already absorbed into the body (typically 0.050.2 g/kg- 1 continuous intragastric infusion for 1-3 days; use only plain activated charcoal; do not use combination product that also contains sorbitol). Studies in animals and human volunteers have shown that MDAC increases chemical elimination significantly for a number of substances, but not all (e.g., substances with short half-lives and with extensive plasma protein binding may show little enhancement of rate of elimination
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from the body). Few prospective randomized controlled clinical studies of MDAC in poisoned patients have been published. The most important endpoint to be measured is a possible reduction in morbidity and/or mortality. The early use of MDAC is an attractive alternative to more complex methods of enhancing toxin elimination, such as hemodialysis and hemoperfusion, even though the latter two modalities are indicated in only a relatively small subset of patients. Generally, patients with mild to moderate into xi cations may benefit the most from MDAC. The decision to use MDAC depends on the physician's clinical judgment regarding the expected clinical outcome, the efficacy of MDAC for the specific condition of the patient, the presence of contraindications (e.g., intestinal obstruction) to the use of MDAC, and the effectiveness of alternative methods of therapy. 26.3.3 GASTRIC LAVAGE The efficacy with which gastric lavage removes gastric contents decreases with time after ingestion. Gastric lavage should be considered only if a patient has ingested a life-threatening amount of a toxic agent, usually within 1 h or a very few hours previously. Gastric lavage with a large bore tube (36-40 French size) inserted orally is a rapid way to remove some of the contents of the stomach; however, some contents may be pushed beyond the pylorus into the small intestine, increasing unwanted absorption of the chemical. There is no strong clinical evidence to support the opinion that lavage later than 1 h or a few hours after a toxic ingestion will benefit patients; this includes ingested substances with anticholinergic activity and/or substances that delay gastric emptying. Nevertheless, uncommonly in individual patients, large amounts of gastric contents have been removed by gastric lavage many hours after ingestion, usually tablet formulations, not liquids or powders.
involves the enteral administration of large volumes of an isosmotic electrolyte lavage solution that contains poly(ethylene glycol) (PEG-ELS; e.g., Colyte or GoLytely) by orogastric tube at a rapid rate until the rectal effluent becomes clear. The purpose of this procedure is to irrigate out the contents of the GI tract to prevent or decrease the absorption of toxic substances. The PEG-ELS is isosmotic and results in minimal or no detectable electrolyte and fluid changes in most patients. Published results typically show an approximately 65-75% reduction in the bioavailability of ingested drugs (solid/tablet formulations) in volunteer studies. Adverse effects include vomiting from overly rapid infusion rates. Contraindications include ileus, GI hemorrhage, GI obstruction, and GI perforation. Relative contraindications include compromised circulation and a compromised airway. The WBI procedure consists of orogastric tube administration of PEG-ELS at about 2540 ml/kg- l . The duration of infusion is determined by the goal of therapy, which may include passage of a clear rectal effluent. 26.3.6 EYE CONTAMINATION Absorption of liquid pesticide formulations may be very rapid from the eye or conjunctiva or the mucosa of the lacrimal duct and nose to which pesticides may drain from the eye. Especially lipophilic pesticides may absorb rapidly. Thus, begin as soon as possible to wash the eye with generous amounts of normal saline or water (normal saline is less irritating to the eyes). Use a gentle stream of water. Water may be poured into the eyes from a bottle or pitcher of water. Eye irrigation for 15 minutes or more utilizing a clock to assure adequate time spent irrigating is recommended.
26.4 ACUTE POISONING BY PESTICIDES 26.3.4 CATHARTICS The use of cathartics is no longer recommended routinely in the management of orally ingested poisons. There is no strong evidence to indicate that the use of cathartics improves patient outcome. Data suggest that cathartics may not alter patient outcome when adequate continuous intragastric activated charcoal is infused. When cathartics are used alone, the absorption of some substances (measured as AUC) is increased. Cathartics must not be administered repeatedly to a patient with the absence of bowel sounds. The uncommon pseudo-obstruction of the intestine during the administration of MDAC may not be prevented with the cathartic sorbitol (Longdon et al., 1992). 26.3.5 WHOLE-BOWEL IRRIGATION
Whole-bowel irrigation (WBI) is not a routine procedure, but may be used in selected patients with a reasonable expectation of good efficacy (when measured as reduction of bioavailability) and relatively less evidence of adverse effects. WBI
26.4.1 EPIDEMIOLOGY Estimates of acute human planetwide pesticide poisoning derived from mathematical models and projections range from 500,000 cases in 1972 to 25 million cases in 1990 (Levine and Doull, 1992). This compares to the 3 million cases cited at the beginning of this chapter. In 1985, an estimate of 1 million cases of acute pesticide poisoning and 20,000 deaths was accepted at a WHO informal meeting (Blondell, 1997; Levine and Doull, 1992). When arranged by individual country, greater numbers of cases and deaths from pesticides occur in less developed countries; for example, approximately 90% of acute unintentional pesticide deaths occur in less developed countries. The wide range of overall estimates occurs because of greatly varying assumptions made during calculations. Such calculated estimates often are made out of public necessity to assist public health officials in the absence of formal detailed accurate epidemiologic studies. Such estimates may not meet the usual epidemiologic standards of validity. Clearly, there is a need for better information. Regardless, the global magnitude
26.4 Acute Poisoning by Pesticides
of acute poisonings by pesticides is enormous in costs of human suffering and human deaths.
593
Table 26.4 Symptoms of Cholinesterase-Inhibitor Poisoning and Possible Drug Treatment Muscarinic (some may respond to atropine)
26.4.2 SYMPTOMS AND SIGNS
Excessive pulmonary tract secretions Sweating
Diagnosis of acute organophosphate poisoning is based on a history of exposure, clinical symptoms and signs, and a blood test of red cell cholinesterase and plasma cholinesterase. Tables 26.3 and 26.4 list symptoms and signs of cholinesteraseinhibitor poisoning. Table 26.5 lists a classification of organophosphate poisoning based on plasma pseudocholinesterase activity. Table 26.6 lists a few carbamate insecticides classified as moderately or highly toxic. The acronym MUDDLES, which stands for miosis, urination, diarrhea, diaphoresis, lacrimation, excitation of central nervous system, and salivation, may be helpful in remembering the cholinergic excess signs that may occur to cholinesterase inhibitors. Another acronym, SLUD, which stands for salivation, lacrimation, urination, and defecation, also may assist remembering cholinergic excess signs. However, both acronyms do not emphasize the most critical clinical signs that may be life-threatening; these are the pulmonary signs of bronchorrhea (wet lungs) and bronchospasm, associated with bradycardia (usually). Severe central nervous Table 26.3 Symptoms and Signs of Mild, Moderate, and Severe Cholinesterase-Inhibitor Poisoning Exposure Mild
Symptoms Anorexia Headache Dizziness Weakness Anxiety Tremors of tongue and eyelids Miosis
Moderate
Nausea Vomiting Salivation Tearing Abdominal cramps Diaphoresis
Salivation Lacrimination Miosis Bradycardia Hypotension Urinary incontinence Gastrointestinal spasms Nicotinic (may respond to pralidoxime) Muscular fasciculations followed by weakness Central nervous system (some may respond to lorazepam/diazepam) Restlessness Anxiety Insomnia Tremors Convulsions Ventilatory/pulmonary depression Circulatory collapse
system signs of coma and seizures also may occur in severe cases. Life-threatening apnea also may occur due to nicotinic receptor depolarization with resultant chest wall muscle paralysis; a fully atropinized patient may require mechanical ventilator support due to this phenomenon. It is crucial that treatment endpoints with antidotes emphasize monitoring of and improvement in bronchorrhea, bronchospasm, and bradycardia, and not miosis (small eye pupil size), because the latter is not life-threatening. Perhaps a different acronym, BBB(u), which stands for bronchorrhea, bronchospasm, and bradycardia (usually), should be offered to assist in the diagnosis of acute organophosphate or carbamate pesticide poisoning. Exposure to organophosphates may produce a broad spectrum of clinical effects that are indicative of massive overstimulation of the cholinergic system. These effects may present clinically as feelings of headache, weakness, dizziness, blurred vision, psychosis, respiratory (pulmonary) difficulty, paralysis, convulsions, and coma. A small percentage of patients
Bradycardia Muscular fasciculations Severe
Diarrhea
Table 26.5 Classification of Organophosphate Poisoning Based on Plasma Pseudocholinesterase Activity
Pinpoint and nonreactive pupils Pulmonary/ventilatory difficulty
Classification
Enzyme activity (% of normal)
Pulmonary edema Cyanosis
Mild
Loss of sphincter control
Moderate
10-20
Heart block
Severe
<10
20-50
Convulsions Coma, possible death
Note: Most patients who require pralidoxime will have a 50% or greater decrease in red blood cell (true) cholinesterase activity.
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Table 26.6 Toxicity Classification of a Few Carbamate Insecticides Moderately toxic Bufencarb Carbaryl Methiocarb Primicarb Promecarb Propoxur Highly toxic Aldicarb Aminocarb Carbofuran Dimetilan Methomyl
may fail to demonstrate miosis, a classic diagnostic hallmark. Also, a few patients may not have bradycardia. The onset of clinical manifestations of organophosphate poisoning usually occurs within 12 hours of exposure. Measurement of red cell cholinesterase usually is diagnostic; when there is a reduction to 50% or less of control values, this indicates significant poisoning and the institution of pralidoxime (2-PAM), a cholinesterase-regenerating agent. Efforts must be made to ensure that the patient does not become reexposed through contaminated clothing or reexposure to the contaminated environment. Children may show a different frequency of symptoms and signs compared with adults. In children, initial symptoms and signs of organophosphorate poisoning more often include central nervous system depression, coma, stupor, dyspnea, and flaccidity, whereas miosis, excessive salivation, cold sweaty skin, and gastrointestinal symptoms may be less frequent (Sofer et aI., 1989). Zweiner and Ginsburg (1988) reported the following pattern of clinical signs in infants and children: miosis (73%), excessive salivation (70%), muscle weakness (68%), lethargy (54%), tachycardia (49%), seizures (22%), and pulmonary failure (38%). Young children may more frequently accidentally access pesticides stored in the home or garage (Reigart, 1995). Very limited data have been published regarding human reproductive/teratology risks for pesticides (Nurminen, 1995; Restrepo et aI., 1990; Sever et aI., 1997). Some symptoms and signs that occur with exposure to organophosphates may be due to impurities. For example, irritant (skin and upper airway tract) and odor effects associated with organophosphates are likely to be due to low molecular weight mercaptans and sulfides (O'Malley, 1997). These mercaptans and sulfides also may cause headache and nausea; at times these substances also may result in bronchospasm. Thus, it may be necessary in a few cases to distinguish whether the bronchospasm is caused by the organophosphate or an impurity. An organophosphate-induced delayed neuropathy (OPIDN) may occur after an acute severe poisoning episode. Onset is
about 7-21 days after exposure. Duration may be as short as many days or as long as months to years with full recovery or effective permanent impairment. Initial flaccidity with muscle weakness in the arms and legs that results in a clumsy shuffling gait is followed by spasticity, hypertonicity, hyperreflexia, clonus, and abnormal reflexes that indicate damage to the pyramidal tracts with a permanent upper motor neuron paralysis. In some patients, recovery occurs only in the arms and hands with no recovery in the lower extremities (foot drop, spasticity, and hyperactive reflexes), which is consistent with damage to the spinal cord (Ecobichon, 1996; Keifer et aI., 1997). OPIDN occurs primarily after a very large acute dose exposure, although it may occur after accumulated repeated lower doses. The intermediate syndrome is distinct from OPIDN in the following ways: onset within 24-96 hours after recovery from acute cholinergic crisis, muscle weakness that primarily affects muscles innervated by the cranial nerves and proximal muscles, tetanic fade instead of denervation potentials on electromyography, and more rapid clinical recovery over 4-18 days compared to 6-12 months typical in OPIDN. Persistent (for several months), often mild, central nervous system neuropsychological dysfunction may occur after an episode of acute organophosphate poisoning. Examples include persistence of decreased ability to concentrate on tasks, memory impairment, lethargy, emotional lability, blurred vision, muscle weakness, nausea, headaches, night sweats, decreased performance (up to 2 years) on a World Health Organization (WHO) neuropsychological test battery (Rosenstock et aI., 1991). 26.4.3 TREATMENT
The acute excess cholinergic symptoms and signs of organophosphate (and N-methyl carbamate) insecticides may be lifethreatening in severe cases. The main therapeutic emphasis in severe cases is to treat pulmonary and cardiovascular dysfunction and to treat with specific pharmacologic drugs [i.e., atropine and pralidoxime (2-PAM)]. Decontamination of the patient's skin is another emphasis in cases of skin exposure. Oxygen and intravenous fluid therapy usually is needed in patients with severe cases of organophosphorus poisoning. Mechanical ventilation also may be needed. Modern intensive care unit monitoring and therapy available in developed countries is indicated initially in most hospitalized patients. Clinical examination and arterial blood gas monitoring guides the use of oxygen and mechanical ventilation. Intensive care treatment more typically is required after deliberate or accidental ingestion of organophosphates than for occupational exposures. Atropine (which is antimuscarinic; no antinicotinic action) is the primary drug in the initial treatment of severe organophosphorus poisoning. It is given in small but frequent repeat doses (e.g., 1-5 mg intravenously over 5 minutes, repeated at 15 minute to several hour intervals; minimal pediatric doses are 0.05-0.1 mg/kg) and titrated to achieve the therapeutic goal of decreasing life-threatening excess cholinergic signs such as
26.5 Chronic Poisoning by Pesticides Table 26.7
Table 26.8
Herbicides: Some Common Classes and Names
Examples of Chlorophenoxy Herbicides
Class
Name
2,4-D (2,4-dichlorophenoxyacetic acid)
Acetanilides
Alachlor, Metolachlor
Amides
3,4-Dichloropropionanilide
Arylaliphatic acids
2-Methoxy-3,6-dichlorobenzoic acid
595
2,4-Dichlorophenoxybutyric acid
Carbamates
Isopropyl carbanilate
Dinitroanilines
a,a,a- Trifiuor-2,6-dinitro-N, N -dipropyl-
Nitriles
2,6-Dichlorobenzonitrile
Dichlorprop Silvex (2,4,5-trichlorophenoxy propionic acid) MCPA Mecoprop
p-toluidine Substituted ureas
3-(p-Chlorophenyl)- I, I-dimethylurea
Triazines
2-Chloro-4-( ethy lamino)-6(isopropylamino loS-triazine
bronchorrhea (pulmonary rales and rhonchi heard on auscultation of the chest), bronchospasm (wheezes heard on auscultation of the chest), and bradycardia. A decrease in pulmonary secretions is evidence of clinical response once adequate doses have been administered. Doses of atropine may be titrated to maintain clear breath sounds and a heart rate (in adults) of 80100 beats per minute. The total dose of atropine per day often markedly exceeds that which is given for other purposes: for example, total doses of 1-30 g of atropine administered over several days have been reported in individual organophosphate poisoning cases (compared to the usual therapeutic dose of 1 or 2 mg). The endpoint of therapy is lack of symptoms in the absence of atropine dosing. Atropine should not be given until adequate ventilation and oxygenation have reversed hypoxia. Pralidoxime (2-PAM) is given in moderate and severe organophosphate poisoning cases. Its mechanism of action is to split apart the cholinesterase-organophosphate complex and regenerate active cholinesterase enzyme. It is effective against nicotinic, muscarinic, and central nervous system signs and symptoms. Pralidoxime has a short plasma half-life. It is perhaps best to administer it as a continuous infusion. One initial regimen suggested is a loading dose of pralidoxime of 4 mg/kg intravenously over 15 minutes followed by 3-mg/kg- I continuous intravenous infusion. The dose is adjusted subsequently based on clinical response. Current recommendations are to not administer pralidoxime for a carbamate pesticide poisoning (Medic is et al., 1996; Tusk et al., 1997). Skin decontamination may be achieved by using soap washings followed by alcohol-soap washings with tincture of green soap or a similar mixture. Rescuers and medical personnel should be protected from contamination by using rubber gloves and aprons. A separate room in the emergency department of the hospital should be used if available and washing solutions should be appropriately discarded. For poisoning by phenoxy acid herbicides (e.g., 2,4-D), plasma kinetic evidence documents considerable shortening of plasma half-life by treatment with alkaline diuresis (see Table 26.7 for a list of some herbicides; see Table 26.8 for examples of chlorophenoxy herbicides). Benefit for clinical outcome perhaps might be expected. The risk of intravenous
sodium bicarbonate fluid therapy and diuresis is relatively small compared to the toxicity of massive doses of phenoxyacid herbicides. Thus, alkaline diuresis therapy in selected cases of phenoxy acid herbicides appears to be therapeutically rational (Flanagan et al., 1990).
26.5 CHRONIC POISONING BY PESTICIDES The incidence and severity of chronic pesticide poisoning is unknown. Estimates of the prevalence of chronic pesticide poisoning apparently are not known and certainly not widely published (Maroni and Fait, 1993). One reason for this is the difficulty of defining chronic pesticide toxicity in humans. Is it the incidence of neuropsychological deficits caused by chronic pesticide toxicity? If so, how do we measured such adverse effects objectively and quantitatively? Is it the incidence of chemically induced cancer caused by chronic pesticide exposure and long-term body burdens of certain long-lasting pesticides? Quantitation of the risk for chemically induced cancer in humans is difficult in part because of the long (often 5-20 years or more), but unknown, lag time period for the cancer to appear clinically and the widely variable genetic susceptibility in different individuals. Despite these difficulties, there are certain principles and guidelines to assist in the evaluation of patients who claim to be or suspected chronic exposed to pesticides. A large number and quantity of chemicals, including pesticides, result in a likely exposure risk for all individuals. However, the magnitude of this risk for overall morbidity and mortality may be small compared to the risks of excessive alcoholic beverage intake and the risk of cigarette smoking. The production of organic pesticides is enormous: annual manufacture in the United States exceeds 1 billion pounds per year. There are approximately 1500 different ingredients in pesticides; 50,000 chemicals are in use (not including pesticides, pharmaceuticals, and food additives); there are about 4000 active ingredients in drugs; 2000 other compounds are used as excipients to promote stability and inhibit bacterial growth; 2500 additives are used for nutritional value and flavoring; and 3000 chemicals are used to prolong product life (Maugh, 1978).
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26.5.1 TYPES OF CHRONIC PESTICIDE TOXICITY
The types of chronic pesticide toxicity may be categorized by chemical class (e.g., organophosphate or organochlorine), by the presence or absence of current ongoing exposure, or by the extent of the long-term body burden (i.e., a chemical with a very long total body elimination half-time measured in months to years). Organophosphate poisoning usually is considered to be relatively acute, but following termination of exposure, clinical toxicity resolves. However, clinical data clearly document adverse effects of organophosphates that persist long beyond the last period of exposure. There are two general types of persistent organophosphate toxicity: delayed peripheral neurotoxicity and neuropsychological deficits. So-called neurotoxic esterase (NTE) or organophosphorusinduced delayed neurotoxicity poisoning is distinct from inhibition of plasma pseudocholinesterase or inhibition of erythrocyte cholinesterase. In the classic presentation of NTE-type toxicity, ataxic signs occur 1-3 weeks after an acute exposure. Painful transient paresthesias develops in a stocking-glove distribution over the lower extremities, rapidly extending to motor weakness and ataxia, with later involvement of the upper extremities. Evidence in laboratory animals points toward greater symptoms from multiple subchronic doses compared to a large single dose. Subchronic dermal exposure appears to be the most potent route of dosing. Structured-activity relationships of organic phosphonates versus phosphenates versus phosphates, chirality, and the ability to have a chemical leaving group correlate with ability to inhibit the NTE enzyme. Thus, not all organophosphates pose a risk for NTE inhibition and only a few have been documented to cause clinical delayed neurotoxicity. A lymphocyte NTE test, if available, may help to assess exposure of an individual patient (Steenland et al., 1995). The second type of delayed organophosphate toxicity consists of neuropsychological deficits. Abnormalities of memory, abstraction, mood, intellectual functioning, and flexibility of thinking can be demonstrated in subjects with a last prior organophosphate poisoning of 7 years or more in a few cases. These individuals may have neuropsychological test results similar to subjects with cerebral damage or dysfunction. These sequelae are sufficiently subtle that clinical neurological examination, clinical EEG, audiometric, ophthalmic, and blood chemistry testing cannot discriminate poisoned subjects from controls. The physician cannot rely solely on the standard neurological examination or on clinical intuition to evaluate a patient who has been poisoned chronically by organophosphates. The usual clinical neurological examination does not thoroughly assess higher level cognitive skills, but rather focuses on sensory and motor functioning. The major deficits in these patients are cognitive and appear on neuropsychological tests of abilities that receive limited emphasis in the usual neurological examination. Thus, the two methods, clinical neurological examination and neuropsychological testing, provide
a more thorough evaluation of such patients with chronic cognitive deficits following organophosphate poisoning (Eyer, 1995). Nerve conduction studies in acute organophosphate poisoning have been reported usually as normal. The muscle response to single stimulation shows a repetitive response. This repetitive response disappears on repeated stimulation and this finding is thought by some to be characteristic. Organochlorine chronic pesticide poisoning presents equally dificult challenges for evaluation. Halogenated hydrocarbons include those used as pesticides as well as many fumigants and solvents. The fumigants and solvents, usually with molecular weights less than approximately 250 Da, generally are considered to have shorter half-lives, in part due to their greater vapor pressures. The halogenated hydrocarbon insecticides, usually with molecular weights of approximately 290-550, have lower vapor pressures and include lindane (y-BHC or y-benzene hexachloride), chlordane, heptachlor, methoxychlor, aldrin, mirex, toxaphene and many others. Other halogenated hydrocarbon environmental contaminants also pose a toxicity risk for humans and include TCDD (dioxin or 2,3,7,8-tetrachlorodibenzop-dioxin), PBBs (polybrominated biphenyls), and PCBs (polychlorinated biphenyls). Many of these less volatile halogenated hydrocarbons share the properties of high lipid solubility, extensive distribution and storage into body fat, and poor metabolization with a resulting long total body half-life. 26.5.2 BACKGROUND ACCUMULATION
An especially important factor in chronic pesticide poisoning is the potential for chronic persistent accumulation in the body. This accumulation occurs in all individuals because of nonoccupational exposure through food (even though the absolute amount may be very low), inhalation, or skin absorption. Thus, a background concentration (e.g., in body fat) of certain halogenated hydrocarbons, including PCBs, DOT, DDE (metabolite of DDT), hexachlorobenzene, dieldrin, lindane and its isomers, and sometimes TCDD (dioxin), pentachlorophenol, and BHT (buty lhydroxytoluene), is detectable in the general human population. Also in body fat, lower concentrations of consistently in nonoccupationally exposed individuals chemicals such as aldrin, heptachlor, oxychlordane, mirex, chlordecone (Kepone), DDD, PBBS, chlorinated benzenes, polychlorinated terphenyls, tetrachlorophenol, toxaphene, phthalate esters (e.g., DEHP, used as a plasticizer in plastics), and a variety of more volatile chlorinated aliphatic hydrocarbons are detected (sometimes less consistently). Thus, a definite body burden of halogenated hydrocarbons exists, a priori, in every patient. Subtle neuropsychological testing of absolutely unexposed individuals (controls), including infants and children, does not exist; all individuals have some exposure to halogenated hydrocarbons. The potential for accumulation can be assessed by use of a bioconcentration factor (BCF). The BCF is the ratio between the concentration of a chemical in tissue (e.g., fat) and the concentration of the same chemical in the diet (both measured
26.5 Chronic Poisoning by Pesticides
in mg/kg). A BCF of 1 or less means that no accumulation has occurred. The BCF for DDT plus metabolites is 1279; the BCF for hexachlorobenzene is 674; the BCF for lindane is 18. Some of these numbers are rather sobering. Such data are valuable in assessing relative risk for body accumulation. Another conceptually useful method for assessment of chronic toxicity, particularly in occupational exposure, is the chronic exposure index (CEI), CEI
= log (base 10)[y x
D divided by age - 18, then - 1]
where y is the number of years of exposure to a pesticide and D is the most recent estimate of number of days of usage of pesticides per year. Index values from the median value to the highest value for a particular group of subjects is defined as high chronic exposure; those from the lowest to the median value are called low chronic exposure. In addition, knowledge of relative chronic toxicity data also is useful in assessing overall risk as is information on specific mechanisms such as metabolite-related tissue binding of halogenated hydrocarbons that result in more selective tissue toxicity. 26.5.3 SYMPTOMS AND SIGNS Symptoms and signs of chronic organochlorine and/or halogenated hydrocarbon pesticide poisoning are illustrated by the cyclodiene pesticides, which include chlordane, heptachlor, aldrin, and dieldrin. Two types of chronic exposure syndromes occur. In one, the exposure is continuous and leads to a slow accumulation of insecticide along with progressive symptoms. In the second type of chronic syndrome, insecticide exposure remains below that needed to cause symptoms; however, the individual experiences adverse effects with further intake. In 10-20% of sprayer operators who apply dieldrin and became poisoned, the earliest clinical signs developed in 3 months of exposure, but most required 8 months or more of exposure. Mild illness consisted of headache (that often was unresponsive to drugs and was persistent), dizziness, general malaise, insomnia, nausea, increased sweating, nystagmus, diplopia, tinnitus, slight involuntary movements, and blurred vision. Severe illness included progression to myoclonic jerking involving one or more limbs, sometimes accompanied by brief loss of consciousness. Approximately half of the 10-20% of cases progressed to convulsions. Persistent neurological sequelae from cyclodiene chronic poisoning include EEG abnormalities that last long after objective signs of toxicity have resolved. Normalization of the EEG after exposure has stopped may require up to 3 months for endrin, a year for dieldrin, and more than a year for telodrin (Ecobichon and Joy, 1982). During continued chronic exposure to a mixture of chlorinated hydrocarbons, organophosphates, and carbamates in asymptomatic occupationally exposed workers, chronic elevation of serum epinephrine and of serum glucose (77 mg/dl versus 127 mg/dl, controls versus exposed) occurred. Thus, neurologic effects of chronic pesticide exposure apparently include sympathetic stimulation of the adrenal glands.
597
Similar data for altered neurological and altered dopamirie and pituitary function are reported for exposure to the hydrocarbon styrene. Chronic poisoning from halogenated hydrocarbons results in measurable neurophysiological and neuropsychological abnormalities. Chronic toxic encephalopathy, once established, improves only slightly or not at all with time. Older individuals are more severely affected and less likely to recover. In one study, psychometric retesting 4 years later (4 years exposurefree) showed significant deterioration in verbal memory with improvement in visual memory. Computed tomography of the brain mayor may not demonstrate loss of brain substance. Similarly, persistent long-term neuropsychological effects of TCDD, a halogenated hydrocarbon contaminate of 2,4-D but not itself a pesticide, have been studied. No clinically evident neurotoxicity was noted in an unreported study cited by Young (1984) of prison volunteers in the mid-1960s who were exposed dermally to 2,3,7,8-TCDD. The experiment was done to determine the dose of TCDD required to induce chloracne. A suspension of 1% TCDD in chloroform/ethanol was applied to the backs of 10 subjects on alternate days for 1 month. The cumulative applied dose was 7500).!g of TCDD. Eight of the ten subjects developed chloracne, which lasted 4-7 months. Blood chemistries including hematologic, liver, and kidney function remained within normal limits. TCDD is highly persistent with a whole-body half -life of approximately 5 years. Poiger dosed himself orally with 105 ng of 3H-TCDD (tritium-TCDD) in corn oil. The dose of 1.14 ng/kg contained only 13 ).!Ci of tritium. Approximately 88.5% of the dose was absorbed. Radioactivity was detected only in the feces and was excreted at a rate of 0.03% of the body burden per day; the wholebody half-time was estimated to be 5.8 years. Pharmacokinetic approaches to such long half-time chemicals still are being developed. The Bhopal incident of massive methyl isocyanate exposure is an example of chronic chemical exposure. Methyl isocyanate is not a pesticide, but does have a chemical structure that is similar to some pesticides. Although often considered a relatively short-term exposure, increasing evidence exists that the methyl isocyanate exposure at Bhopal has resulted in persistent longterm sequelae in survivors. The primary sequelae is chronic pulmonary damage, a combined obstructive and restrictive type. Methyl isocyanate-specific antibodies have been detected in the serum of some of the victims upon follow-up and the antibody titer seems to correlate with the severity of the lung injury. Of over 200,000 people exposed to methyl isocyanate at Bhopal, estimates are that up to 5000 died within 2 days and about 60,000 individuals require long-term medical management. 26.5.4 ASSESSMENT OF EXPOSURE HISTORY One of the most difficult aspects in evaluating a patient who presents with or claims to have chronic pesticide poisoning is obtaining a meaningful medical history. Individuals with legitimate medical toxicologic events may be unable to reconstruct
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Poisoning Due to Pesticides
a completely useful history despite skillful questioning. Those individuals who present initially and have already decided they will pursue legal redress for their exposure often have attributed (sometimes unknowingly) many minor and nonspecific complaints to the alleged exposure. In an attempt to obtain a more consistent and thorough history, utilization of somewhat standardized clinical toxicology patient history and physical examination forms may be useful. The primary purpose of such a form is to assist in the completeness of the evaluation and, secondarily, to expedite the process. Such a form is particularly useful in the setting of an environmental-occupational toxicology clinic. The Agency for Toxic Substances Disease Registry has a useful form that is available. Risk analysis is another process the medical toxicologist enters into as he or she obtains an exposure history. Scientists have formulated various approaches to attempt to add precision and provide a more solid basis for governmental regulatory decisions regarding exposures to chemicals. However, for the physician evaluating an individual patient, obtaining as complete as possible the extent of exposure history and general knowledge of human toxicity effects for the individual chemical(s) usually is the roost common historical information available for making a diagnosis and patient management recommendations. Equally important to chemical exposure information is information regarding the patient's general medical status, past medical history, and any specific susceptibilities, for example, known history of allergies or family history of cancer. Knowledge of the individual chemical(s) bioconcentration factor may be of use for selected chemicals. For infants it is useful to know that breast-feeding newborn infants potentially may be subjected to much higher halogenated hydrocarbon pesticide intake per kilogram of body weight than are adults due to the much higher concentration of some of these chemicals in human milk. Indeed, milk is the major route of excretion of body burdens of highly fat-soluble chemicals such as dioxin. To quote Matsumura (1985), "If a 5-kg infant consumed 0.7liter of human milk daily containing DDT at an average concentration of 0.08 ppm, the resulting dose would be 0.0112 mg/kg/day. This value may be compared with the average adult daily intake of 0.0005 mg/kg/day. Assuming a 2- to 3-fold increased sensitivity to the toxic effects of DDT in infants compared to adults, this amounts to 8 times more than the FAO-WHO recommended maximum acceptable dose!" Similar considerations for lindane isomers (total benzene hexachloride) reveal that up to 1000 times more than the daily adult per kilogram intake may be "dosed" to infants (Matsumura, 1985)! Subcutaneous fat concentrations of chlorinated hydrocarbons from infants and children sampled in 1982 reflected a highly significant association with the quantity of mother's milk consumed. Some individual fat concentrations were higher than the mean values for adults from other areas. Because hydrocarbons have been documented to alter the developing brain biochemically in young laboratory animals, these human data have to be viewed with concern.
Increasing data are available on quantitation of pesticide exposure in various situations. Home gardening exposure to the carbamate carbaryl is approximately 8.5 mg per 10 g pesticide applied to garden plants. Protective clothing reduced exposure by 20-fold. Whereas over 90% of surburban residents in the United States are reported to use pesticides around the home, such exposure has considerable significance (Kurtz and Bode, 1985). Similarly, aerial pesticide spraying requires further published quantitative data that are lacking currently. 26.5.5 ASSESSMENT OF SYMPTOMS Assessment of symptoms is a subjective evaluation. This is particularly difficult in chronic pesticide poisoning for at least two reasons: (1) many symptoms are nonspecific (e.g., headache, intermittent dizziness) or vague (e.g., general malaise) and (2) in the United States the patient may be planning legal action or be seeking workmen's compensation relating to the pesticide exposure and thus may perceive his or her symptoms differently. It is important for the physician to exercise skillful judgment and, if possible, rely on prior experience. In the end, a judgment of the relevance of the complaints to the chemical exposure has to be made. The lay public's perception of the relative risks of a variety of possible hazards may not correlate with actual relative risks (Baird, 1986). Interestingly, pesticides rank higher for greater perceived risk than, for example, motor vehicle accidents or alcoholic beverage-related risks. Physicians and scientists likely might rank certain hazards quite differently compared to the general public. However, such information about how the lay public perceives risk is useful in assessing which item and to what degree that item is likely to concern many patients. Risks due to naturally occurring contaminants are accepted about 20 times more readily that those due to man-made sources. Risks taken voluntarily are about 100 times more acceptable than risks taken involuntarily (Baird, 1986). 26.5.6 ASSESSMENT OF SIGNS Assessment of signs generally is an objective evaluation. In chronic pesticide poisoning, the physical examination should be done with particular attention toward target organs for toxicity of a particular chemical, if known. The neurological examination and close scrutiny for signs compatible with cancer are important when there is a long history of exposure. Biological monitoring is becoming increasingly sophisticated and helpful in the initial assessment of chronic pesticide exposure and for follow-up (He, 1993). Biological monitoring includes a variety of approaches, for example, direct measurement of blood, urine, or fat levels of a specific chemical, measurement of a metabolite in urine or of a toxicologically relevant (e.g., thioether compounds) metabolite in urine, or measurement of a biochemical effect relevant to toxicity risk (e.g., DNA adducts in urine; urine porphyrins). Unlike environmental monitoring (e.g., air level measurements of a specific
26.5 Chronic Poisoning by Pesticides
chemical), biological monitoring can assess exposure by all routes, not just by the inhalation route. Biological monitoring has the potential to assess the actual uptake of a chemical by an individual. Also, biological monitoring, depending on the specific parameter, may take into account, at least in part, individual variability for risk from a given level of exposure. Biomonitoring of carbamate pesticides has been reported utilizing hemoglobin adducts (Sabbioni et aI., 1990). A biological exposure index (BEl) theoretically can be developed for many specific chemicals. For trichloroethylene (not a pesticide), the BEl is 100 mg trichloroacetic acid per liter of urine with the specimen collected at the end of the work week; that is, this is the acceptable upper limit of exposure, with greater concentrations indicating excessive exposure that reguires further investigation (Lowry, 1987). One useful approach to biological monitoring for organophosphate pesticides, especially for low-level exposure, has been the repeated sequential measurement of plasma pseudocholinesterase to monitor organophosphate exposure of agricultural field workers and those exposed to pesticide spray drift. Small but definite intraindividual plasma pseudocholinesterase differences over time were found in individuals who were intermittently exposed to organophosphate spray drift compared to those not exposed. Measurement of urinary alkylphosphates may be a more sensitive method for biological monitoring of low level organophosphate exposure (Sunaga et al., 1989). Sequential monitoring for specific pesticides also appears to be indicated for indoor work areas such as greenhouses. Biological monitoring for risk of developing OPIDN (neurotoxic esterase inhibition) appears possible with the development of a lymphocyte and/or platelet NTE assay. Validation by careful clinical studies is needed prior to widespread application of this test. The porphyrin excretion pattern in the urine may become a useful biological monitor for exposure to certain pesticides and chemicals. Compounds that are considered to be porphyrinogenic include some organophosphorus and organochlorine pesticides as well as PCBs, PBBs, TCDD, vinyl chloride, and chlorinated naphthalenes (Strik, 1987). Chemically induced porphyria occurs due to inhibition of uroporphyrinogen decarboxylase, an enzyme that is part of heme synthesis. Inhibition of this enzyme results in accumulation and increased excretion of uroporphyrin and heptacarboxylic porphyrin. Only small elevations in urinary porphyrins have been seen generally in individuals exposed to chemicals such as PCBs. In one study, a few individuals following exposure to PCBs had a urinary uroporphyrin excretion of 66-106 I-lg per 24 hours compared to a nonexposed group that had less than 60 I-lg per 24 hours (Osterloh et aI., 1987). Other studies have suggested that a total porporphyrin (not uroporphyrin) excretion in adults of up to 200 I-lg per liter of urine is normal (Strik, 1987). The pattern of porphyrins excreted rather than the total amount appears to be more significant for biological monitoring purposes. An increase in uroporphyrin and heptacarboxylic porphyrin may be an early finding in chemically induced porphyria. An increase
599
in uroporphyrin and a decrease in coproporphyrin was seen following PCB exposure. Another potentially useful and rational approach to biological monitoring for selected pesticides and chemicals is the measurement of urinary thioethers. A usual but not necessary requirement is that the chemical be metabolically activated to an electrophilic (positively charged) short-lived toxic metabolite by the cytochrome P-450 microsomal mixed function oxidase drug metabolizing enzymes. In addition to electrophilic intermediates, such chemicals as aliphatic and aromatic halides and a-f3 unsaturated ketones may undergo direct (nonglutathione transferase) conjugation with glutathione. The technique exploits the fact that conjugation with glutathione, followed by urinary elimination as mercapturic acid metabolites, is a significant metabolic pathway, especially for putative toxic metabolites. Urinary thioethers potentially may indicate the extent of internal contamination of an exposed individual. Gasoline station attendants who dispensed gasoline directly (fully-serve) versus those attendants who did not (self-serve) had urinary thioether concentrations (median value of 7-8 pm single urine void) of approximately 8 I-lmol-SHlmmol creatinine versus approximately 2 I-lmol-SHlmmol creatinine, respectively (Stock and Priestly, 1986). Although urinary thioether output detects dose-related increases in chemical exposure in animal models, this apparently has not been demonstrated yet in published clinical studies. Inhibition of red blood cell glutathione-S-transferase potentially may be another useful marker of toxic chemical body burden. Some organochlorine and halogenated hydrocarbon pesticides (many of which are no longer marketed in the United States are known to be metabolically activated and also known, upon chronic high dose exposure, to be carcinogenic in laboratory animals. Assessment of chronic exposure to these or similar pesticides by urinary thioether measurement in the future may become a useful biologic monitor. Certainly this approach deserves further careful, well-designed, prospective, clinical studies. An exciting and potentially highly significant biological monitor is DNA adduct formation. DNA adduct formation represents a direct chemical covalent bond between one of the bases of DNA and a chemical. The combination may be measured as a DNA adduct. Such adduct formation may be one of the initial steps necessary for chemically induced carcinogenesis. One has to be very cautious in attempting to apply such information clinically unless solid data validate a correlation between a given monitoring technique and the risk of human cancer. DNA adduct measurement in the white blood cells of iron foundry workers correlated quantitatively with air concentrations of polycyclic aromatic hydrocarbons such as benzpyrene (not a pesticide). The mean adduct levels (femtomoles of adduct per microgram of DNA) ranged from 0.24 to 1.5 fmol/).Lg (or, expressed differently, 0.80-5.0 mean adducts per 10 exponent 7 nucleotides) and correlated with air concentrations of <0.05 to >0.2 I-lg of benzpyrene per cubic meter of air in the workplace. Similarly, vinyl chloride (not a pesticide), known to cause angiosarcoma of the liver in humans, reacts
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Poisoning Due to Pesticides
to form DNA adducts. One has to be cautious in interpreting data. For example, although styrene (not a pesticide; used in the manufacture of plastics) produces chromosome aberrations in cultured human blood lymphocytes in vitro, workers exposed to low levels of styrene show no measurable chromosome abnormalities. Hemoglobin adducts of certain chemicals represent another biological monitoring endpoint and are described for ethylene oxide, benzene, benzpyrene, and dimethylnitrosamine (Farmer et aI., 1987; Perera et aI., 1988). Thus, many techniques exist or are being developed for biological monitoring. Some will prove to be more predictive than others. Biological monitoring of chronic pesticide exposure will expand as mechanisms of toxicity are elucidated further and new techniques advance. 26.5.7 WORKUP OF NEUROTOXICITY
The workup of neurotoxicity in individual patients can be divided into the assessment of the peripheral nervous system (PNS) and the central nervous system. In general, somewhat more objective measurements (neurophysiologic testing) can be made in examination of the PNS compared to the CNS (neuropsychologic testing) with the exception of objective CNS evoked potential measurements. For PNS function, it is important to remember that axonal neuropathies often are associated with deficits of both sensory and motor function. Motor function testing includes (1) inspection for muscle atrophy, unusual movements, and an analysis of coordination; (2) testing muscle tone and resistance to passive stretch of an extremity; (3) the Babinski reflex; and (4) analysis of the strength of individual muscles. Sensory function testing includes evoking the sensations produced by warmth and cold, pinprick, joint movement, tuning fork vibration, and shapes of complex objects. Cranial nerve examination, especially optic nerve (cranial nerve 11) and trigeminal nerve (cranial nerve V) function, is important in evaluation of toxic exposures. Similarly, evaluation of the autonomic nervous system for bladder, bowel, and sexual functions, pupil response, lacrimation, salivation, sweating, and supine-upright blood pressure is important (Spencer et aI., 1985). Knowledge of the specific toxin is helpful in planning and analyzing PNS evaluation. For example, in acrylamide (not a pesticide) neuropathy, sensory symptoms and signs are prominent. By contrast, NTE neuropathy from organophosphates may show retention of sensory function in the face of significant distal wasting and weakness. Nerve conduction velocity may not be altered in organophosphate NTE neuropathy, but a large drop in muscle action potential amplitude (EMG testing) may be observed. Acrylamide toxicity can be detected earlier by monitoring vibration sensation in fingers (e.g., using a portable Optacon device; Spencer et aI., 1985). In hexacarbon solvents (e.g., n-hexane or methylbutyl ketone; not pesticides), toxic neuropathy, use of nerve motor conduction velocity measurements is valuable because these solvents affect conduction. Blue-yellow color vision loss from solvent exposure is well
documented. The Lanthony D-15 de saturated panel test for color vision loss has been validated for use at the workplace as an initial screening tool. Perhaps some of the foregoing testing procedures should be studied for their possible validity to monitor workers exposed to various pesticides. For CNS function, evaluation of mental status includes assessments of the level of consciousness, orientation, concentration, memory, cognitive functions, behavior, mood, and affect. The most frequently reported behavioral effect of chemicals is a disturbance in psychomotor functioning. Usually, this is characterized by a delay or slowness in response time, clumsy or awkward eye-hand coordination or dexterity, or a combination of these. Diminished attention also has been found (Feldman et al., 1980). For a thorough evaluation of suspected neuropsychological deficits, use of a standard battery (group) of psychological tests is the best available procedure. Examples of such tests include the Halstead-Reitan battery, the LuriaNebraska battery and the Pittsburgh Occupational Exposure Test (POET) battery. Formal testing is time-consuming and usually is done by psychologists trained in the use of such tests. There is a need for standardized neuropsychological tests that have a sufficient degree of simplicity and speed of administration such that they can be administered more readily to blue collar industrial workers as well as individuals with nonoccupational environmental exposures to pesticides. Furthermore, such test results must be adjusted for age and educational level (Feldman et aI., 1980). One example of a neuropsychological test battery with published normative data from a control population with no known previous history or occupational exposure to industrial chemicals is the POET battery (Ryan et aI., 1987). Other descriptions of neuropsychological testing results support the use of computed tomography x-ray scans of the head and targeting specific neuropsychological tests for specific exposure situations. A few specific examples of chemicals for which formal neuropsychological testing for exposure effects has been done include organophosphates, inorganic lead, carbon disulfide, mercury, styrene, perchloroethylene and trichlorethylene, chlorinated solvent mixtures such as that of methylene chloride plus Freon 113 plus trichloroethylene, and the fumigants methyl bromide and sulfuryl fluoride. 26.5.8 RISKS OF PESTICIDE-INDUCED CANCER
Risks for cancer from occupational exposure to some chemicals, including pesticides, have been estimated. However, for a given individual, risk is nearly impossible to estimate with currently available data. Individuals vary widely in their genetic susceptibility, their exposures to other chemicals (e.g., cigarette smoke and ethanol), diet content, and other less well defined parameters. Ecogenetics (distinct from pharmacogenetics) is the study of genetically controlled variations in response to environmental substances other than drugs. One example of the toxicological consequences of genetic differences is the 40-fold range among normal subjects in
References the capacity to metabolize chemical carcinogens by a specific cytochrome P-450 dependent isozyme, aryl hydrocarbon hydroxylase (AHH; Vesell, 1987). High activities are associated with lung carcinomas. It is likely that increased AHH activity would yield higher amounts of toxic carcinogenic reactive metabolites with exposure to polycyclic aromatic hydrocarbons (and perhaps pesticides with similar chemical structures. Such reactive metabolites bind covalently to DNA and initiate the development of cancer. Similar considerations apply to bladder cancer risk from aromatic amine exposure in the rubber industry (what about aromatic amine herbicides?). Numerous examples of chemical-induced cancer exist. An increased risk for leukemia has been reported in children whose parents use pesticides in the home or the garden, or whose fathers are exposed occupationally to chlorinated solvents, or dyes or pigments (Lowengart et aI., 1987). Phenoxy acid herbicide agriculture use appears to pose a risk for non-Hodgkin's lymphoma in farmers and forestry worker who use sprays (Zahm et aI., 1997). Workers engaged in the storage and bulk handling of agricultural grains and peanuts in Sweden were reported to have an elevated risk for primary liver cancer; aflatoxin exposure is suspected. Use of insecticides and fumigants such as aluminum phosphide, carbon disulfide, carbon tetrachloride, ethylene dibromide, ethylene dichloride, malathion, and methyl bromide in flour milling may be implicated in the increased risk of non-Hodgkin's lymphoma in flour mill workers. 26.5.9 TREATMENT Treatment of chronic pesticide poisoning or exposure requires careful medical evaluation and a thorough assessment of the exposure situation. The recommendation that the patient be removed from further exposure may be followed more easily for nonoccupational exposures. For occupational exposures, removal of the worker to another area of the worksite where lower exposure occurs may, on occasion, be acceptable medically. Alternatively, temporary removal while clean-up measures are begun also, on occasion, may be acceptable medically. Monitoring exposure at the worksite by a medically suitable means should be a part of all recommendations and follow-up evaluations of workers with chronic pesticide toxicity. Treatment of symptoms, other than removal from further exposure, often may be symptomatic only. Protective clothing can be recommended and should be worn for certain chemical exposure situations. More published data are becoming available to identify improved protective clothing materials, for example, polyvinyl alcohol polymer material to resist permeation of methylene chloride (Stampfer et aI., 1984). Gastrointestinal dialysis may offer the possibility of enhancing the rate of removal of the body burden of certain chemicals/pesticides that have long residence times (measured in months to years) and are sequestered in kinetically deep storage sites (usually body fat). A well documented example was the use of repeated oral cholestyramine therapy to increase the
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removal of chlordecone (Kepone), a highly neurotoxic chemical. Other examples include the use of repeated oral low dose mineral oil for a relatively short limited time to enhance excretion of chlordane and lindane (Snodgrass et aI., 1983, 1986). 26.5.10 SUMMARY In summary, diagnosis and management of both acute and chronic pesticide poisoning presents a significant challenge to current clinical capabilities. With careful attention to the details of exposure history and clinical findings in each individual patient, coupled with knowledge of the specific pesticide involved, an individualized management plan can be formulated that should be beneficial for most patients.
REFERENCES Albertson, T. K, et al. (1989). Superiority of activated charcoal alone compared with ipecac and activated charcoal in the treatment of acute toxic ingestions. Ann. Emerg. Med. 18,56-59. Baird, B. N. (1986). Tolerance for environmental health risks: The influence of knowledge, benefits, voluntariness and environmental attitudes. Risk Anal. 6,425. BlondeII, J. (1997). Epidemiology of pesticide poisonings in the United States, with special reference to occupational cases. Occup. Med. 12, 209-220. Boylan, J. L., Egle, J. L., and Guzelian, P. D. (1978). Cholestyramine: Use as a new therapeutic approach for chlordecone poisoning. Science 199, 893895. Coye, M. J., Lowe, J. A., and Maddy, K T. (1986). Biological monitoring of agricultural workers exposed to pesticides. 1. Cholinesterase activity determinations. J. Occup. Med. 28,619-627. Ecobichon, D. J. (1996). Toxic effects of pesticides. In "Casarett and DouII's Toxicology" (C. D. Klaassen, ed.), pp. 643-689. McGraw-HiII, New York. Ecobichon, D. J., and Joy, R. M. (1982). "Pesticides and Neurological Diseases." CRC Press, Boca Raton, FL. Edwards, R., Ferry, D. G., and Temple, W. A. (1991). Fungicides and related compounds. In "Handbook of Pesticide Toxicology. Classes of Pesticides" (w. J. Hays, Jr. and E. R. Laws, Jr., eds.), Vol. 3, pp. 1409-1470. Academic Press, New York. Eyer, P. (1995). Neuropsychological changes by organophosphorus compounds: A review. Human Exper. Toxicol. 14,857-864. Farmer, P. B., et al. (1987). Estimation of exposure of man to substances reacting covalently with macromolecules. Arch. Toxicol. 60,251. Feldman, R. G., Ricks, N. L., and Baker, E. L. (1980). Neuropsychological effects of industrial toxins: A review. Amer. J. Indus. Med. 1,211-227. Ferrer, A., et al. (1995). Recent epidemics of poisoning by pesticides. Toxicol. Lett. 82-83, 55-63. Flanagan, R. J., et al. (1990). Alkaline diuresis for acute poisoning with chlorophenoxy herbicides. Lancet 335, 454-458. F1eming, L. E., et al. (1997). Emerging issues in pesticide health studies. Occup. Med. 12,387-397. Fredriksson, T. (1961). Percutaneous absorption of parathion and paraoxon. IV. Decontamination of human skin from parathion. Arch. Environ. Health 3, 185-188. He, E (1993). Biological monitoring of occupational pesticide exposure. Int. Arch. Occup. Environ. Health 65(Suppl. 1), S69-S76. He, E, Wang, S., Liu, 1., Chen, S., Zhang, Z., and Sun, J. (1989). Clinical manifestations and diagnosis of acute pyrethroid poisoning. Arch. Toxicol. 63, 54-58. Hoar, S. K, Blair, A., Holmer, E E, Boysen, c., Robel, R. J., Hoover, R., and Fraumeni, J. E (1986). Agricultural herbicide use and risk of lymphoma and soft tissue sarcoma. J. Amer. Med. Assoc. 256, 1141-1147.
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leyaratnam, 1. (1990). Acute pesticide poisoning: A major global health problem. World Health Stat. Q. 43, 139-144. Keifer, M. c., et al. (1997). Chronic neurologic effects of pesticide overexposure. Occup. Med. 12,291-304. Kulig, K. W, Bar-or, D., Cantrill, S. v., et al. (1985). Management of acutely poisoned patients without gastric emptying. Ann. Emerg. Med. 14,562-567. Kurtz, D. A., and Bode, W. M. (1985). Application exposure to the home gardener. In "Dermal Exposure Related to Pesticide Use" (R. C. Honeycutt, G. Zweig, and N. N. Ragsdale, eds.), American Chemical Society, Washington, DC. Levine, R. S., and Doull, 1. (1992). Global estimates of acute pesticide morbidity and mortality. Rev. Environ. Contam. Toxicol. 129,29-50. Longdon, P., et al. (1992). Intestinal pseudo-obstruction following the use of enteral charcoal and sorbitol. Drug Safety 7, 74-77. Lowengart, R. A., et al. (1987). Childhood leukemia and parents occupational and home exposure. J. Nat. Cancer Inst. 79,39-46. Lowry, L. K. (1987). The biological exposure index: Its use in assessing chemical exposures in the workplace. Toxicology 47, 55-69. Maroni, M., and Fait, A. (1993). Health effects in man from long-term exposure to pesticides. Toxicology 78, 1-174. Matsumura, F. (1985). Hazards to man and domestic animals. In "Toxicology of Insecticides", 2nd ed. (F. Matsumura, ed.). Plenum, New York. Maugh, T. H. (1978). Chemicals: How many are there? Science 199, 162. Medicis, 1. 1., et al. (1996). Pharmacokinetics following a loading dose plus a continuous infusion of pralidoxime compared with the traditional short infusion regimen in human volunteers. J. Toxicol. Clin. Toxicol. 34, 289295. Merigian, K. S., et al. (1990). Prospective evaluation of gastric emptying in self-poisoned patients. Amer. J. Emerg. Med. 8,479-483. Nurminen, T. (1995). Maternal pesticide exposure and pregnancy outcome. J. Occup. Med. 37, 935-940. Olson, D. K., et al. (1991). Pesticide poisoning surveillance through regionl poison control centers. Amer. J. Public Health 81, 750-753. O'Malley, M. (1997). Clinical evaluation of pesticide exposure and poisonings. Lancet 349, 1161-1166. Osterloh, 1., et al. (1987). Pilot survey of urinary porphyrins from persons transiently exposed to a PCB transformer fire. J. Toxicol. Clin. Toxicol. 24, 533. Perera, F. P., et al. (1988). Detection of polycyclic aromatic hydrocarbon DNA adducts in white blood cells of foundry workers. Cancer Res. 48, 2288. Reigart, 1. R. (1995). Pesticides and children. Pediatric Ann. 24,663-668. Restrepo, M., et al. (1990). Birth defects among children born to a population occupationally exposed to pesticides in Colombia. Scand. 1. Work Environ. Health 16, 239-246. Rosenstock, L., Keifer, M., Daniell, W. E., et al. (1991). Chronic central nervous system effects of acute organophosphate pesticide intoxication. Lancet 338,223-227. Rowley, D. L., Rab, M. A., Hardjotanoho, W., et al. (1987). Convulsions caused by endrin poisoning in Pakistan. Pediatrics 79, 928-934. Rumack, B. H. (1997). "Poisindex." Micromedex, Denver, CO. Ryan, C. M., et al. (1987). Assessment of neuropsychological dysfunction in the workplace: Normative data from the Pittsburgh Occupational Exposures Test Battery. J. Clin. Exper. Neuropsychol. 9,665-679.
Sabbioni, G., et al. (1990). Biomonitoring of arylamines: Hemoglobin adducts of urea and carbamate pesticides. Carcinogenesis 11, 111-115. Sever, L. E., et al. (1997). Reproductive and developmental effects of occupational pesticide exposure: The epidemiologic evidence. Occup. Med. 12, 305-325. Snodgrass, W. R., Kisker, S., and Rozman, K. (1983). Enhanced elimination of an environmental chlorinated hydrocarbon in man: Use of oral mineral oil and cholestyramine. Veterin. Human Toxicol. 25(Suppl. 1),59. Snodgrass, W R., Morgan, D. P., Winsett, 0., and Roy, D. (1986). Mobilization of a halogenated hydrocarbon pesticide from body fat in man: Lindane. Veterin. Human Toxicol. 28,471. Sofer, S., Asher, T., and Shahak, E. (1989). Carbamate and organophosphate poisoning in early childhood. Pediatric Emerg. Care 5, 222-225. Spencer, P. S., Arezzo, 1., and Schaumburg, H. (1985). Chenicals causing disease of neurons and their processes. In "Neurotoxicity of Industrial and Commercial Chemicals" (1. L. O'Donoghue, ed.), Vol. 1, pp. 1-14. CRC Press, Boca Raton, FL. Stampfer, 1. E., et al. (1984). Permeation of eleven protective garmet materials by four organic solvents. Amer. Indust. Hyq. Assoc. J. 45, 642-654. Steenland, K., et al. (1995). Chronic neurological sequelae to organophosphate pesticide poisoning. Amer. J. Public Health 84, 731-736. Stock, 1. K., and Priestly, B. G. (1986). Urinary thioether output as an index of occupational chemical exposure in petroleun retailers. Brit. J. Indust. Med. 43,718. Strik, 1. 1. T. (1987). Porphyrins in urine as an indication of exposure to chlorinated hydrocarbons. Ann. N. Y. Acad. Sci. 514,219. Sunaga, M., et al. (1989). Urinary alkylphosphate levels as an index of exposure to organophosphate insecticides in pest control operators. Nippon Eiseigaku Zajsshi 44,763-770. Tusk, G. M., et al. (1997). Pralidoxime continuous infusion in the treatment of organophosphate poisoning. Ann. Pharmacother. 31,441-444. U.S. Environmental Protection Agency (1992). "Pesticide Industry Sales and Usage. 1990 and 1991 Market Estimates." Office of Pesticide Programs, U.S. EPA, Washington DC. Vesell, E. S. (1987). Pharmacogenetic perspectives on susceptibility to toxic industrial chemicals. Brit. J. Indus. Med. 44, 505-509. Wester, R. c., and Maibach, H. 1. (1985). In vivo percutaneous absorption and decontamination of pesticides in humans. J. Toxicol. Environ. Health 16, 25-37. Wolff, M. S., et al. (1992). Dermal levels of methylparathion, organochlorine pesticides, and acetylcholinesterase among formulators. Bull. Environ. Contam. Toxicol. 48,671-678. World Health Organization (WHO) (1990). "Public Health Impact of Pesticides Used in Agriculture." WHO, Geneva. Young, A. L. (1984). Determination and measurement of human exposure to the dibenzo-p-dioxins. Bull. Environ. Contam. Toxicol. 33,702. Zahm, S. H., et al. (1997). Pesticides and cancer. Occup. Med. 12,269-289. Zweiner, R. 1., and Ginsburg, C. M. (1988). Organophosphate and carbamate poisoning in infants and children. Pediatrics 81,121-126.
CHAPTER
27 Surveillance of Pesticide-Related Illness and Injury in Humans* Geoffrey M. Calvert and Wayne T. Sanderson Centers for Disease Control and Prevention
Margot Bamett Strategic Options Consulting
Jerome M. Blondell V.S. Environmental Protection Agency
Louise N. Mehler California Environmental Protection Agency
27.1 INTRODUCTION A simple concise definition for surveillance is "data for action" (Giesecke, 1999). Surveillance data are vital for targeting public health resources. Traditionally, surveillance includes the ongoing collection, analysis, interpretation, and dissemination of data to prevent and control disease (Thacker and Berkelman, 1988). Surveillance data are useful for identifying the nature and magnitude of public health problems and evaluating the effectiveness of interventions to address those problems. In this context, the term public health surveillance is used. Public health surveillance is directed at national or regional (e.g., state or province) popu1ations. It can be distinguished from "local" surveillance or medical screening programs. Medical screening programs are directed at a more limited population (e.g., workplace or community) and are implemented to enable early recognition of individuals needing treatment, prophylaxis, or additional education and training. The most important use of surveillance data is to guide prevention activities, including regulatory, enforcement, consultative, or educational interventions. Surveillance can produce many data products that are useful for directing preventive action. These data products include the following: (1) estimation of the magnitude of the problem, (2) identification of trends in disease occurrence, (3) identification of epidemics or clusters of *This chapter has been reviewed by the National Institute for Occupational Safety and Health, the V.S. Environmental Protection Agency, and the California Environmental Protection Agency. However, the contents of this chapter do not necessarily reflect the views of these agencies. Handbook of Pesticide Toxicology
Volume 1. Principles
disease, (4) identification of emerging problems or new populations at risk of disease, and (5) evaluation of the effectiveness of prevention and intervention efforts. Through the dissemination of these data, public health surveillance focuses attention on important health problems. The toxicity of pesticides continues to raise public concern and is the focus of much media attention. The importance of pesticides to protect the food supply and to control disease vectors is well recognized. However, it is also recognized that there is no perfectly safe form of pest control. Because society allows pesticides to be disseminated into the environment, society also incurs the obligation to track the health effects of pesticides. As such, surveillance of pesticide-related illness and injury continues to be important. Surveillance for pesticide poisoning identifies primarily two groups of cases, each of which requires different approaches for intervention. The first group consists of cases that are preventable by following the precautionary measures specified on product labels and in government regulation. The appropriate interventions for these cases include enhanced education and enforcement. The second group of cases occurs despite compliance with label instructions and regulatory measures and therefore requires interventions aimed at changing pesticide use practices and/or modifying regulatory measures. This chapter will describe state-based, national, and international surveillance systems for pesticide-related illness and injury. Surveillance systems are the network of individuals and activities that engage in the process of surveillance. There are no comprehensive, national surveillance systems for pesticide-
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from individuals and health care professionals seeking information on how to manage an exposure to a poison. Approximately 13% of their calls come from doctors treating exposed patients. The other 87% come from victims of the exposure or their relatives (e.g., mother of an exposed child). Typically, a PCC is run by a hospital or university. "Poison Centers function primarily to provide poison information, telephone management and consultation, collect pertinent data, and deliver professional and public information" (AAPCC, 1988). Most of the cases (83%) in TESS are submitted by certified PCCs. To be certified, a PCC must fulfill the following criteria (AAPCC, 1988):
Figure 27.1 Greenhouse worker dressed in protective clothing, gloves, and full-face respirator applying insecticide to potted plants. (Courtesy of Wayne T. Sanderson. )
related illness or injury. Therefore, none of the surveillance systems described in this chapter provides a complete understanding of the pesticide-related illness problem. However, each system has strengths and weaknesses and each system provides data that are useful for directing active intervention. The focus of this chapter will be on surveillance systems that operate in the United States (both state based and national); however, some information is provided on international surveillance efforts. The chapter also describes some of the tools of surveillance (e.g., regulations that facilitate surveillance, efforts toward standardization of case definitions and variables, and guidelines for evaluating surveillance systems). In addition, the chapter provides a general discussion of the limitations and strengths of surveillance data, with specific reference to the surveillance of pesticide-related illness and injury. Finally, the chapter provides an exploration of the role played by epidemiologic studies in the surveillance of pesticide-related illness and injury.
27.2 SURVEILLANCE SYSTEMS 27.2.1 TOXIC EXPOSURE SURVEILLANCE SYSTEM
Description Most of the nation's poison control centers (PCCs) participate in a national data collection system, known as the Toxic Exposure Surveillance System (TESS). Previously, this system was called the National Data Collection System. TESS is maintained by the American Association of Poison Control Centers (AAPCq. Of the 75 PCCs in the United States, the number that participate varies from year to year. Between 1993 and 1996, the number of participating PCCs ranged from 64 to 67 (Litovitz et al., 1994,1995,1996,1997). Typically, the PCCs serve a population of 1 to 10 million people, and each receives a minimum of 10,000 calls per year (Felberg et aI., 1996). Poison centers receive telephone calls
1. Have a board-certified physician on call at all times with expertise in medical toxicology. 2. Have poison information specialists available on site at all times to handle all calls. The poison information specialists are required to complete a training program and are certified by the AAPCC. 3. Maintain a comprehensive file of toxicology information sources and have ready access to a major medical library. 4. Maintain written operational guidelines that provide a consistent approach to evaluation and management of toxic exposures. The guidelines must include a provision for follow-up of each case to determine the patient's final disposition or medical outcome. 5. Have an ongoing quality assurance program, including regularly scheduled conferences, case reviews, and audits. 6. Keep records on all cases handled by the center with standard data elements and sufficient narrative to allow for peer review. Records must be handled using a standardized form that is acceptable as a medical record. Standard data elements include the date of the call, age and sex of the victim, location of victim at time of exposure (e.g., home, workplace), substance exposed to, route of exposure, initial symptom assessment, source of treatment (e.g., self-treated, referred to physician, hospitalized), and an evaluation of medical outcome after case follow-up (telephone contact with patient or health care professional). Beginning in 1993, information about specific symptoms experienced by the victim was also collected. 7. Submit all case data to the Toxic Exposure Surveillance System within deadline (2-4 times per year) and meet quality requirements. The AAPCC prepares a computer record that is returned to the local PCC (AAPCC, 1988). Taken together, all these criteria help assure the quality of the data. Case Definition Cases are identified through telephone calls from individuals and health care professionals seeking information on how to manage an exposure to a pesticide. Those with symptoms are categorized into minor, moderate, or major, depending on symptom severity and whether recovery is complete. Definitions used by the poison control centers to categorize medical outcome are summarized as follows (Veltri et aI., 1987):
27.2 Surveillance Systems
60:,
Table 27.1 Selected Severity Measures Used in the Analysis of the TESS Data Severity measure % with symptoms
Tables where used
= Cases reporting pesticide-related symptoms or clinical effects x 100 Cases with follow-up to determine medical outcome
% with life-threatening or fatal outcome
= Cases classified as having a major or fatal outcome x 100 Cases with follow-up to determine medical outcome
% seen in a health care facility =
Cases seen in a health care facility x 100 Exposures reported to poison control centers
% hospitalized =
Cases hospitalized x 100 Cases seen in a health care facility % treated in a critical/intensive-care unit = Cases admitted to a critical/intensive-care unit x 100 Cases seen in a health care facility
No effect. Patient developed no symptoms as a result of the exposure. Minor. Symptoms are minimal with no residual disability (e.g., mild gastrointestinal symptoms, skin irritation, drowsiness). Moderate. Symptoms are more pronounced, prolonged, or more of a systemic nature than minor symptoms with no residual disability. Usually some form of treatment is indicated. Examples include high fever, disorientation, hypotension that rapidly responds to treatment, and isolated brief seizures. Major. Symptoms are life-threatening or result in residual disability or disfigurement. Examples include patients who require intubation plus mechanical ventilation and patients who sustain repeated seizures, cardiovascular instability, or coma.
PCC poison specialists rely on their experience and judgment to determine whether cases have symptoms consistent with the toxicology, dose, and timing of the pesticide exposure. No standardized criteria are used to make this determination. Patients treated at home or any other non-health care site are classified as "managed on site" (AAPCC, 1994). Those seen in a health care facility may be classified as either treated and released or admitted for medical care. "Admitted for medical care" is used when "the patient is observed and/or treated and subsequently admitted as an inpatient primarily to receive medical care rather than psychiatric evaluation." For the purpose of this review, unintentional exposures to single pesticide products are emphasized because they constitute the overwhelming majority of cases and risk mitigation measures can be more easily targeted based on the known use pattern of individual pesticides. As a result, excluded from all analyses were attempted suicides, cases with intentional malicious use (e.g., attempted homicide or child abuse), cases exposing themselves for euphoric or other psychotropic effect, and cases where intent could not be determined. Together, these intentional categories account for 2.4% of all exposures. Also excluded were cases exposed to multiple products (an additional 4.1 % of all exposures) and confirmed nonexposures (0.04% of all exposures).
Tables 27.3. 27.4, 27.6 Table 27.4 Table 27.5 Table 27.5 Table 27.5
Several measures of severity were developed for this review of PCC data (Table 27.1). The first severity measure was the percentage of cases reporting pesticide-related symptoms or clinical effects. The second measure was the percentage of cases that had a major medical outcome (as defined previously). With the first two measures, the denominator was the number of pesticide exposure cases whose medical outcome was determined (PCCs were able to determine the medical outcome for only 49% of reported exposures). A third severity measure was the percentage of all exposure cases that were seen in or referred to a health care facility (HCF). Among those seen in or referred to an HCF, the proportion hospitalized represents the fourth severity measure. Typically, cases are not admitted unless the attending physician feels the case is likely to require clinical observation and/or extensive treatment to prevent further adverse effects. Finally, among those seen in or referred to an HCF, the proportion admitted for critical care or treated in an intensive-care unit (lCU) represents the fifth severity measure. Data Source Most of the nation's poison control centers participate in TESS. Between 1993 and 1996, the number of participating PCCs ranged from 64 to 67 (AAPCC, 1998a; Litovitz et al., 1994, 1995, 1996, 1997). Target PopUlation A large proportion of the U.S. population is served by the PCCs included in TESS. Between 1993 and 1996, an average of 81 % of the popUlation was covered by participating PCCs. However, PCCs do not ascertain all cases within their catchment areas. Therefore, data from TESS are representative of the universe of exposures reported to PCCs and cannot be generalized to the entire universe of all poison exposures. Period of Data Collection times per year.
PCCs submit data to TESS 2-4
Periodicity of Reports Data are available by calendar year. The TESS annual report, which includes information on all types of poison exposures, is published annually in the September issue of the American Journal of Emergency Medicine. Additional TESS data can also be purchased from the AAPCC.
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Calendar year data become available 10 weeks after the conclusion of the year. Findings The current review is based on 424,628 records of pesticide-related exposures reported to PCCs participating in TESS from 1993 through 1996 (Table 27.2). Pesticides accounted for 6% of all exposures reported to PCCs; in contrast, 42% of all exposures were due to pharmaceuticals. Of the 424,628 exposures, 392,209 (92%) occurred in a residential setting. The leading cause of exposures was from insecticides. Among insecticides, the organophosphates were the most common exposure (68,496 exposures). Although attempted suicide cases made up only 2.4% of pesticide-related calls to PCCs (and are not included in the detailed review that follows), they may account for nearly 10% of the cases seen in a health care facility. The leading types of products involved in PCC cases of attempted suicide include anticoagulant rodenticides (20% of the total suicide attempts), pine oil disinfectants (14%), organophosphate insecticides (11 %), pyrethrins/pyrethroids (6%), unknown rodenticides (5%), carbamate insecticides (4%), and phenol disinfectants (3%). A primary determinant of risk to pesticides is the age of the case. Children under age 6 accounted for 91 % of the rodenticide exposures, reflecting the fact that rodenticides are mostly baits used inside homes where young children have access to them. By contrast, only 28% of the herbicide exposures and 32% of the fungicide exposures involved children under age 6. These pesticides are more likely to be used and stored outside or at
Table 27.2 Number of Unintentional Exposures Reported to PCCs, 1993-1996, by Pesticide Class and Age Categorya Pesticide class
Child < 6
Children 6--19
Disinfectants
53,617
6,714
18,889
79,777
Fumigants
172
145
1,794
2,156
Fungicides
1,445
494
2,566
4,569
Herbicides
7,948
2,850
17,507
28,728
Insecticides C
121,545
24,762
96,179
245,596
Rodenticides
57,094
2,095
3,219
62,733
242,609
37,127
140,362
424,628
Total d
Adults
aExcluded were suicides, cases with intentional malicious use, cases where intent could not be determined, cases exposed to multiple products, and confirmed nonexposures. bRow totals include cases with undetermined age. cIncludes insect and moth repellents. dColumn totals include cases involving certain other pesticide classes (e.g., molluscicides) not listed individually.
least where young children have less access. Half of the insecticide cases and two-thirds of the disinfectants (67%) involved exposure to a child younger than 6 years of age. Table 27.3 presents the information on those cases that were followed up and determined to have symptoms related to their exposure. It should be noted that only 49% of cases receive follow-up to determine final medical outcome (follow-up was determined for 51 % of children under 6 years and for 46% of adults and children aged 6 years and older). Exposures that are expected to have no effect or minor effects, in the judgment of the poison specialist, often receive no follow-up and account for 38% of all exposures reported to the PCc. An additional 4% of all exposures are lost to follow-up. The remaining 9% of exposures involve individuals with health effects judged to be unrelated to their exposure. Among cases that received follow-up to determine medical outcome, the poison specialist determined that 42% developed exposure-related symptoms. The percentage of exposures with symptoms varied across the various classes of pesticides, from only 4% for anticoagulant rodenticides up to 68% for fumigants. It should be noted that some cases avoid developing symptoms as a result of intervention measures implemented by the caretaker. The low percentage of symptomatic rodenticide exposures was likely related to the high percentage of exposures that occur among children under age 6. Parents of these children will contact a poison center even before symptoms have developed and the overwhelming majority of exposures are to anticoagulants, which have very low toxicity to humans unless ingested in high quantity or repeatedly over a short period of time. Overall, 13% of the symptomatic cases for all pesticides were moderate, major, or fatal, with percentages ranging from 6% for boric acid insecticides to 21 % for fumigants (Table 27.3). Of the symptomatic cases, ingestion was the route of exposure in 51 % of children under 6 years of age, 22% of children aged 6-19, and 14% of adults. Tables 27.4 and 27.5 summarize the data on severity. Adults and children aged 6-19 were combined because their exposure patterns were similar and because the severity distribution differed little between these two groups. The three chemicals with the highest proportion for a given severity measure are indicated by a superscript of 1-3 (i.e., a 1 was assigned to the chemical with the highest proportion falling into a given severity measure). Among young children, organochlorines ranked highest in four of the five severity measures (Tables 27.4 and 27.5). This is primarily due to lindane hair shampoos intended to kill head lice. Children who accidentally ingest this product can develop life-threatening seizures. Organophosphate insecticides also ranked high (in the top 3) for young children on three of the five severity measures: percentage with life-threatening or fatal outcome, percentage hospitalized, and percentage seen in an intensive-care unit (Tables 27.4 and 27.5). Among young children, organophosphate insecticide exposures are four times more likely to result in a fatal or major outcome compared to all other pesticide exposures combined. Once in a health care facility, children exposed to organophosphates are 3 times more likely to be hospitalized and 5 times more likely to be admit-
27.2 Surveillance Systems
607
Table 27.3 Number of Symptomatic Unintentional Cases and Proportion in Each Severity Category by Pesticide Class, 1993-1996a Number of Pesticide class
symptomatic
Minor
Moderate
Major
Fatal
cases b
(%)
(%)
(%)
(%)
Disinfectants Hypochlorites
7,663
84.7
15.1
0.183
0.000
Phenols
4,134
93.4
6.4
0.145
0.024
Pine oil
8,646
92.7
7.0
0.301
0.023
Otherc
2,446
87.1
12.6
0.204
0.041
22,889
89.5
10.2
0.223
0.017
Fumigants
823
78.8
20.2
0.850
0.122
Fungicides
1,172
85.9
13.8
0.256
0.000
Total
Herbicides
7,573
86.9
12.7
0.370
0.026
Insect repellents
6,887
93.0
6.6
0.378
0.029
Insecticides 573
94.2
5.8
0.000
0.000
Carbamates
Boric acidlborates
4,854
85.0
14.4
0.515
0.041
Organochlorines
2,503
84.2
13.8
1.96
0.080
Organophosphates
16,062
84.7
14.5
0.784
0.025
Pyrethrins/pyrethroids
13,680
84.3
15.3
0.417
0.007
7,069
86.6
12.8
0.552
0.028
44,741
85.0
14.3
0.662
0.024
1,559
88.3
11.2
0.449
0.000
OtherC Total Moth repellents Rodenticides Anticoagulants
865
83.4
15.4
1.27
0.000
OtherC
516
79.4
18.0
2.33
0.194
1,381
81.9
16.4
1.66
0.072
87,100
87.0
12.5
0.507
0.024
Total Total d
aSee footnote a in Table 27.2.
bIncludes cases with undetermined age. COther includes other and unknown for that category. dTotal includes other pesticides (e.g., molluscicides) not listed individually.
ted for critical care than children exposed to all other pesticides combined. Among older children and adults, five groups of pesticides ranked in the top 3 for two severity measures: fumigants, carbamate insecticides, organochlorine insecticides, organophosphate insecticides, and other rodenticides (Tables 27.4 and 27.5). Adults seen for organophosphate exposure also showed a pattern of increased risk similar to children but at a much lower level. For example, the percentage hospitalized was 1.5 times that of all other pesticides combined, and the percentage seen in an intensive-care unit was 1.7 times higher. Not unexpectedly, carbamates, which have many of the same use patterns and also poison through cholinesterase inhibition, show a pattern similar to organophosphates. Those with organochlorine exposures had a high risk for life-threatening outcomes and for hospitalization. This was largely due to accidental ingestion and dermal reactions from lindane-based shampoos. Fumigants and non-anticoagulant rodenticides (these categories include methyl bromide, strychnine, phosphides, and
other highly toxic compounds) had among the highest risks for fatal or life-threatening outcomes. Table 27.6 lists the number of symptomatic cases reported by age group and site of exposure. For children under age 6, 97% of the cases occur at a residence and this figure is very consistent regardless of the class of pesticide. For adults and older children, 80% of the cases occur at a residence, 14% occur at the workplace, and 6% occur at other locations, primarily public areas such as a park or store. Symptomatic cases involving phenol disinfectants, other disinfectants, fumigants, fungicides, herbicides, and other rodenticides had the highest proportion of cases occurring at the workplace, accounting for 22%,37%,64%,30%,20%, and 23% of all exposures, respectively. Tables 27.3-27.5 utilize proportionate hazard to rank pesticides by their potential for causing problems. Another method to examine severity is to use the incident rate, defined as the number of individuals who become ill divided by the number at risk over some time period. A surrogate measure for the population at risk can be estimated by the extent of pes-
608
CHAPTER 27
Surveillance of Pesticide-Related Illness and Injury in Humans
Table 27.4 Among Pesticide-Exposed Individuals Who Were Successfully Followed to Determine Medical Outcome, the Number and Percentage (in Parentheses) with Pesticide-Related Symptoms and with a Major or Fatal Outcome by Age Category for Selected Pesticide Classes, 1993-1996a Children under 6 years
Pesticide class
Adults and children 6-19 years
Number (%)
Number(%)
Number(%)
with
major or fatal
with
Number(%) major or fatal
symptoms
outcome
symptoms
outcome
Disinfectants H ypochlorites
2,528 (39.9)
Phenols
2,493 (42.3 1)
Pine oil Other
5,306
(0.034) (0.080) 1
11,237 18
(0.046) (0.066)
(77.4)
0
793 (0.0)
Fungicides
117
Herbicides
1,033
Insect repellents
3,756
(23.9)
(70.6) 1,035
0
(19.2)
(0.0) (0.046)
(0.216) 8 (0.712 3 ) 3 (0.219) 27
(71.5) 3,087
5
(0.259) 32
(75.5) 6,453
2
(0.261) 5
(78.6) 11,492
(19.8)
(0.230) 13
(66.1) 1,519
21
(0.156) 5
(74.1) 3,296
14
9
(88.1 1) 1,612
2
(30.4)
(35.3) Fumigants
5,065
(0.063)
910 (42.1 2 )
Total
4
(41.4 3 )
(0.055)
(79.3 2 )
267
0
302
(0.299) 23 (0.591)
Insecticides Boric acid/borates
(0.0)
(6.9) Carbamates
1,007 (20.3)
Organochlorines
766 (27.7)
Organophosphates
2,871 (23.2)
Pyrethrins/pyrethroids
3,031
Other
1,788
Total
9,730
(33.2) (20.3)
Moth repellents
7 (0.141) 30 (1.086 1) 47 (0.3792 ) 10 (0.109) 21 (0.238 3 ) 115
0
(44.4) 3,799
(0.0) 20 (0.382)
(72.6) 1,714 (59.1) 13,021
21 (0.7242 ) 82
(71.0) 10,515
(0.447) 48
(79.2 3 ) 5,221
(0.362) 20
(73.5) 34,572
(23.2)
(0.274)
(72.8)
673
6
867
(7.2)
(0.065)
(51.6)
714
(0.281) 191 (0.402) 1 (0.059)
Rodenticides 4
145
(3.1)
(0.018)
(11.8)
Other
188
5
328
Total
902
Anticoagulants
(6.0) (3.5) Totalb
27,495 (22.3)
(0.161)
(48.3)
9
473
(0.035) 158 (0.128)
6 (0.490) 8 (U78 1) 14
(24.8) 58,817 (72.2)
(0.735) 300 (0.368)
aThe three chemicals with the highest proportion for a given severity measure are indicated by a superscript of 1-3 (i.e., a 1 was assigned to the chemical with the highest proportion of exposures that fell into a given severity measure). Excluded were suicides, cases with intentional malicious use, cases where intent could not be determined, cases exposed to multiple products, and confirmed nonexposures. bTotal includes other pesticides (e.g., molluscicides) not listed individually.
27.2 Surveillance Systems
609
Table 27.5 Among All Individuals Exposed to a Particular Pesticide Class. the Number and Percentage (in Parentheses) Who Were Seen in a Health Care Facility (HCF), and of Those Seen in an HCF the Number and Percentage (in Parentheses) Who Were Hospitalized, or Treated in an Intensive-Care Unit (ICU) by Age Category, 1993-1996a Adults and children 6- I 9 years
Children under 6 years
Among those seen in HCF
Among those seen in HCF Number
Number
Number
(%)
(%)
seen in
(%)
treated
seen in
(%)
treated
HCF
hospitalized
inlCU
HCF
hospitalized
inlCU
Number (%)
Pesticide class
Number Number
(%)
Disinfectants Hypochlorites
1,023 (10.2)
Phenols
712 (6.5)
Pine oil
3,974 (13.7)
Other
470 (12.8)
Total
6,179 (11.5)
Fumigants
45 (26.2)
Fungicides
97 (6.7)
Herbicides
678 (8.5)
Insect repellents
1,301 (7.6)
36 (3.52) 27 (3.79) 240 (6.04) 19 (4.04) 322
5
2,065
90
(0.49)
(22.4)
(4.36)
8
902
( 1.12) 71 (1.79)
(21.1) 1,429
92
41
(6.44)
(2.87)
(34.5 2 ) 5,482
89
44 (4.05) 284
(5.21)
(1.44)
(21.4)
(5.18)
6
0
99
(13.3 3 )
(0.00)
998 (51.5 1)
5
2
(5.15)
(2.06)
52 (7.67) 50 (3.84)
810
12
(26.5) 4,772
(1.77) 19 (1.46)
19 (2.11)
(16.0) 1,086
5 (1.06)
58 (6.43)
33 (1.60)
(9.92) 34 (4.20) 329
(23.4)
(6.89)
1,167
40
(14.5)
(3.43)
12 (1.10) 105 (1.92) 36 (3.61) 16 (1.98) 132 (2.77) 16 (1.37)
Insecticides Boric acidlborates
775 (7.6)
Carbamates
1,071 (9.6)
Organochlorines
1,593 (39.6 1)
Organophosphates
3,277 (13.0)
Pyrethrins/pyrethroids
2,339 (13.1)
Other
1,956 (10.1)
Total
11,011 (12.5)
Moth repellents
3,141 (18.9)
29 (3.74) 72 (6.72) 258 (16.2 1 ) 504 (15.42 ) 105 (4.49) 257 (13.1) 1,225 (11.1) 182 (5.79)
5 (0.64) 26
222 (11.1)
9 (4.05)
5 (2.25)
2,420
247
123
(19.4)
(10.2 2 )
(5.08 1)
75
1,529
167
(4.71 2 )
(27.1)
(10.9 1)
(2.43)
218 (6.65 1 ) 29 (1.24)
8,863
898
438
(21.0)
(10.1 3 )
(4.942 )
5,770
308
133
(19.7)
84
3,097
(4.29 3 )
(17.9)
437 (3.97) 43 (1.37)
59 (3.86)
21,901 (20.1) 500 (12.3)
(5.34) 232 (7.49) 1,861 (8.50) 39 (7.80)
(2.30) 92 (2.97) 850 (3.88) 16 (3.20)
Rodenticides Anticoagulants Other
17,177 (34.4 3 )
354
2,642
112
(37.1 2 ) Total
19,819 (34.7)
Totalb
42,401 (17.5)
(2.06) (4.24) 466 (2.35) 2,321 (5.47)
aSee footnote a in Table 27.4.
bTotal includes other pesticides (e.g., molluscicides) not listed individually.
51 (0.30) 23 (0.87) 74 (0.37) 682 (1.61)
1,008 (28.7 3 ) 485 (26.9) 1,493 (28.1) 37,156 (20.9)
73 (7.24) 45 (9.28) 118 (7.90) 2,807 (7.55)
25 (2.48) 22 (4.54 3) 47 (3.15) 1,219 (3.28)
610
CHAPTER 27
Surveillance of Pesticide-Related Illness and Injury in Humans
Table 27.6 Among Pesticide-Exposed Individuals Who Were Successfully Followed to Determine Medical Outcome, the Number of Cases (and Percentage) with PesticideRelated Symptoms by Age Category and Site of Exposure for Selected Pesticide Classes, 1993-1996a Adults and children 6-19
Children under 6 Pesticide class
Residential
Other
Residential
Workplace
Other
Disinfectants Hypochlorites
2,464 (97)
Phenols
2,427 (97)
Pine oil Other
5,220
Fumigants Fungicides Herbicides Insect repellents
(3) 66 (3) 86
4,003 (79) 1,105 (68) 2,955
(98)
(4)
(90)
847
63
720
(7)
(47)
(93) Total
64
10,958
279
8,783
(98)
(2)
(76)
15
3
247
(83)
(17)
(31)
111
6
672
(95)
(5)
(65)
987
46
(96)
(4)
3,594
162
4,728 (73) 2,597
(4)
(84)
259
8
275
(97)
(3)
(91)
986
21
(96)
708 (14) 348 (22) 179 (5) 562 (37) 1,797 (16) 504 (64) 307 (30) 1,295 (20) 62 (2)
354 (7)
159 (10) 162 (5) 237 (16) 912 (8) 42 (5) 56 (5) 430 (7)
428 (14)
Insecticides Boric acidiborates Carbamates Organochlorines Organophosphate Pyrethrins/pyrethroids Other Total Moth repellents
3,220 (85)
21 (7) 456 (12)
(98)
(2)
746
20
(97)
(3)
(88)
2,792
79
10,930
(97)
(3)
(84)
2,949
82
9,014
(97)
(3)
1,725
63
(96)
(4)
9,457
273
(97)
(3)
(84)
(12)
651
22
706
82
(97)
(3)
(81)
(9)
683
31
120
13
(96)
(4)
(83)
(9)
181
7
233
76
(96)
(4)
(71)
(23)
864
38
353
89
(96)
(4)
(75)
(19)
1,517
(86) 4,276 (82) 29,232
117 (7) 1,637 (13) 1,138 (11) 608 (12) 3,977
6 (2) 123 (3) 80 (5) 454 (3) 363 (3) 337 (6) 1,363 (4) 79 (9)
Rodenticides Anticoagulants Other Total Totalb
26,666 (97)
829 (3)
aS ee footnote a in Table 27.2. bTotal includes other pesticides (e.g., molluscicides) not listed individually.
47,352 (80)
8,120 (14)
12 (8) 19 (6) 31 (6) 3,345 (6)
27.2 Surveillance Systems
611
Table 27.7 Ratio of ResidentiaI Symptomatic Cases per Million Pesticide Containers in U.S. Homes by Age Category and Pesticide Class a Adults and children 6+
Children < 6 years Residential Pesticide class
symptomatic casesb
Estimated containers in U.S. homes c
Residential
Estimated containers
symptomatic casesb
in U.S. homes c
(millions)
Ratio
(millions)
Ratio
2,767
145.08
19.1
2,364
145.08
16.3
247
32.98
7.5
1,182
32.98
35.8
2,364
176.45
13.4
7,308
176.45
41.4
216
4.83
44.7
88
4.83
18.3
Disinfectant Fungicide Herbicide Insecticide Rodenticide
aSee footnote a in Table 27.2. bRow totals include cases with undetermined age. "Includes insect and moth repellents. bThe average annual number of symptomatic exposures in the residential setting for the years 1993-1996 is provided. CObtained from survey data collected in 1990 (Whitmore et aI., 1992).
ticide use in residential households. The U.S. Environmental Protection Agency (EPA) survey of home and garden pesticide use provided estimated numbers of containers of pesticides for all households in the United States in 1990 (Whitmore et aI., 1992). Table 27.7 provides the number of symptomatic cases in young children and adults/older children, the estimated number of pesticide containers in U.S. homes and provides the ratio of the number of symptomatic pesticide exposures per million containers in U.S. homes. Unfortunately, the EPA survey of home and garden pesticide use does not contain the detailed pesticide information found in TESS. Table 27.7 shows a different pattern among young children compared to adults and older children. For children under age 6, the ratio is highest for those products typically used or stored close to the floor. The survey of households in 1990 found that about 58% of the disinfectants were stored less than 4 feet off the floor without a child-resistant closure. For fungicides, the figure was 48%; herbicides, 19%; insecticides, 32%; and rodenticides, 46%. Unlike the other pesticide classes, rodenticides are used almost exclusively as poison baits placed on the floor where children later find them. Greater use of child-resistant packaging would likely reduce the number of cases significantly, particularly for disinfectants and other products stored in bottles and not used in baits. For baits, tamper-resistant bait stations can be used to reduce exposures to children. One measure of a pesticide's potential hazard is the frequency of cases due to exposure to residues left after application or use (as opposed to exposures directly related to application, accidental ingestion, or other direct contact). The category "environmental exposure" is used by poison control centers to capture this kind of hazard. An environmental exposure is any passive, nonoccupational exposure that results from contamination of air, water, soil, or other surfaces. Environmental exposures account for 7% of all pesticide exposures reported to PCCs (Table 27.8). Not surprisingly, fumigants and moth repellents, which are intended to con-
taminate the air, had the highest proportion of cases due to environmental exposures. Organophosphates were 2.5 times more likely to be involved in environmental exposures than all other pesticides combined (13% vs. 5%). In addition, a larger proportion of symptomatic cases with serious outcomes (which included moderate, major, or fatal outcomes) were due to environmental exposures compared to minor cases or asymptomatic exposures. Among symptomatic cases with a serious outcome (moderate, major, or fatal) involving organophosphates, nearly one-quarter of them were due to environmental exposures. Discussion TESS is an important source of surveillance data of pesticide-related illness and injury. The system has several strengths. Among these are the large number of cases that are reported. It addition, approximately half of the reported exposures receive successful follow-up. As part of this follow-up, the severity of the medical outcome is determined. The TESS also provides data for most areas of the United States. The TESS also has some limitations. For example, PCCs do not ascertain all cases. The extent of underreporting is not known. For example, many poisoning cases seen in emergency rooms or by private physicians do not result in calls to a PCc. In a study comparing hospital data with PCC data, cases identified in a review of inpatient and outpatient medical records of all Utah acute-care hospitals were matched to records from the PCC serving Utah. It was found that only about one-third of the hospital cases were matched to PCC cases (Veltri et aI., 1987). Another limitation is that not all states are included in TESS. Examination of AAPCC annual reports from 1993 through 1996 found that most or all of seven states were not served by a PCC that reported to TESS (Litovitz et aI., 1994, 1995, 1996, 1997). These states were Arkansas, Illinois, Maine, Mississippi, Oklahoma, South Carolina, and Vermont. Another five states (Iowa, Minnesota, Nevada, North Carolina, and Texas) had PCCs that did not report for one or two of the four years. Of
612
CHAPTER 27
Surveillance of Pesticide-Related Illness and Injury in Humans
Table 27.S Among All Exposure Cases, All Symptomatic Cases, and All Symptomatic Cases with a Serious Outcome, the Proportion That Were Due to Exposure to Environmental Residues of Selected Pesticide Classes, 1993-1996
Pesticide class
Among all
Among all
Among all
exposure
symptomatic
symptomatic
cases, the
cases, the
cases with a
proportion
proportion
serious
due to
due to
outcomeb , the
environmental
environmental
proportion
residues
residues
due to
(%)
(%)
residues
Disinfectants Hypochlorites
5
Phenols
7
9
2
2 2
Pine oil Other
2
2
4
Total
2
3
6
Fumigants
26
24
26
Fungicides
10
13
12
Herbicides
11
12
18 4
Insect repellents Insecticides Boric acidlborates
2
6
6
Carbamates
9
13
18
Organochlorines Organophosphates
5
6
9
13
18
24
Pyrethrins/pyrethroids
11
14
19
Other
13
19
27
Total
11
16
21
6
22
32
Moth repellents Rodenticides Anticoagulants
0.4
2
Other Total Total
C
7
9
12
4
6
11
17
aSee footnote a in Table 27.2. bModerate, major, or fatal outcomes were considered serious. cSee footnote b in Table 27.6.
the 67 pesticide-related deaths from 1993 through 1996 identified in the multiple cause-of-death public use tapes (see Section 27.2.5 for more information on these data), 16 deaths (24%) occurred in one of these 12 states during a year when the state lacked coverage by the AAPCC. Thus, cases of poisoning are underestimated by TESS data. It is also not known whether or how reported cases differ from unreported cases. In the Utah study described previously, the characteristics of unmatched cases were not studied, thereby making it impossible to determine how PCC cases differ from hospital cases that do not result in a call to a PCC (Veltri et aI., 1987). As a result of this selection bias, any study using
PCCs as a source for cases can only be judged as representative of the universe of exposures reported to PCCs and not the entire universe of all poison exposures. PCC data are a simple form of a case series and therefore are not appropriate for complicated statistical analysis. However, given the large proportion of the D.S. population served by PCCs participating in TESS (81 %) and the large number of poison exposures, factors identified within this selected series are likely to be helpful for targeting particular types of exposure situations for risk mitigation. Misclassification may also occur when symptoms are reported over the phone and are not confirmed by a physician or laboratory tests. Although about 13% of calls to PCCs arise from health care professionals, the majority are calls made by the victims or their relatives. The PCC poison specialists must rely on their experience and judgment to determine which cases have symptoms consistent with the toxicology, dose, and timing of the exposure. Unlike other surveillance systems (e.g., some state-based systems participating in the Sentinel Event Notification for Occupational Risk), the AAPCC has not provided standardized criteria by which to make this determination. However, the APPCC requires that when health effects are judged to be unrelated to exposure, all available evidence should support this determination. Although some misclassification can be expected to occur, it is assumed to be nondifferential across pesticides. That is, there is no reason to believe that poison specialists are likely to misclassify one pesticide more or less than another. It should be noted that reports in TESS are labeled "poison exposures." Many of the cases in the database never develop health effects as a result of the exposure. There are several potential explanations for the lack of health effects in these cases. These explanations include the following: Advice provided by the PCC led to prompt treatment and/or decontamination; the exposure agent was relatively nontoxic; the exposure dose was not great enough to produce toxicity; or, the exposure was suspected but actually never occurred [cases in which there is sufficient evidence that exposure never occurred are removed from TESS (Litovitz, 1998)]. Tracking symptomatic as well as asymptomatic exposure cases can be useful. Finally, limitations involving some severity measures used in this review may be present. The severity measures may have higher or lower values because of perceptions concerning certain pesticide classes rather than the actual risks associated with them. For example, both the public and many in the health care profession are likely to perceive ingestion of rat poison as more dangerous than any other class of pesticide. As a result, such cases are more likely to be seen in a health care facility even though the overwhelming majority of cases involve minor exposures (e.g., just a taste) to anticoagulants, which pose relatively little risk. Similarly, decisions about which cases become hospitalized or are transferred to an intensive-care unit may be due to inaccurate perceptions of risk. Differences in the experience and style of care may affect a health care professional's choice of care. In addition, health care-seeking behavior by patients is affected by the availability and extent of health insurance cover-
27.2 Surveillance Systems
age or workers' compensation. Often, the PCC provides advice on appropriate use of health care resources, which can result in a significant savings in health care costs. Validity of the data collected by different poison centers is an important concern of TESS. The AAPCC conducted an audit of 588 randomly selected pesticide charts based on records submitted to the TESS in 1996 (AAPCC, 1998b). A total of 58 of these records were excluded (34 were from a center that was overrepresented in the data set, and 24 were from three centers that had closed since 1996). After these exclusions, requests for 530 cases were sent to the PCCs and 512 records were located and returned to the AAPCC for a response rate of 97%. (Thirteen records could not be located, one center did not send the three requested records, and the wrong record was sent in two cases.) Cases were reviewed to determine how accurately the information coded in TESS matched the information in the original medical record. Five fields important to this analysis were selected for the audit: reason for exposure, route of exposure, management site of case, medical outcome, and accuracy of specific and generic substance category. Results from the audit found that the majority of cases were coded correctly (AAPCC, 1998b). Of those cases that contained errors, the most common error was insufficient follow-up to accurately code the fields. "Reason for exposure" was coded correctly 90% of the time, incorrectly coded in 4.5%, and insufficient information to determine coding in 5%. "Route of exposure" was coded correctly in 96% of cases and incorrectly coded in 3.7% [1.7% incorrect route and 2.0% route(s) omitted]. "Health care facility use and referral" was correctly coded for 93.5%, incorrect in 1.8%, and unable to determine correct coding in 4.7%. "Outcome" was correctly coded in 83%, coded incorrectly in 5%, and unable to determine correct coding in 12% (due to inadequate follow-up or missing information). "Specific" substance was correctly coded 93.3% of the time, incorrectly coded 6.5%, and unable to determine if correct 0.2%. "Generic substance" was coded correctly 98% of the time and incorrectly coded 2% of the time. 27.2.2 STATE-BASED SURVEILLANCE SYSTEMS Description Thirty states in the United States require some form of physician, laboratory, or hospital reporting of pesticiderelated illness (Freund et aI., 1989; Zeitz et aI., 1998). These states are listed in Table 27.9 along with information on the specifics of the reporting rule. Only eight states (Arizona, California, Florida, Louisiana, New York, Oregon, Texas, and Washington) routinely conduct more comprehensive case investigation and surveillance activities. In response to public concern, other states are considering initiating or expanding pesticide poisoning surveillance activities. Until very recently, the existing surveillance systems used a variety of methods for collecting and categorizing data that did not permit the routine pooling and analysis of multistate data. The National Institute for Occupational Safety and
613
Health (NIOSH), with funding assistance from the EPA, has begun to address the issue of standardization of pesticide poisoning surveillance. A collaboration involving experts from federal agencies (NIOSH, EPA, National Center for Environmental Health), nonfederal agencies (Council of State and Territorial Epidemiologists, Association of Occupational and Environmental Clinics), and state health departments or other state designees developed a standardized set of variables for pesticide-related illness and injury surveillance. This standardized set of variables is now in use by the eight states mentioned previously. These collaborating entities also developed a case definition and classification scheme that is described elsewhere in this chapter. The large number of pesticide products on the market and difficulties in obtaining case reports make the pooling of all available data particularly desirable. Having standardized variables and a standardized case definition will facilitate the aggregation of these data. The aggregated data will be useful to regulatory agencies, public health policymakers, researchers, worker education programs, the public, and the medical community. This standardization will also be beneficial to other states when they initiate surveillance systems. This section briefly describes the surveillance systems in Arizona, Florida, New York, Oregon, Texas, and Washington. These states have much in common and are not described separately in detail. The surveillance system in California that is maintained by the California Department of Pesticide Regulation (DPR) is described in a separate section, because it is the largest state-based system, has been in existence longer than the other surveillance systems, and uses a slightly different case definition. The California Department of Health Services (CDHS) recently initiated a surveillance system for occupational pesticide-related illness that is similar to the other surveillance systems described in this section. Because the CDHS system is relatively new, it will not be discussed further. The current design of these surveillance systems involves using multiple sources for case ascertainment (these sources are listed in the following Data Source section) and active case follow-up performed either directly by the surveillance system or by partner state agencies. Several of these state systems originally included a system of sentinel health care professionals who were contacted on a regular basis. This approach was labor intensive and did not yield many cases. Five states with pesticide-related illness and injury surveillance systems are partially funded by NIOSH through the Sentinel Event Notification System for Occupational Risk (SENSOR) program. The SENSOR program promotes statebased surveillance of selected occupational conditions, including occupational pesticide-related illness and injury. Besides tabulating the number of cases, these SENSOR-supported surveillance systems perform in-depth investigations for case confirmation, conduct screening of other workers at a case patient's workplace, and develop interventions aimed at particular industries or hazards.
614
CHAPTER 27
Surveillance of Pesticide-Related Illness and Injury in Humans
Table 27.9 Pesticide-Related Illness Mandated Reporting Requirements and Entities by Statea Entities mandated to report Pesticide reporting
Emergency
requirementb
Physician
Alaska
ANYOCCDZ
x
Arizona
ANY PEST
Arkansas
ANY PEST
x x
California
ANY PEST
Connecticut
ANYOCCDZ
Florida
ANY PEST
Hawaii
ANY TOXIN
x x x x
Iowa
ANY PEST
Kansas
ANYOCCDZ
State
Louisiana
ANY PEST
Maine
ANYOCCDZ
Maryland
ANYOCCDZ
Massachusetts
OCCPEST
Michigan
ANYOCCDZ
Mississippi
ANY TOXIN
Missouri
ANY PEST
New Hampshire
ANYOCCDZ
New Jersey
OCCTOXIN
New Mexico
ANY PEST
New York
ANY PEST
Ohio
OCCPEST
Oregon
ANY PEST
South Carolina
ANY PEST
Texas
OCCPEST
Utah
ANY TOXIN
Virginia
OCCPEST
Hospital
room
Clinic
x
x
x
Laboratory
x
x
x
x
x x
x x x
x
x x x x x x
x x x
Poison
Other health
control
care
center
professional
x x
x
x x
x x
x x x
x
x x
x x
x x
x
x
x
x
x x x x x
x
x
x
x
x
x x x
x
x
x
x x x
x
x
Washington
ANY PEST
x x
West Virginia
ANYOCCDZ
x
Wisconsin
OCCPEST
x
Wyoming
ANY PEST
x
x
x
x
aThis table does not include states with only voluntary pesticide reporting requirements (Idaho, Illinois, North Dakota, South Dakota, Vermont). b ANY PEST, reporting of any pesticide-related illness (whether occupational or nonoccupational) is mandated; OCC PEST, only reporting of occupational pesticide-related illness is mandated (there are no requirements for reporting poisoning from nonoccupational toxic exposures); ANY OCC DZ, reporting of any occupational disease is mandated (there are no specific requirements for occupational pesticide-related illness reporting nor are there requirements for reporting poisoning from nonoccupational toxic exposures); ANY TOXIN, reporting of any poisoning from toxic exposures is mandated (there are no specific requirements for pesticide-related illness reporting); OCC TOXIN, reporting of any poisoning from occupational toxic exposures is mandated (there are no specific requirements for occupational pesticide-related illness reporting nor are there requirements for reporting poisoning from nonoccupational toxic exposures). Sources of data: Freund et al. (1989), GAO (1993), Zeitz et al. (1998), and calls to selected states to clarify inconsistencies.
Case Definition These states began using the National Public Health Surveillance System case definition and classification scheme to evaluate reports starting with 1998 data. This case definition is described in detail elsewhere in this chapter (see Section 27.5). Data Source All of these state-based surveillance systems require reporting of pesticide-related illness and injury cases from physicians (Table 27.9). Other sources of case reports vary
by state and include poison control centers, emergency medical services, medical laboratories, hospital emergency rooms, other health care providers, clinics, migrant legal aid, selected community contacts, and state agencies with jurisdiction over pesticide use (e.g., state agricultural departments, state structural pest control boards). Some states also accept self-reports as a trigger for investigation. States also routinely review other data sources to identify additional potential cases and to evaluate the completeness of reporting. The other data sources in use
27.2 Surveillance Systems include workers' compensation claims, hospital discharge data, and death certificates. Many states accept reports from sources other than those required to report through regulation. Both Oregon and Washington maintain interagency boards that are required to coordinate the investigation of reported adverse impacts from pesticides, review incidents, and develop strategies to prevent exposures. The interagency board in Oregon is called the Pesticide Analytical and Response Center (PARC) and the Washington board is called the Pesticide Incident Reporting and Tracking Review Panel (PIRT). Both interagency boards are composed of representatives from agencies with jurisdiction over pesticides, health, and the environment. In addition, these interagency boards include a state poison control center representative and an appointed generalpublic member. PIRT also includes a practicing toxicologist and representation from the state universities. The state universities serve as consultants to the PARC board but are not members. Target Population These systems strive to capture any pesticide-related acute illness or injury occurring in the state population. The systems capture illness and injuries resulting from both occupational and nonoccupational exposures. Although the emphasis in California and Texas has been on occupationally related cases, both the California and the Texas surveillance systems have current projects aimed at enhancing the reporting of nonoccupational cases. Period of Time of Data Collection The commencement of acute pesticide-related illness and injury surveillance varies by state. Oregon has required health care providers to report pesticide poisoning since 1987, and surveillance data are considered complete beginning with the calendar year 1988. In Texas, acute occupational pesticide poisoning has been reportable since 1986, and surveillance data are considered complete beginning with the calendar year 1987. In Arizona, although the reporting rule went into effect in 1987, surveillance data are considered complete beginning with the calendar year 1992. Likewise, in Florida, the reporting rule went into effect in 1987; however, reporting had been limited until the more comprehensive surveillance system was initiated in 1997. In Florida, surveillance data are considered complete beginning with the calendar year 1998. In New York, acute pesticide poisoning has been a reportable condition since August 1990, and surveillance data are considered complete beginning with the calendar year 1991. Finally, Washington State has had an acute pesticide poisoning reporting requirement since 1989, and surveillance data are considered complete beginning with the calendar year 1991. In each of these states, complete data for a calendar year are generally available 12-24 months after the end of the calendar year. Periodicity of Reports lished annually.
Printed reports are generally pub-
Findings Table 27.10 provides a summary of 5 years of data from the five states with the most comparable systems and
615
similar case classification schemes (Arizona Department of Health Services, 1992, 1993, 1994, 1995, 1996; New York State Department of Health, 1995, 1997; Pesticide Analytical and Response Center (PARC), 1991-1992, 1993, 1994, 1995, 1996; Washington State Department of Health, 1994, 1995, 1996, 1997, 1998; 1. Shannon, personal communication). The data in Table 27.10 reflects the reports that have been investigated and classified as likely to be illnesses or injuries related to pesticide exposure. [The table excludes intentional poisonings (i.e., suicide and attempted suicide]).] All of these states classify between 30 and 55% of the reports received as noncases. Insecticides are clearly the most common source of pesticide-related illness. In Arizona and Texas, herbicides, fungicides, and fumigants are responsible for a lower proportion of the total cases (in Arizona and Texas, these three pesticide classes account for approximately 12% of all pesticide-related illness cases, whereas in the other three states combined they account for 27% of all cases). These differences may be due to variations in climate and crop distribution among the states. Washington reports substantially more cases on an annual basis compared to the other four states. This may represent a true difference; however, additional factors responsible for these higher case counts may include the greater resources available in Washington to conduct both investigations and outreach activities. In addition, in Washington, like California, there is a timely referral of workers' compensation cases to the surveillance system. Occupational exposures represent 25-97% of all cases. Texas (where only occupational pesticide poisoning cases are reportable by law) and Washington are at the upper end of this range. From 1992 to 1996, occupational cases involving agricultural exposures accounted for 13-46% of all cases in those states providing information on the location of exposure. Occupational cases caused by agricultural exposures most frequently involved pesticide mixing, loading, and application. The occupational cases that resulted from nonagricultural exposures represent a broad range of occupational classifications but were predominated by pesticide applicators, office workers, and workers in retail establishments. Often, occupational cases among office workers and retail store workers resulted from "bystander" exposures, whereby these workers did not apply the pesticide(s) but worked in proximity to where pesticides were applied or spilled. Nonoccupational exposures that resulted in illness are most commonly related to structural and surface pesticide treatments in and around homes (i.e., bystander exposures) as well as accidental ingestions. Exposures to children (individuals less than 19 years old) represent 7-26% of all cases. Discussion Among the strengths of these state systems is their reliance on a variety of sources for case ascertainment. For example, the development of close ties with regional poison control centers has served to provide more complete reporting, particularly for illnesses from nonoccupational exposures. In addition, workers' compensation systems can be an important source of occupational cases as documented in California and
Table 27.10 Number of Acute Pesticide-Related Illness Cases by Pesticide Class in Arizona, New York, Oregon, Texas, and Washingtona
Pesticide class
Arizona
New York b
Oregon
Texas
Washington
Imlmlmlml~
Imlmlmlml~
Imlmlmlml~
~lmlmlml~
1992 1993 1994 1995 1996
24 14
25 14
26 18
37
18
9
lO
6
2
2
14
2
Other insecticides
o
2
o
1
3
I1
4
6
10
Combination of
4
7
6
13
3
28
5
12
9
6
4
Insecticides Cholinesterase-inhibiting
62 23
25 16
38 20
39 20
30 15
41 32
31 16
26 19
16
16
44
18
2
5
24
2
3
2
5
2
2
2
13
o
7
o
2
6
10
5
6
10
4
12
7
15
13
4
3
2
9
2
2
11
13
2
2
o
2
42 23
30
38 22
80 35
94 58
178 112
116 70
98 44
128 67
3
16
22
23
11
28
47
7
11
10
5
6
10
10
8
12
2
19
9
37
25
16
6
8
21
37 25
36
48
8
6
20
49 14
1l
6
insecticides Pyrethrinsl Pyrethroids
o
insecticides
Other
o o o o o
Combination of
3
Herbicides Fungicides Fumigants Rodenticides
....
Q\ Q\
4
2
1
o
o
2
4
6
lO
6
0
35
o
o o
0
o
2
o
0
o 10
o
o
0
3
46 1
o
o 17
18
o
6
o o
50
20
8
4
o o o
o
0 84
o
o
72
56
3
5
o
3
4
4
4
2
o
2
1
5 0
o o o
4
o o
o o
10
6
5
5
o
o
4
o
2
1
2
1
3
7
o
8
6
o
8
2
17
88
38
29
25
14
4
6
2
4
6
1
67 106
160
347
210 214 1162 114 (10%)
231
o 3
4
o o o
5
pesticide classa Unknown Annual total 5-year total Number of cases that
1
2
28
33
5
48 38 179 21 (12%)
o
3
32
124
56
93 96 453 119 (26%)
48 42 263 36 (14%)
5
45
24
12 35
7
58 290 20 (7%)
were children (less than 19 years) (%)
Number of cases
35 (23%f
217 (48%)
143 (54%)
277 (96%)
853 (73%)
Data not available
60 (13%)
45 (17%)
97 (33%)
534 (46%)
classified as occupational (%) Number of occupational cases that involved agricultural exposures (%) aIntentional poisonings (i.e, suicides and attempted suicides) were excluded. bData on cases related to pyrethrins, pyrethroids, and fumigants were unavailable for New York. Such cases may be induded in the categories Other insecticides and Other. cFor Arizona, data on occupation are available only for 1993-1996.
27.2 Surveillance Systems Washington. Other states should consider improving access to workers' compensation data as they are an important source of cases and can be used to periodically evaluate the completeness of the surveillance system. Another strength of state-based surveillance systems is their access to personal identifiers. By knowing the identity of a case and the location of the exposure, prompt appropriate followup and intervention can be instituted. For example, in the case of an occupational pesticide-related illness, identification of the responsible workplace can result in an investigation to identify other workers with illness and to precisely target appropriate prevention programs. This is in contrast to national surveillance systems, which provide only anonymous data without personal identifiers. Many state systems have found that maintaining physician (or any health care provider-based) reporting is resource intensive. When some states have attempted to promote physician reporting through outreach activities, they found that case reporting increased but only as long as the outreach activities persisted. For physician reporting to be successful, the health care professional must be able to recognize pesticide-related illness and must comply with reporting requirements in a timely fashion. Considering that pesticide-related illness is relatively rare and that health care professionals may not be trained in its recognition, the expectation that the preceding steps will occur is rather optimistic. These issues are explored in more depth later in this chapter (see Section 27.6). Some states require laboratories to report when test results yield evidence for pesticide-related illness. The cholinesterase test is probably the most common laboratory test for recognizing pesticide-related illness; however, it is only useful for organophosphate and carbamate pesticides. An additionallimitation of cholinesterase reports is that only a minority of them are associated with known organophosphate or carbamate exposures. For example, in New York, where state law mandates that clinical laboratories report abnormally depressed cholinesterase levels, of the 198 laboratory reports of abnormally depressed cholinesterase levels received in 1995-1996, only 62 (31 %) had known pesticide exposures (New York State Department of Health, 1997). The remainder of the cases had either congenitally low levels or illnesses associated with low cholinesterase levels [e.g., liver disease, malnutrition, acute infections, and pernicious anemia (Vorhaus and Kark, 1953)]. Before deciding to adopt laboratory reporting for cholinesterase levels, consideration must be made for the resources needed to follow up on abnormal results. The data provided by these state surveillance systems have proved useful. These data have fostered recognition of new populations at risk from exposure [e.g, workers using flea control products (Ames et aI., 1989)], emerging problems with pesticide products [e.g., sulfotepp (Weldon et aI., 1996)], and the persistence of previously identified problems [e.g., reentry interval violations resulting in illness (CDC, 1999)]. These findings are helpful for instituting needed regulatory change and for targeting education programs. Data from these state surveillance systems are marginally able to measure the magnitude
617
of pesticide-related illness. However, the data collected adds to our understanding of the nature and extent of the problem. The standardization of data collection will allow data from these systems to be aggregated, further enhancing our understanding of acute pesticide-related illness and injury. 27.2.3 CALIFORNIA DEPARTMENT OF PESTICIDE REGULATION Description Since 1971, the state of California has required physicians to report pesticide illnesses. During the mid-1970s, responsibility for collecting and evaluating the data moved from the Department of Public Health (later the Department of Health Services) to the Department of Food and Agriculture. Pesticide safety functions, including surveillance, were placed in the California Environmental Protection Agency's Department of Pesticide Regulation (DPR) in 1991. Physicians are required to report by telephone to the local health officer any disease or condition that they know or have reason to believe to be caused by pesticide exposure. Within 7 days, the local health officer is required to transmit this information to the DPR. The California Pesticide Illness Surveillance Program (PISP), which is maintained by DPR supplements these physician reports with reports of occupational cases forwarded to the California Bureau of Labor Statistics (CBLS) by workers' compensation insurers. Beginning in 1987, the program has made an effort to track cases related to antimicrobiaVdisinfectant products. Practically all recorded antimicrobial cases are identified through the CBLS. All identified cases (including those involving nonagricultural exposures) are referred to the agricultural commissioner in the county where the exposure occurred. The commissioner investigates the exposure circumstances and, where appropriate, takes action to enforce pesticide safety regulations. The investigation involves attempts to locate and interview the affected people and also to interview the case's supervisor and other witnesses to the event. When appropriate, investigators request the affected people to sign releases allowing them to obtain copies of relevant medical records. The state of California also supports several specialized analytic laboratories to which the commissioners can submit samples for chemical analysis, further verifying the nature and extent of exposure. DPR scientists abstract data from the commissioners' reports, including any available analytic data and medical records, and maintain it in a computerized database.
Case Definition Physician reporting is required for any "pesticide poisoning or any disease or condition caused by a pesticide." This requirement of California Health and Safety Code Section 105200 specifically states that such consultations may not be dismissed as "first aid;" doctors must report all pesticide cases. Cases involving suicide and attempted suicide are also included. The DPR recognizes that pesticide products are complex mixtures with various possible mechanisms of action. It is DPR
618
CHAPTER 27
Surveillance of Pesticide-Related Illness and Injury in Humans
policy to consider any adverse health effect that results from pesticide exposure to be a pesticide-related illness or injury. For purposes of overall classification, the primary toxic effects of the active ingredient(s) are not distinguished from incidental effects such as nausea in response to odor. The DPR has established a standardized system for classifying the relationship between pesticide exposure and illness or injury. The classification categories are as follows: Definite. The signs and symptoms exhibited by the affected person are such as would be expected to result from the exposure described. Both medical evidence (e.g., blood cholinesterase levels or allergy testing) and physical evidence (e.g., residue on leaf samples or contaminated clothing) support the conclusion that the illness or injury was the result of the pesticide exposure. Probable. There is a close correspondence between the pattern of exposure and the illness or injury experienced. Medical and/or physical evidence may not be available. For example, although symptoms may be highly suggestive of cholinesterase inhibition, without results of cholinesterase testing, the case would have to be entered as probable rather than definite. Possible. There is some correspondence between the pesticide exposure described and the illness or injury experienced. The information available may be ambiguous. Headaches, nausea, and skin rashes, for example, all can be caused by many different things; and sometimes people are uncertain exactly where they were working when a problem began. Such uncertainty will cause a case to be entered as possible. Unlikely. The exposure may be uncertain; the signs and symptoms reported are not typical of the exposure suspected, but the possibility that the victim is suffering the effects of pesticide exposure cannot be dismissed. Uncertain exposures may involve people far from the application site or people who only handled tightly closed packages or thoroughly cleaned containers. Unrelated. Evidence is available to demonstrate that the illness or injury was caused by factors other than exposure to pesticides. Sometimes, a product that initially was thought to be a pesticide turns out to be something else, such as a fertilizer or cleaner. Other times, the attending physician determines that the problem is infectious, not toxic. Asymptomatic. The subject of the investigation was exposed to one or more pesticides, but suffered no illness or injury. Cholinesterase depression without symptoms falls in this category. Such cases may, however, reflect lapses from good work practice. Indirect. The illness or injury reported appears to have been caused not by pesticide exposure, but by measures prescribed for avoiding pesticide exposure. People who develop heat stress through performing vigorous work in heavy protective clothing fall into this category, as do those who develop allergic reactions to rubber gloves.
Definite and probable cases generally are combined in data analyses conducted by PISP. Similarly, cases classified as unlikely, unrelated, asymptomatic, and indirect often are discussed as a group. The category of possible relationship is the most ambiguous. In practice, the possible category indicates that the people involved are known to have had contact with pesticides shortly before becoming ill or injured, but firm evidence is not available to indicate whether or not pesticide exposure caused their illness or injury. Because the possible category may contain more false positives compared to the definite/probable category, the PISP generally discusses possible cases separately from other categories. However, when determining the magnitude of acute pesticide-related illness and injury, cases classified as either definite, probable, or possible are included in the sum.
Data Source A total of 60-75% of the cases PISP reviews originate from workers' compensation reports reviewed at CBLS. Reporting by physicians provides the majority of the remainder. Additional cases may come to light via news media, by citizen complaints to agricultural commissioners, or by recognition in the course of an investigation. Target Population This program attempts to capture any pesticide-related health problem evaluated by a California physician. Because of reliance on workers' compensation, occupational exposures are more fully reported than nonoccupational exposures. Complaints registered by citizens and not by physicians are investigated and forwarded to PISP at the discretion of the county agricultural commissioner. Period of Time of Data Collection Annual data are made available 18-24 months after the end of the calendar year. Periodicity of Reports Printed reports are published annually. The text of the report and tabulations summarizing the year's data are also available on the Internet. The Internet address for the California Department of Pesticide Regulation is http://www.cdpr.ca.gov. The PISP report can be located among the department's publications. Findings Figure 27.2 provides data on the number of acute pesticide-related illnesses and injuries for the years 19911996. Acute pesticide-related illnesses and injuries include those cases classified as either definite, probable, or possible. The total annual number of acute pesticide-related illnesses ranged from 1332 to 2239 during this period, and an overall downward trend is evident. From 1991 to 1993, antimicrobials/disinfectants were responsible for the largest proportion of cases. Between 1994 and 1996, insecticides became the dominant source of cases. Among insecticides, organophosphates and insecticide combinations were most commonly responsible (Fig. 27.3). A train derailment that caused a metam-sodium spill accounted for the high number of cases involving fumigants in 1991. In that incident, approximately 19,000 gallons of met amsodium spilled into a river, liberating methylisothiocyanate
27.2 Surveillance Systems 2500
800 700
2000
600 500
1500
~
2-
400 1000
300
200
~
500
100 0
0 1991
1992
1993
1994
1995
1996
.ANTIMICROBIALS INSECTICIDES D HERBICIDES/DEFOLlANTS . FUNGICIDES D FUMIGANTS -TOTAL (Right Axis) Indudes definite. probable. and possible cases reported to the California Pesticide Illness
Surveillance Program
Figure 27.2 Number of acute pesticide-related illnesses by pesticide category in California, 1991-1996.
vapor. That incident was associated with illness in 435 individuals. The high number of cases involving insecticide combinations in 1996 was largely due to a single large drift episode that involved exposure to an organophosphate/pyrethroid combination. In that episode, 230 farm workers became ill when the insecticide combination that was aerially sprayed on a neighboring field drifted onto the field where the farmworkers labored. Between 1991 and 1996,34% of all pesticide-related illnesses were associated with agricultural exposures. Since the mid-1970s, county agricultural commissioners have usually investigated 2000-3000 cases each year. Among all cases that were investigated between 1991 and 1996, 18% were classified as definite, 25% were classified as probable, and 23% were classified as possible. The remaining 34% of cases were classified as unlikely, unrelated, asymptomatic, or indirect. From 1991 to 1996, an average of 9.4% of pesticide-related illness was associated with exposure to lingering field pesticide residues [in contrast to illnesses associated with the pesticide application process (including exposures from drift), accidental ingestion, or other direct contact (e.g., exposure to residue from structural applications)]. After 1988, when California imposed
400
300
200 100
o
1991
1992
1993
1994
1995
1996
_ ORGANOPHOSPHATES . CARBAMATES PYRETHRIN/PYRETHROID OTHER INSECTICIDES D INSECTICIDE COMBINATIONS
Figure 27.3 Number of acute insecticide-related illnesses by insecticide category in California, 1991-1996.
619
regulations to lengthen restricted entry intervals for several pesticides, the average annual number of illnesses attributed to field residue exposures dropped 44% (from an annual average of 282 illness cases between 1982 and 1988, to an average of 158 cases between 1989 and 1996). The decrease occurred primarily among complaints of skin irritation (from an annual average of 221 illness cases between 1982 and 1988 to an average of 103 cases between 1989 and 1996). The number of systemic illnesses associated with field residue exposures has changed little between 1982 and 1996, averaging 58 cases per year. Discussion The California PISP provides the most comprehensive source of verified information on adverse health effects of pesticides in the United States. Alone among state surveillance systems, it collects data on all types of pesticide products, including antimicrobials. The PISP database can be searched based on dozens of variables, including pesticide identity, type of formulation, toxicity category, type of health effect, circumstances of exposure, and age and sex of the people affected. The program attempts to capture only those events that result in medical consultation, which provides both a threshold of severity and a preliminary screening by a trained professional. The PISP only rarely includes individuals who did not seek medical attention. The large numbers of cases identified through workers' compensation reports, and not by direct physician reporting, demonstrates that physician reporting is incomplete. This suggests that nonoccupational cases may not be fully reported. The surveillance program has been instrumental, however, in identifying opportunities for mitigation through the regulatory program. 27.2.4 BUREAU OF LABOR STATISTICS (BLS) Description Since the early 1970s, the Bureau of Labor Statistics (BLS) has published annual reports on the number of illnesses and injuries in private industry. Beginning in 1992, the information provided in these reports was enhanced to provide additional information on occupational injuries and illnesses. Among the enhancements was the commencement of reporting occupational pesticide-related illnesses and injuries. However, estimates of total numbers of specific occupational injuries and illnesses are not provided. Instead, data are available only when the condition resulted in the worker being away from work for one or more days. The illness and injury estimates provided in these reports are obtained through an annual survey of employers (Bureau of Labor Statistics (BLS), 1998). The survey collects data that employers are required to maintain under the Occupational Safety and Health Act of 1970. The disease estimates provided by the survey are based on a scientifically selected probability sample, rather than a census of the entire population. The sample is selected to represent all private industry in the United States. Because the data must meet the needs of participating state agencies, an independent sample is selected for each
620
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Surveillance of Pesticide-Related Illness and Injury in Humans
914 (Fig. 27.4). In each of these years, insecticides were responsible for most cases (with 1995 being the exception when fumigants were responsible for most cases).
1000 800 600 400 200
J j1
o• 1992
1993
1994
Total Herbicides
m Rodenticides Figure 27.4 data.
1995
m
Insecticides
~
Fum igants
IIlII
Fungicides
Number of pesticide-related illness cases by pesticide cIasslBLS
state. Employers are stratified by their standard industrial classification (SIC) code and by employment size. Employers are then sampled from these strata. For the classes that contain the largest employers, the allocation procedure places all of the establishments of the frame in the sample; as employment size decreases, smaller and smaller proportions of establishments are included in the sample. The response rate is generally over 90% for sampled establishments. By a weighting procedure, sample units are made to represent all units within a sampling strata. Case Definition Occupational pesticide-related illness and injury cases resulting in days away from work and recorded by employers as required under the Occupational Safety and Health Act of 1970. Data Source An annual survey of employers (see the preceding Description section). Target Population This survey provides an estimate of the number of serious, nonfatal pesticide-related illnesses and injuries in private industry that involved days away from work. Excluded from the survey are self-employed individuals; farms with fewer than 11 employees, which account for approximately 50% of farms (NASS, 1998); employers regulated by other federal safety and health laws (i.e., railroad transportation and coal, metal, and nonmetal mining); and federal, state, and local government agencies.
Discussion The BLS provides data on occupational pesticiderelated illness only. In addition, pesticide-related illness data are available only for cases that result in lost work time (suggesting that only the more severe cases are recorded). Data on pesticide-related illness are available beginning in 1992. Because the number of identified cases is relatively small, and because this is a weighted sample and not a census of the entire population, the estimates have the potential to vary widely from year to year. These limitations may explain the high number of cases in 1995 associated with fumigant exposure. 27.2.5 VITAL STATUS STATISTICS: MULTIPLE CAUSES OF DEATH Description The National Center for Health Statistics (NCHS) of the Centers for Disease Control and Prevention (CDC) releases a public-use vital statistics tape file for each data year. This file contains a data record of all deaths occurring annually in the United States. Each data record contains the underlying cause of death, other mentioned causes of death, and demographic data. The public-use tapes can be purchased from the National Technical Information Service or from the Government Printing Office. Additional information on obtaining these tapes can be obtained on the Internet at hup://www.cdc.gov/nchswww/. Case Definition Any of the following causes of death mentioned on the death certificate (International Classification of Diseases, 9th revision): E863.0, E863.l, E863.2, E863.3, E863.4, E863.5, E863.6, E863.7, E863.8 and E863.9 (the pesticide class that corresponds to each of these codes is provided in Table 27.11). Separate codes are used for suicidal poisonings and poisonings possibly related to suicide. Suicides and possible suicides were excluded from the following analysis. In addition, data on specific pesticide products are not available. Data Source Multiple-cause-of-death public-use tape files have been released for each data year beginning in 1968. Only incomplete data are available for 1972, 1981, and 1982.
Period of Time of Data Collection Approximately 16 months are required to collect, compile, and publish findings following a given calendar year.
Target Population The 50 states, New York City, and the District of Columbia.
Periodicity of Reports Printed reports are published annually. Data are also available on the Internet at http://stats.bls. gov/oshhome.htm.
Period of Time of Data Collection Approximately 18 months are required to collect, compile, and publish findings for a given calendar year.
Findings Between 1992 and 1996, the annual number of pesticide-related illness and injury cases ranged from 504 to
Periodicity of Reports annually.
Public-use data tapes are available
27.2 Surveillance Systems
621
Table 27.11
Counts of Pesticide-Related Deaths by Pesticide Class Using Multiple-Cause-of-Death Data, 1987-1996a ,b Pesticide-related cause of death ("EH code) Insecticides of
1987
1988
1989
1990
1991
2
o
o
o
4
4
o
5
2
3
5
4
4
4
o o
o
o o
o
o
o o
o o
o
3
3
4
5
7
o 2
1992
1993
1994
1995
1996
Total
o
o
11
4
4
5
40
I
o o
o o
o
4
29
2
25
o o
19
organochlorine compounds (E863.0) Insecticides of organophosphorus compounds (E863.l) Carbamates (E863.2) Mixtures of
4
insecticides (E863.3) Other and unspecified insecticides (E863.4) Herbicides (E863.5)
4
5
6
1
2
Fungicides (E863.6)
o
2
o
o
2
1
Rodenticides (E863.7)
2
2
5
4
3
o
8
o
3
o
o
4
12
4
4
7
2
o
4
3
4
3
4
35
21
19
28
15
13
20
22
15
11
19
183
Fumigants (E863.8) Other and
1
1
Unspecified (E863.9) Total
aThe underlying and all mentioned causes of death were coded using the International Classification of Diseases, 9th revision (ICD-9) (World Health Organization, 1977). bSuicides and possible suicides were excluded.
Findings Between 1987 and 1996, there were a total of 183 deaths related to pesticides. During this lO-year period, the annual number of death certificates that mentioned pesticiderelated illness and injury ranged from 11 to 28 (Table 27.11). Organophosphate insecticides were mentioned on 40 death certificates, making this the most common pesticide class associated with pesticide-related deaths (Table 27.11). Although most of the pesticide-related deaths were among adults, 19 (10%) were among children under the age of 10 (17 of whom were under the age of 6). Most of the pesticide fatalities were among whites (71 %) and males (73%). However, blacks accounted for a disproportionate number of cases (26%). There are only limited data available on the circumstances of these pesticide-related deaths. A total of 67 (37%) occurred in the home, 5 (3%) occurred in a public building, 3 (2%) occurred in industry, 9 (5%) were noted to have occurred in an "other" location, and one (1 %) occurred on a farm. Data on the location of the poisoning were not available for 98 (54%) of the deaths. Discussion The multiple-cause-of-death data are a useful source of data on pesticide-related deaths. However, only the most severe poisonings are included in this data source. There has been little change in the number of pesticiderelated deaths over the past 10 years. However, the numbers are lower than those reported in the 1970s and earlier (Hayes and Vaughn, 1977). Pesticide-related deaths numbered 97 in 1961 and 33 in 1974 (Hayes and Vaughn, 1977).
These data also demonstrate declines in the number of pesticide-related deaths among children under lO-years of age. Between 1987 and 1996, these children accounted for 10% of the pesticide-related deaths. This is low when considering that children under the age of lO comprised 15% ofthe U.S. population (U.S. Bureau ofthe Census, 1995). (It should be noted that children under the age of 6 accounted for 9% of all pesticiderelated deaths and also comprised 9% of the U.S. population.) Furthennore, children less than 10 years of age accounted for a lower proportion of all pesticide-related deaths during this 10-year period compared to the more distant past. In 1974, the proportion was 32%, whereas, in the more distant past, it was over 50% (Hayes and Vaughn, 1977). Nonetheless, any childhood poisoning fatality is a tragedy and highlights the need for ongoing efforts to prevent pesticide access among children. As has been documented in the past, blacks account for a disproportionately high number of cases. In 1990, 12% of the U.S. population was black (U.S. Bureau of the Census, 1995), which contrasts with the fact that blacks accounted for 26% of the pesticide-related deaths between 1987 and 1996. The limitations of this data source are many. First, only the most severe cases are included. Second, it is likely that not all pesticide-related deaths were included, because some may have been coded to other nonspecific causes of death. For example, Hayes and Vaughn (1977) found that in 1973-1974 only 63% of accidental pesticide-related deaths were coded with the correct "E" code. Finally, details on the circumstances of exposure are not available on the multiple-cause-of-death data file. Col-
622
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lection of such details would require direct queries to the health care provider, as has been done in the past (Hayes, 1976; Hayes and Vaughn, 1977). However, information on each decedent in the multiple-cause-of-death data file includes age, race (including Hispanic origin), gender, state of residence, marital status, and place of the accident. 27.2.6 NATIONAL HOSPITAL DISCHARGE STUDIES: COLORADO STATE UNIVERSITY Description Three previous studies were conducted by Colorado State University to estimate the nationwide incidence rates for hospitalized acute pesticide poisoning. These studies covered the intervals 1971-1973, 1974-1976, and 1977-1982. No similar national studies have been conducted using data after 1982. We provide information on the most recent study (Keefe et aI., 1990). The nationwide poisoning estimates provided by this study are based on a stratified random-sampling procedure involving all general-care hospitals. States were placed in three strata based on the state's average rate of hospitalized pesticide poisonings for the years 1971-1976. Hospitals were sampled from each state; however, a higher proportion of hospitals were sampled from those states in the stratum with the highest hospitalized pesticide poisoning rates. Approximately 6% (368 hospitals) of all U.S. general-care hospitals were included in the study. Case Definition
Hospitalized pesticide poisoning case.
Data Source A survey of general-care hospitals (see the preceding Description section). For the sampled hospitals, all medical records were reviewed. For patients with designated diagnoses from the International Classification of Diseases, 8th revision, Adaptedfor Use in the United States (ICDA), medical records were reviewed and appropriate data were abstracted. Target Population This survey provides an estimate of the nationwide incidence of hospitalized pesticide poisoning cases in the United States. Period of Time of Data Collection The last survey covered the time period 1977-1982. Periodicity of Reports Printed reports are available for each of the later two studies (Keefe et aI., 1990; Savage et aI., 1980), and a review of the findings from the first two studies is also available (Keefe et aI., 1985). Findings Between 1977 and 1982 in the United States, the average annual number of hospitalized unintentional pesticide poisoning cases was estimated at 2380 (range: 2127-2991). The estimated annual number of occupational cases averaged 814 (range: 513-1077). The occupations with the greatest
number of pesticide poisonings were farmers and commercial applicators. Organophosphates were most often involved in occupational pesticide poisoning cases, accounting for approximately 43% of occupational cases. Following are the top 10 pesticides, listed in order (highest to lowest), responsible for the most number of occupational pesticide poisonings that required hospitalization: parathion, malathion, methomyl, carbofuran, 2A-dichlorophenoxyacetic acid (2A-D), mevinphos, methyl parathion, disulfoton, aldicarb, and glyphosate (Blondell, 1997). Children 0-4 years of age accounted for 57% of all unintentional nonoccupational hospitalized pesticide poisoning cases. Discussion These hospital discharge data are a useful source of data on severe pesticide poisoning cases. Unfortunately, the most recently available data are from the period 1977-1982. Funding for this study was cut by the EPA in the early 1980s due to agency budget cuts and redirected priorities. It is interesting that the number of hospitalized pesticide poisoning cases estimated using these data are similar to the average annual number of such cases identified by TESS (see Table 27.5). As there are data to suggest that TESS underestimates the number of pesticide poisonings cases seen in hospitals (Veltri et aI., 1987), these data from Colorado State University may similarly provide underestimates of the true incidence. Furthermore, because the data are obtained from a weighted sample and not from a census of all hospitals, one is led to suspect that some imprecision exists in the estimates. 27.2.7 NATIONAL HOSPITAL DISCHARGE SURVEY: NATIONAL CENTER FOR HEALTH STATISTICS Description The National Center for Health Statistics (NCHS) conducts an annual survey of nonfederal, short-stay hospitals in the United States. The survey has been conducted annually since 1965. Data are collected from a sample of inpatient records acquired from a national sample of hospitals. The data represent a sample of discharges and not patients. Therefore, persons with multiple discharges can be sampled more than once. Only general hospitals, children's general hospitals, or hospitals with an average length of stay of less than 30 days are included. Federal, military, and Veterans Administration hospitals are excluded, as are hospital units of institutions (e.g., prison hospitals), and hospitals with less than six beds. The poisoning estimates provided by the study are based on a stratified three-stage design. For 1996, data were collected for 282,000 patient discharges from the 480 responding hospitals. Items available in this data set include age, sex, race, ethnicity, marital status, geographic region of the hospital, discharge diagnoses (up to 7) coded into International Classification of Diseases, 9th revision (ICD-9) (World Health Organization, 1977) categories, and expected source of payment. Occupation of the patient is not available.
27.2 Surveillance Systems
Case Definition Hospitalized accidental pesticide poisoning case. Any of the following diagnoses were eligible for inclusion (International Classification of Diseases, 9th revision): E863.0, E863.1, E863.2, E863.3, E863.4, E863.5, E863.6, E863.7, E863.8, and E863.9 (the pesticide class that corresponds to each of these codes is provided in Table 27.11). Separate codes are used for suicidal poisonings and poisonings possibly related to suicide. Suicides and possible suicides were excluded from the following analysis.
Data Source A survey of general-care hospitals (see the preceding Description section). Target Population This survey attempts to provide an estimate of the nationwide incidence of hospitalized pesticide poisoning cases in the United States. Period of Time of Data Collection Annual surveys have been conducted since 1965. Periodicity of Reports Approximately 18 months are required to collect, compile, and publish findings for a given calendar year. Public-use data tapes are published annually. The E-code data are considered too unreliable to be included in the printed annual report. Findings In 1996, there were an estimated 936 discharges with an E863 discharge diagnosis (accidental poisoning by agricultural and horticultural chemical and pharmaceutical preparations other than plant foods and fertilizers). Of these, 339 involved organochlorines (E863.0), 95 involved organophosphates (E863.1), 129 involved other insecticides (E863.4), and 373 involved rodenticides (E863.7) NCHS, (unpublished data). Discussion The National Hospital Discharge Survey is not a reliable source of data for acute pesticide-related illness and injury. Only 50-60% of the sampled hospitals provide data on E codes. For this reason, E-code data are not published in the annual reports of the National Hospital Discharge Survey. The findings we provide are not consistent with other sources of pesticide poisoning data. For example, organochlorines accounted for 36% of hospital discharges involving pesticide poisonings. In contrast, the National Hospital Discharge Study conducted by Colorado State University for 1977-1982 found that chlorinated hydrocarbon pesticides accounted for 9% of hospitalized occupational pesticide poisonings. In addition, the TESS data for 1993-1996 found that organochlorine insecticides accounted for 8% of hospitalized pesticide poisonings (see Table 27.5). These findings, together with the trend in declining use of the organochlorines, provide additional support for the unreliability of the National Hospital Discharge Survey data for surveillance of acute pesticide-related illness and injury.
623
27.2.8 SOUTH CAROLINA HOSPITAL DISCHARGE SURVEYS Description The Medical University of South Carolina has periodically conducted a survey of hospitalized pesticiderelated illness and injury in South Carolina. The initial survey was published in 1975 and included data from the period 19711973 (Caldwell and Watson, 1975). In the most recent survey (which included the years 1992-1996), all primary-care hospitals in South Carolina were invited to participate. Sixty-two hospitals (90%) participated. Case Definition Any inpatient medical record that contains one of the following ICD-9 codes: 989.2 (chlorinated pesticides), 989.3 (organophosphate and carbamate pesticides), and 989.4 (other pesticides). Intentional poisonings (e.g., suicides and suicide attempts) were included.
Data Source A survey of all primary-care hospitals in South Carolina that involved a detailed review of case charts by a board-certified physician (see the preceding Description section). Target Population This survey provides an estimate of the incidence of hospitalized pesticide poisoning cases in South Carolina. Period of Time of Data Collection Periodic surveys have been conducted since the early 1970s. Periodicity of Reports Reports and published papers provide the findings of the periodic surveys. Findings The most recent survey found that there were 112 hospitalized pesticide-related cases during the years 1992-1996 (with a mean of 22 cases/year) (Caldwell et aI., 1997). Intentional ingestions accounted for 40% of the cases. Unintentional nonoccupational poisonings accounted for 57% of the cases (this category included 26 adults and 31 children). Only 9 (8%) cases had an occupational etiology, 8 of which occurred in an agricultural setting. Discussion Using hospital inpatient data, a decline in pesticide-related illness and injury has been observed in South Carolina. The average annual number of cases has declined from a peak of 79 cases/year during 1979-1982 to 22 cases/year during 1992-1996. The decline in occupationally related cases is thought to have played the largest role in this decline (from a mean of 20 cases/year during 1979-1982 to a mean 2 cases/year during 1992-1996). The authors of the most recent survey attribute this to the success of pesticide applicator training programs, the licensing and certification of applicators using restricted pesticides, and the increasing use of pesticides with lower toxicity (i.e., pyrethrinlpyrethroid insecticides). Because this survey includes data only on hospitalized cases, only the most severe pesticide-related illnesses are captured.
624
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Surveillance of Pesticide-Related Illness and Injury in Humans
The most recent survey included information on outpatient emergency room visits from the 56% of participating hospitals that had this information in computerized format. As more hospitals move to electronic data storage, the outpatient data will become an important source of surveillance information. 27.2.9 NATIONAL AGRICULTURAL WORKERS SURVEY
Description Since the early 1950s, the U.S. government has attempted to monitor the size, composition, and needs of the agriculturallabor force. This function was transferred from the Department of Agriculture to the Department of Labor in 1987, and the National Agricultural Workers Survey (NAWS) reached its present form in 1989. The survey interviews about 1500 agricultural workers in each of three annual interview cycles. Cycles begin in February, June, and October and last 15-16 weeks. Interviewees are selected by a multi stage stratified process. The program defines 12 regions, each of which is further divided into several farm labor areas. Each farm labor area is an aggregate of counties that have similar agricultural and economic characteristics. At least three farm labor areas are selected from each region. One county is then selected from each selected farm labor area. Staff then compile lists of all farms within the selected counties and solicit cooperation from a random sample of the farms. At participating farms, the employees are sampled with probability proportional to the square root of the size of the farm workforce. By sampling and recruiting workers at their worksite, this survey minimizes the undercounting of this population. After they are identified and recruited, the farm workers are interviewed outside of working hours and are offered a $10.00 honorarium for responding to a personal interview that lasts about an hour. The NAWS cooperates with other government agencies by including questions that help measure the needs that can be addressed by migrant education, migrant health, and Census Bureau programs, among others. Since 1993, the EPA has contributed a limited set of questions designed to elicit information about health effects associated with pesticide exposure. Beginning with the October 1998 cycle, the health effects questions were expanded and improved. The survey now includes questions on medical history, use of medical services, participation in pesticide training, and housing conditions. A complete occupational history for the year preceding the interview is also obtained. Although the survey includes questions on general pesticide exposure (e.g., "In the last 12 months, did you load, mix, or apply pesticides?"), information on exposure to specific pesticide products or pesticide classes is not obtained. Periodic reports presenting aggregate findings are available from the Office of the Assistant Secretary for Policy of the U.S. Department of Labor. A public-use tape containing data on approximately 13,000 farmworkers collected between 1993 and 1998 is available. Finally, if the desired data are not available in published reports or in the publicuse tape, specialized analyses can be requested from NAWS
staff. Additional information on NAWS is available on the Internet at hUp://www.dol.gov/dol/asp/public/programs/agworker/ naws.htm. 27.2.10 INTERNATIONAL SURVEILLANCE EFFORTS
Estimates of the worldwide toll of pesticide toxicity are based on mathematical models. The summary by Levine and Doull (1992) provides an overview of the approaches taken and the available observational inputs. The observational inputs generally consist of mortality records and estimates of the proportion of pesticide-related illnesses that are fatal. Other worldwide estimates use hospitalization data along with mortality data. Finally, some estimates are based on projections of the number of workers exposed to pesticides and the proportion of these workers who experience symptoms of pesticide-related illness. The World Health Organization (WHO) estimates that there are up to 5 million acute unintentional pesticide-related illnesses and injuries per year and that annually there are 20,000 deaths related to unintentional pesticide poisoning (Levine and Doull, 1992). Many surveys and case series concerning the health effects of pesticides are published in local or regional journals not readily accessible in other parts of the world. Others summarize the experience of referral or resource centers. Examples of this approach include a review of pesticide fatalities in Spain (Garcia-Repetto et aI., 1998) and a review of calls to the Greek Poison Center (Vlachos et aI., 1982). The following discussion summarizes primarily reports in internationally available publications that describe systematic, ongoing efforts to record health effects attributable to pesticide exposure. Some surveys are referenced where they provide insight into conditions in areas not covered by ongoing surveillance systems. Among European nations, the United Kingdom (UK), Romania, and Russia attempt formal ongoing surveillance. Additionally, a review of hospital records for Finland (Lamminpaa and Riihimaki, 1992) found pesticides to be a rare cause of hospitalization. Persson et al. (1997) collected reports from the Swedish Poisons Information Center, as well as reviewing hospital and death records. They concluded similarly that "the incidence of acute pesticide poisonings in Sweden is low" (Tables 27.12 and 27.13). The UK surveillance system for pesticide poisoning is implemented primarily through a Pesticide Incidents Appraisal Panel (Health and Safety Executive, 1997). Episodes of pesticide exposure are reported to the panel by the Health and Safety Executive and by local authorities (Table 27.12). The panel includes medical doctors, toxicologists, and poison information specialists. It classifies cases on a consensus basis as confirmed, likely, or unrelated. Because not all investigations obtain the information needed to assign cases to these categories, some cases receive open assessments or are left pending or designated as providing insufficient information. Since 1994,
27.2 Surveillance Systems
625
Table 27.12 Annual Fatalities Ascribed to Pesticide Intoxication by Country
Average annual number of deaths (SD)
Percent self-inflicted
Approximate population (millions)
Country (source of data)
Time period
Costa Rica
1980-1986
40.4
84
2
1988-1993
82 (24.5)
56
<10
1975-1983
1032 (200)
66
1969-1994
0.84
Annual rate of all pesticide fatalities! million 20.2
Annual rate of unintentional pesticide fatalities! milIiona 3.2
(Wesseling et aI., 1993)b Romania
>8
>3.5
14.5
71.2
24.2
85
8
0.1
0.02
0.4
0.01
(Fabritius and Balasescu, I 996)C Sri Lanka (de Alwis and Salgado, 1988)d Sweden (Persson et aI., 1997) United Kingdom
1990-1991
22
66
50
1987-1996
18
0
249
(Thompson et aI., 1995) United States
0.07
(see Section 27.2.5)e SD, standard deviation; - , data not available. aExcludes self-inflicted poisoning deaths. bData presented are crude figures derived from official sources and thought more likely to be comparable to the other entries in this table. A chart review conducted by Wesseling et al. identified additional cases (revised annual number of deaths!year, 61.3) and called into question the percentage of deaths that were self-inflicted (revised estimate, 62%). CData are presented for 28 counties, nonoccupational intoxications only. The covered area comprises less than half the population of Romania, but has higher rates of toxicity than the rest of the country. dpercentage self-inflicted based on review of a sample. eThese data contain only unintentional cases.
Table 27.13 Annual Hospitalizations Ascribed to Pesticide Intoxication
Country (source of data)
Time period
Costa Rica
1980-1986
Average annual number of admissions (SD) 476
Percent self-inflicted
Approximate population (millions)
Annual rate of all pesticide poisoning hospitalizations! million
Annual rate of unintentional pesticide poisoning hospitalizations! milliona
24
2
238
181
75
14.5
944
236
31
4.7
16.6
8
7
(Wesseling et aI., 1993) Sri Lanka
1975-1983
13,688 (1,921)
(de Alwis and Salgado, 1988) Finland (Lamminpaa and Riihimaki, 1992) Sweden (Persson et aI., 1997) United States
1980,1981,
78
11.5
1982,1987,1988 1978-1983
50-60
1988-1993 1977-1982
(Keefe et aI., 1990) SD, standard deviation; - , data not available. "Excludes hospitalizations due to self-inflicted poisoning.
2,834
16
227
11.4
9.6
626
CHAPTER 27
Surveillance of Pesticide-Related Illness and Injury in Humans
the panel has also assessed the character of health effects, classifying them as acute or chronic; local or systemic; and mild, moderate (requiring medical intervention), or severe (requiring inpatient therapy). In most years, this system records fewer than 100 cases and classifies fewer than half of them as confirmed or likely. Several sources have voiced concern that reporting may be seriously incomplete and recommended efforts to improve notification of the panel about pesticide poisoning cases. The purpose of the UK surveillance system is to identify "unanticipated adverse effects of new pesticides or previously unrecognized effects of established products" and to help "inform decisions about appropriate control measures, research needs and resource allocation." A recent review of the surveillance system concluded that the data collected from accident and emergency departments, data from inpatient stays, and death data could provide an indication of the overall magnitude of the problem, but did not supply enough specific information to fulfill the purpose of identifying the effects of particular pesticides (G. Jones, Health and Safety Executive, personal communication). Records of the National Poisons Information Service sometimes include information to assess the effects of particular pesticides. However, only those cases investigated by the Pesticide Incidents Appraisal Panel regularly recorded the critical features of pesticide-related illness and injury episodes. Fabritius and Balasescu (1996) describe the situation in Romania. There, sanitary police and preventive health care units file reporting cards for pesticide intoxications, among other conditions. A review of these reports for 26 counties identified hundreds of deaths annually. In Romania as in the United Kingdom, unexplained variations in the data suggested incomplete reporting. In Russia, the Federal Ministry of Public Health's Center on Sanitary Epidemiological Surveillance cooperated with the Computer Center of the Russian Academy of Sciences to develop an enhanced version of a database on pesticides (Cilovtsev et aI., 1998). The updated software, known as PESTOTEST, supports decision making at federal, regional, and local levels. Estimates of pesticide morbidity and mortality in eastern Asia are sparse. They rely heavily on a review of hospital records for the year 1979 in Sri Lanka (Jeyaratnam et aI., 1982) and an interview survey of agricultural workers in Indonesia, Malaysia, Sri Lanka, and Thailand (Jeyaratnam et aI., 1987). In addition, Sri Lanka's judicial medical officer and his deputy reviewed deaths and hospital admissions for the years 1975 through 1983 (de Alwis and Salgado, 1988). During that period, pesticide poisoning accounted for 4% of all autopsy cases (over 1000 deaths per year) and for more than 10,000 hospital admissions each year (Tables 27.12 and 27.13). Chan et al. (1996) mention surveillance for vegetable-borne pesticide poisoning in Hong Kong in the context of a review of inquiries to the Drug and Poisons Information Bureau and of medical records at a major Hong Kong hospital. Hong Kong has been subject to repeated outbreaks of methamidophos poisoning from contaminated vegetables. In 1992 alone, 47 outbreaks of methamidophos toxicity were recorded, presumably affect-
ing hundreds of individuals. Otherwise, few poisoning episodes were identified in Hong Kong. In Africa, surveillance data are available for only two countries (South Africa and Kenya). In South Africa, the Health Act makes pesticide poisoning a notifiable condition. The Department of National Health and Population Development receives any notifications. London et al. (1994) analyzed the reports of pesticide poisoning collected from 1987 to 1991 by one regional office. They also reviewed hospital records for the same period and found that the majority of hospitalizations had not been reported. A small-scale attempt to monitor the health of pesticide users and formulators began in Kenya during 1997 (Kimani and MacDermott, 1998). Amos (1998) reported that a survey of cocoa farmers in Nigeria found some physical complaints to be strongly correlated with pesticide use. In Central America, the surveillance system in Nicaragua serves as a model for other countries in this region. Nicaragua included pesticide poisoning among notifiable conditions in 1979, when the National Unified Health System was established (Cole et aI., 1988). Initially, few cases were reported. By 1987, the system was capable of identifying an epidemic of pesticide poisoning that resulted from increased pesticide usage, which was encouraged by loan subsidies and a favorable exchange rate coupled with deregulation of the price of crops (McConnell and Hruska, 1993). An evaluation by Keifer et al. (1996) still detected substantial underreporting, however. Reviews of medical records in Costa Rica (Leveridge, 1998; Wesseling et aI., 1993) demonstrated that pesticide poisoning was a substantial public health problem (Tables 27.12 and 27.13). Costa Rica declared mandatory reporting of pesticide intoxication in 1983 and subsequently undertook a pilot surveillance project, with support from the Swedish Agency for International Development, that focused on a single county. The Pan American Health Organization and WHO later joined this effort. Surveillance has since extended to the Atlantic region, with additional areas scheduled for inclusion (Castro et aI., 1998). The Costa Rican system is organized hierarchically
Figure 27.5 Team of workers using backpack sprayers to apply herbicide to sugarcane field in South Africa. (Courtesy of Wayne T. Sanderson.)
27.3 V.S. Environmental Protection Agency Regulations
(Antich, 1998). Each health center transmits case records to its immediate hierarchical superior. These collection points forward their data to their superiors, until all reports arrive at the national office. Software identifies possible duplicate entries and flags cases for investigation based on criteria of severity (such as fatality) or social importance (such as pediatric poisoning). In 1997, again with support from the Pan American Health Organization and WHO, Mexico established surveillance for pesticide illness in the states of Sonora and Sinaloa (Alvarez et aI., 1998). This effort used a broad case definition, intending to record "all persons of any age or sex suspected to have acute or chronic adverse health effects related to pesticide exposure." After a period of outreach to physicians, nurses, agricultural specialists, and the community, the system collected case reports passively from health centers, clinics, and hospitals and performed active surveillance in some high-risk situations. Staff also reviewed existing mortality data and medical or hospital registries to identify cases. From March through December of 1997, Sonora reported 42 cases and Sinaloa reported 44 cases of pesticide-related illness. In South America, pesticide-related illness and injury surveillance is conducted in Brazil and Uruguay. In 1982, surveillance efforts began in the Campinas region of the Brazilian state of Sao Paulo (Trape and Zambrone, 1991). A report on improvement efforts (de Oliveira, 1998) implied continued existence of a Brazilian program. Absence of reports from one region led to design of a sentinel event system, with preparation of instruments for follow-up of index cases reported by specially trained health professionals. A single toxicology information center has served all of Uruguay since 1975 (Burger, 1998). The records kept there indicate that during the earliest years pesticides were responsible for one-quarter of the center's consultations. Pesticides still contributed 15% of the case load in 1997.
27.3 U.S. ENVIRONMENTAL PROTECTION AGENCY REGULATIONS The EPA is responsible for implementing several regulations that promote the safe use of pesticides and that facilitate surveillance of pesticide-related injury and illness. These regulations are discussed in the context of collecting information about the acute adverse effects of pesticides and the regulatory programs available to implement risk mitigation. 27.3.1 THE FEDERAL INSECTICIDE, FUNGICIDE, AND RODENTICIDE ACT
The EPA regulates the use of pesticides in the United States under the authority of the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA). No pesticide may legally be sold or used in the United States unless it bears an EPA registration number. It is a violation of the law for any person to use a pesticide in a manner inconsistent with its label. FIFRA gives the
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EPA the authority and responsibility for registering pesticides for specified uses, provided that such uses do not pose an unreasonable risk to human health or to the environment. FIFRA provides a number of remedies that can reduce and mitigate risks from pesticides. If subsequent information indicates that the use of a pesticide would pose unreasonable risks, the EPA has the authority to suspend or cancel its registration. If a pesticide warrants special handling because of its toxicity, it may be classified for restricted use. Pesticides with a restricteduse classification can be applied only by a certified applicator or under a certified applicator's direct supervision. States administer the certification programs that require the certified applicator to demonstrate competency with respect to the use and handling of pesticides. States also have the responsibility for enforcement of FIFRA (e.g., investigating and issuing penalties for a label violation). The pesticide label provides directions on how, when, and where a pesticide can legally be used and which pests can be controlled. The EPA classifies all pesticides into one of four acute toxicity categories based on established criteria (40 CFR Part 156). Toxicity is determined by acute animal tests for oral, dermal, and inhalation lethality and corrosive effects to the skin and eyes. Those pesticides with the greatest toxicity are placed in Toxicity Category I. Other pesticides are placed in the remaining three toxicity categories (Toxicity Categories 11, Ill, and IV). A hazard signal word indicates the toxicity category of a pesticide product. The most hazardous pesticides (i.e., Toxicity Category I) are labeled "Danger," those with moderate toxicity (i.e., Toxicity Category 11) are labeled "Warning," and less toxic pesticides (i.e., Toxicity Categories III and IV) are labeled "Caution." Precautionary statements describe the protective clothing and other equipment that must be used. They also specify the hazards to humans, children, domestic animals, and the environment. A statement of practical treatment may advise on the signs and symptoms of poisoning, provide information on first aid and antidotes, and provide a note to physicians on appropriate treatment. The label specifies directions for safe storage and disposal. The label may have a number of statements designed to reduce risk in addition to the requirements listed previously. Label statements may limit the amount used by specifying the frequency of application, amounts handled, acreage treated, or the rate of application. Application may require enclosed cabs, closed mixing/loading systems for liquids, ventilation, or mechanical flagging devices. Certain more hazardous application methods (e.g., air blast spraying) may be prohibited. Hygiene statements may require washing or not wearing contaminated clothing the next day. After application, a restricted-entry interval may be imposed when unprotected persons are not permitted in the treated area. Posting and/or notification may be required to warn bystanders and others not directly involved in the application. Formulation and packaging requirements can also be imposed under FIFRA to reduce the risk from pesticides. For example, the container size or percentage active ingredient may be limited. Formulation may require ready-to-use solutions in-
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stead of concentrates, warning odors or dyes, or a bitter taste to discourage ingestion. Formulations may be limited to types (e.g., dry flowables) that limit exposure to handlers. Packaging design (e.g., water-soluble packets, lock-and-load container design) can also minimize handler exposure. Depending on acute toxicity, the pesticide may be required to be in child-resistant packaging. 27.3.2 FEDERAL REPORTING REQUIREMENTS FOR RISK INFORMATION
Section 6(a)(2) of FIFRA requires pesticide registrants to submit to EPA information concerning adverse effects of their products. The purpose of this requirement is to help ensure that EPA decisions to register a pesticide as well as terms and conditions of registration were correct and that a pesticide can be used without posing unreasonable adverse effects to human health and the environment. Information submitted under this rule involves toxicological and ecological studies, antimicrobial product efficacy failure data, and incident reports. Incident reports may involve humans, domestic animals, wildlife, plants, surface or ground water contamination, or property damage. When the EPA learns from a third party that a registrant knew about but did not report appropriate risk information, fines of several thousand dollars per case have been imposed on the guilty registrant. For incident reporting, the regulations specify detailed information to be provided with serious or rare incidents. Common or minor incidents, on the other hand, can be summarized as counts by product or active ingredient. For example, for the more serious human incidents, documentation is requested on the pesticide agent, the circumstances of exposure, and evidence of the type and severity of adverse effects. Exposure circumstances include how exposed, use site, situation (e.g., household application, field reentry), and evidence that the label directions were not followed. Adverse-effect information includes the route of exposure, list of signs and symptoms, results from medical laboratory tests, type of medical care sought (i.e., none, clinic, hospital emergency department, private physician, poison control center, or hospital inpatient), time between exposure and onset of symptoms, and estimated duration or amount of exposure, if available. Some registrants have elected to use the services of poison control centers to handle inquiries about adverse-effect incidents concerning their products. 27.3.3 NATIONAL PESTICIDE TELECOMMUNICATIONS NETWORK
The National Pesticide Telecommunications Network (NPTN) is a toll-free telephone service that provides pesticide information to any caller in the United States, Puerto Rico, or the Virgin Islands. The service is funded by the EPA to provide objective, science-based information about a wide variety of pesticiderelated subjects, including pesticide products, recognition and
management of pesticide poisoning, toxicology, and environmental chemistry. The service can provide chemical, health, and environmental information on more than 800 pesticide active ingredients incorporated into over 20,000 different products registered for use in the United States. The NPTN operates 7 days a week, excluding holidays, from 6:30 AM to 4:30 PM Pacific time. The NPTN can be reached by telephone at 1-800858-7378, by fax at 1-541-737-0761, or on the Internet at http://ace.orst.edu/info/nptn/.
Also at the NPTN is the National Antimicrobial Information Network (NAIN), which provides similar information about antimicrobial products-sanitizers, disinfectants, and sterilantsby phone or mail. The NAIN has the same operating schedule as the NPTN. Its telephone number is 1-800-447-6349; the fax number is 1-541-737-0761. The Web site address is http://ace.orst.edu/info/nain/. 27.3.4 WORKER PROTECTION STANDARD
In 1992, EPA revised the worker protection standard concerning protection of agricultural workers from pesticide exposure. The purpose of the revised standard is to reduce pesticide exposure among agricultural workers and thereby reduce the risk of pesticide poisonings and potential chronic effects. Under this standard, workers employed at farms, forests, nurseries, and greenhouses receive notification about pesticide applications so they may avoid treated areas. All workers must receive safety training about pesticides and appropriate protective measures. For hazardous pesticides, decontamination soap and water must be available on site. Emergency assistance in the form of transport to an appropriate medical facility must be provided in the case of a suspected pesticide poisoning, and the employer must provide information about the pesticide to which the person may have been exposed. Depending on the toxicity ofthe pesticide, additional requirements for personal protective equipment and restricted-entry intervals may also be required. Workers are excluded from entering a pesticide-treated area during the restricted-entry interval, with only narrow exceptions.
27.4 EVALUATING SURVEILLANCE SYSTEMS The purpose for evaluating surveillance systems is to ensure that available surveillance resources and funding are directed at important public health problems and to determine whether the surveillance systems are operating efficiently and effectively. Guidelines for evaluating surveillance systems are available (CDC, 1988) and are briefly summarized here. When evaluating surveillance systems, it is important to describe the public health importance of the health event of interest. The information provided earlier in this chapter supports the public health importance of acute pesticide-related illness and injury. Surveillance data indicate that a large number of cases occur annually, some of which are fatal. In addition, epidemiologic data suggest that acute poisoning is associated with
27.5 Case definition for Acute Pesticide-related IIInessand Injury
long-term health effects (Rosenstock et aI., 1990; Savage et aI., 1988; Steenland et aI., 1994). Finally, acute pesticide-related illness and injury is preventable through appropriate training and by taking appropriate safeguards. It is also important to assess the usefulness of the surveillance system. This can include describing the actions taken as a result of using data from the surveillance system (e.g., policy changes, regulatory changes, clinical practice changes, etc.) and describing other anticipated uses of the data (e.g., detecting disease magnitude and trends, detecting new pesticide hazards or new populations at risk, detecting epidemics, stimulating epidemiological research, and assessing the effectiveness of interventions). The attributes ofthe surveillance system should also be evaluated. Because some attributes can conflict with other attributes (i.e., excelling in one attribute may hamper the ability to satisfy another attribute), it is important to identify and strengthen those attributes that are most important to a particular surveillance system. It should be recognized that it may not be possible to fully achieve the less important attributes. The attributes that should be evaluated are as follows: Sensitivity. What proportion of the total number of cases is identified by the system? Substantial resources may be required to evaluate this attribute (e.g., extracurricular efforts to determine the annual incidence of the condition in the community). Systems without high sensitivity can be useful for monitoring trends, as long as the sensitivity remains relatively constant. Flexibility. How adaptable is the system to changing needs or operating conditions? A flexible system can handle changes in case definitions, reporting sources, and outcomes/diseases/exposures. Simplicity. The system should be as simple as possible. There are several measures to be considered. Is the case definition easy to apply? Are there multiple levels of reporting? What is the mechanism for transmitting case information/data? How extensive are the staff training requirements? What type of data analysis is required? Who are the users of the data and what is the mechanism for distributing reports/data? Acceptability. This attribute assesses the willingness of individuals and organizations to participate in the system. Several factors influence acceptability. What is the public health importance of the health event of interest? Is there timely recognition of an individual's contribution? Are the time and personnel costs onerous? Have reporting laws been enacted? Does the system provide useful information? Predictive value. It is important to maximize the proportion of cases reported to the system that actually have the condition. This is because system resources are used to confirm and investigate cases reported to the system. A system with a low predictive value for reported cases suggests that resources may be wasted when those cases are investigated. Inappropriate outbreak investigations may be conducted if a high number of false positives is reported.
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The predictive value is related to the sensitivity and specificity of the case definition, and the prevalence of the condition in the population. Increased specificity and prevalence leads to an increased predictive value. Representativeness. This assesses whether findings from the surveillance system can be generalized to the entire target population. It is also important to attempt to identify any population subgroups that may be systematically excluded from the surveillance system. As surveillance data are often used to calculate morbidity and mortality rates, thought should be given to ensure that the denominator, which is often obtained from a different source (i.e., census data), is comparable with respect to the demographics of the surveillance data in the numerator. As with sensitivity, substantial resources may be required to assess this attribute. Special studies may be useful that seek to identify all cases and then compare them to those cases reported to the system. Timeliness. This refers to the time interval between each step in the surveillance system. Among the more important time intervals to assess are the length of time between an event and its being reported to the surveillance system, the time required to identify trends and outbreaks, and the time then required to institute interventions. The importance of timeliness depends on the urgency of the public health problem and the availability of effective control measures. The growing use of computer technology holds promise for improving this attribute. Discussion When evaluating a surveillance system, conclusions and recommendations should be provided. An assessment should be made as to whether the surveillance system should be continued (i.e., Is the health condition under surveillance important? Can justification be made for the resources used by the system?). If it is to be continued, the need for any modifications to the system should be identified. Finally, when making recommendations for modifications, it is prudent to recall that the costs and attributes of the system are interdependent. Improvements in many of the attributes (sensitivity, representativeness, timeliness) will likely increase the costs of the surveillance system. In addition, improvements in one attribute may affect performance of another attribute. For example, improvement in predictive value may compromise sensitivity and may reduce simplicity. Therefore, these consequences should be considered when recommending modifications.
27.5 CASE DEFINITION FOR ACUTE PESTICIDE-RELATED ILLNESS AND INJURY A case definition is used to identify individuals with a health outcome of interest. It is needed both in epidemiologic studies and to conduct surveillance. The case definition for acute pesticide-related illness and injury can be simple or complex.
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Those that are used in the surveillance systems described earlier in this chapter are not identical but vary across the systems. In some instances, the clinical diagnosis may be the basis of the case definition. For example, conducting surveillance with vital status statistics involves identifying death certificates that contain ICD-9 E codes specific for accidental pesticide poisoning. The assigned ICD-9 codes are based on data supplied by the health care professional who completed the death certificate. Similarly, data from the BLS and the American Association of Poison Control Centers often involve medical outcome data supplied by the health care provider. The case definition and the clinical diagnosis serve different purposed. As such, some case definitions may provide a classification that differs from the clinical diagnosis. The case definition provides guidelines for assessing the certainty of the evidence regarding exposure and health effects. An example of this is the case definition for acute pesticide-related illness and injury developed for the National Public Health Surveillance System. In contrast, the purpose of the clinical diagnosis is to guide the immediate treatment course for an ill individual.
three areas: pesticide exposure, health effects, and evidence supporting a causal relationship between exposure and effect. A case of pesticide-related illness or injury is classified as being either definite, probable, or possible. The specific classification category is chosen depending on the level of certainty of exposure, whether health effects were observed by a health care professional, and whether there is sufficient toxicologic information to support a causal relationship between the exposure and the health effects. The cases classified into these categories must meet the following criteria:
National Public Health Surveillance System Case Definition The National Public Health Surveillance System (NPHSS) is a conceptual framework for all public health surveillance based on a consensus of practicing epidemiologists at the local, state, and national levels (Meriwether, 1996). Goals of the NPHSS include prioritizing surveillance activities and securing the necessary resources to conduct these activities. Acute pesticiderelated illness and injury is one of the conditions identified for inclusion in the NPHSS. The acute pesticide-related illness and injury case definition for the NPHSS was developed using a modified nominal group process (Jones and Hunter, 1995). The group consisted of experts from federal agencies (NIOSH, EPA, National Center for Environmental Health), nonfederal agencies (Council of State and Territorial Epidemiologists, Association of Occupational and Environmental Clinics), and state health departments or other state designees. Prior to the first meeting of the group, a proposed case definition was distributed. During the first meeting in September 1995, the case definition was reviewed and revisions were made. Following the meeting, a revised case definition was provided to each of the participants, along with a classification exercise that consisted of three "test" cases that each participant classified using the case definition. The classification exercise identified the need for several modifications to the case definition. Additional meetings held in April 1996 and November 1997, two subsequent classification exercises, and additional iterations via e-mail continued this process until consensus was achieved. Because public health agencies seek to prevent all adverse effects from regulated pesticides, the case definition is intended to be applied to any acute adverse health effect resulting from exposure to a pesticide product, including health effects due to an unpleasant odor, injuries from explosion of the product, allergic reactions, and effects associated with inert ingredients. The case definition requires the collection of information in
When insufficient toxicologic information is available to determine whether a causal relationship exists between the pesticide exposure and the health effects, a case is classified as "suspicious." This category is assigned when minimal human health effect data are available or when there are less than two published case series or positive epidemiologic studies linking health effects to the putative exposure agent. When convincing evidence for an exposure-health effect relationship is not present, the case is classified as "unlikely." A classification of "not a case" is assigned when there is strong evidence that no pesticide exposure occurred, when no new postexposure abnormal symptoms were reported, or when there is definite evidence of a nonpesticide causal agent. The case definition is complex. It requires knowing how to obtain information on exposure (i.e., knowing what environmental and medical tests should be conducted and what questions to ask); health effects (i.e., ability to review medical records and to solicit a medical history); and the causal relationship (i.e., knowledge about how to find and use appropriate references). It also requires knowledge and experience with assessing whether the exposure was sufficient to produce the observed health effects. Because of the skills, knowledge, and experience that are required to use this case definition, it is likely that, for any given case, the extent of agreement among raters will be not be total. There are several reasons for the complexity of the case definition. One is that there is no "gold standard" for pesticiderelated illness and injury (i.e., there is no symptom, sign, or test that is definitive for this condition). Therefore, identifying cases of pesticide-related illness and injury require assessment of the available information on exposure, health effects, and causal relationship. Unfortunately, because there is no "gold standard", it is difficult to determine the case definition's sensitivity, specificity, and positive predictive value. Likewise, it is difficult to assess the degree of misclassification that arises by using the
• Documentation of two or more new adverse health effects that are temporally related to a documented pesticide exposure • Consistent evidence of a causal relationship between the pesticide and the health effects based on the known toxicology of the pesticide from commonly available toxicology texts, government publications, information supplied by the manufacturer, or two or more case series or positive epidemiologic investigations
27.6 Limitations of Pesticide Poisoning Surveillance Data case definition. However, we think most would argue that this case definition reduces misclassification compared to other less rigorous definitions. Another reason for the complexity of the case definition is that it covers all classes of pesticides. This allows the case definition to be flexible. This case definition is also resource intensive. It requires having trained staff to collect and assess the information needed for case classification. It also requires staff to code and key the data into a database so that the data can be analyzed. A major strength of the standardized case definition and the standardized variables (described in Section 27.2.2) is that they allow data from participating surveillance systems to be aggregated. This aggregation will enhance knowledge about acute pesticide-related illness and injury. With this knowledge, the goal of surveillance can be realized: targeting public health resources toward the prevention of acute pesticide-related illness and injury. In conclusion, a case definition is needed to identify individuals with pesticide-related illness and injury. Currently, the case definition varies across surveillance systems. Where adequate resources exist, it is recommended that the case definition for the National Public Health Surveillance System be used. Regardless of the case definition that is used, all serve the purpose of identifying cases so that appropriate interventions can be targeted.
27.6 LIMITATIONS OF PESTICIDE POISONING SURVEILLANCE DATA When examining surveillance data, one needs to be mindful of their limitations. A discussion of these limitations follows. 27.6.1 DENOMINATORS
. A denominator is needed to calculate rates. Comparing the rate of the condition across different groups is needed to identify high-risk populations and to evaluate risk factors. Counts alone of a condition's occurrence may have little value for identifying disease risk factors. The difficulty of finding appropriate denominator information is one of the most obvious limitations of surveillance data. National populations provide one type of denominator for surveillance data. The absolute counts reveal striking differences in pesticide morbidity between industrialized and developing nations. Adjusting for relative population size, through the use of rates, emphasizes the disparity (see Tables 27.12 and 27.13). More refined comparisons require estimates of the number of pesticide-exposed individuals. Unfortunately, such estimates are generally either imprecise or unavailable. For occupational exposures, the most straightforward denominator is the number of pesticide-exposed individuals in the occupational group of interest (e.g., the number of licensed pesticide applicators). Regrettably, little information is available on unlicensed applicators or on the number of people performing agricultural work.
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The transient nature of the agricultural workforce in many areas and the fact that some agricultural workers may have entered the country illegally further complicate developing reliable denominators for agricultural pesticide exposure. Considering variations in exposure among the agricultural workforce adds complexity to the problem (e.g., what pesticides are used, method of application, duration of use, personal protective equipment). To calculate rates for nonoccupational pesticide-related illness, denominator information is needed on the use of pesticides by homeowners. This information is even more difficult to acquire than estimates of occupational pesticide users. Possible surrogates include data from the EPA survey of home and garden pesticide use that provided estimates of the number of containers and number of applications of pesticides for all households in the United States in 1990 (Whitmore et aI., 1992). Denominators can also be derived from pesticide use databases. Nationally, some survey data are available on the annual quantities of agricultural pesticides that were used (U.S. Department of Agriculture, 1998). Although few states require the reporting of pesticide use, New York and California have comprehensive systems for collecting data on pesticide use. Since 1990, California has required agriculturalists and professional applicators to report each pesticide application made in the state. New York has had a similar system in place since 1996. These reports provide, among other details, the identity of the pesticide used, the amount used, the number of acres or other units treated, and the crop or other site to which the pesticide was applied. Collecting all these data is a massive undertaking and provides an enormous amount of information on pesticide usage. These data have been used to identify risk factors associated with illness from restricted-use organophosphate pesticides in an agricultural setting (Weinbaum et aI., 1997). It should be noted that problems have arisen with denominators derived from pesticide use databases. Among these are the lack of information on the number of exposed workers and their duration and intensity of exposure. In addition, care must be taken when using total poundage as the denominator. Pesticides applied at low rates may exhibit exaggerated risk if the time required for application is similar to or greater than the length of time to apply high-rate pesticides. When evaluating statistics on pesticide-related illness, it is important to be mindful of the various factors that can influence the data. These include the method used to approximate the person-time at risk, the inherent toxicity of the pesticide, the method of application, the amount applied, the equipment used, and the skills of the applicators. Any analysis that compares groups based on only some of these factors makes the implicit assumption that the groups do not differ with respect to the other factors. Generalizing rates of poisoning to those outside the population that was investigated can be problematic. For example, the risks among field workers can vary depending on the crops they tend, the tasks they perform, climatic conditions, and their individual work practices. The most difficult factors to ascertain may be the most critical. Ideally, changes in field worker
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Figure 27.6 Strawberry harvesters in a field sprayed with insecticide. (Courtesy of Way ne T. Sanderson.)
tasks and practices over the time interval of interest should be determined. As such, when generalizing surveillance data on pesticide-related illness, the amounts of pesticide used or number of acres treated may not be adequate to assess the true risks that were experienced by field workers. 27.6.2 LIMITATIONS RELATED TO DEFINITIONS
Enumerating the population at risk is not the only problem in interpreting surveillance data. When using surveillance data or comparing data from various surveillance systems, it is extremely important to understand the case definition that was used. Different surveillance systems monitoring conditions related to pesticide exposure may assign different meanings to the terms "pesticide," "exposure," "case," and "related." In the United States, the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) defines pesticides to include "all substances and mixtures of substances" that "prevent, destroy, repel, or mitigate" any pest. Other statutes clarify that pests include all deleterious organisms, even bacteria, although pharmaceuticals are distinguished from pesticides. Pesticides consequently include sanitizers and disinfectants (e.g., chlorine) along with mothballs, rat baits, weed killers, fumigants, and many other substances. Few programs attempt to track the effects of all these products. Most surveillance systems track illnesses and injuries associated with exposure to insecticides, fungicides, rodenticides, and herbicides. Health effects from exposures to other pesticide products such as disinfectants and antibacterials are not universally included in surveillance systems. Each surveillance system sets its own threshold for the adverse effects recorded. Some systems attempt to identify only cases that resulted in medical consultation or only those that included hospitalization. Other data sources accept self-reports of illness or injury. Even when systems use the same objective
standard, such as hospitalization, criteria for hospital admission may vary with culture and economic circumstances. Additional discrepancies stem from variations in criteria for what constitutes pesticide-related illness and injury. Some have argued that surveillance should consider only those intoxications in which the pesticide acts on human victims by the same mechanism by which it controls pests (i.e., cholinesterase inhibition for organophosphate pesticides). Some critics have also taken issue with the inclusion of health effects that may be related to the "inert" (nonpesticidal) ingredients in a pesticide product. Public health surveillance more typically considers all characteristics of pesticide products that can cause harm. These surveillance systems record illness and injury resulting from exposure to the pesticide products (i.e., the formulated product not just the active pesticide ingredient). Additionally, these surveillance systems record the occurrence of bums from pesticide fires, traumatic injuries from pesticide explosions, illnesses resulting from purposeful (i.e., homicidal or suicidal) and accidental ingestion, allergic reactions to pesticide products, and effects associated with inert ingredients. Most controversially, they may record reactions to a pesticide's noxious odor. Evaluation of the causal relationship between pesticide exposure and adverse health effects is complicated. Some surveillance systems accept the clinician's diagnosis in determining the relationship, whereas other systems have more complex case definitions and classification schemes. It is useful for surveillance systems to independently examine the relationship between exposure and health effects because the clinical diagnosis may not be correct (e.g., health care professionals may report cases that they suspect may be related to pesticide exposure, but for which they are not certain). Systems that examine the relationship between exposure and health effects must take several factors into account, including the wide range of symptoms various pesticides can produce, the nonspecific nature of reported signs and symptoms (especially in less severe illness), limited or nonexistent analytical environmental data on the individual's exposure, lack of clinical/biological measures of pesticide absorption, and inappropriate use of available tests. Evaluating anxiety poses a particular problem, as many common insecticides are neurotoxic and may elicit anxiety pharmacologically; on the other hand, anxiety unaccompanied by physical exposure often mimics toxic effects. Rarely can physical findings or test results clarify this issue. In response to these difficulties, pesticide surveillance systems typically classify cases into one of several categories that reflect the certainty of the relationship between exposure and illness (see Section 27.5). When examining surveillance data, care must be taken not to confuse reports with confirmed cases. This is especially true for surveillance systems that include reports from affected individuals and nonmedical personnel and where no investigation is undertaken to follow up the report. Following an appropriate investigation, reports are classified according to the National Public Health Surveillance System case definition as "confirmed" (i.e., defined as definite, probable and possible cases), "nonconfirmed" (i.e., defined as unlikely cases, those deter-
27.6 Limitations of Pesticide Poisoning Surveillance Data
mined not to be a case, and those where insufficient information is available), or "suspicious" (i.e., insufficient toxicologic information is available to determine whether a causal relationship exists between the exposure and the health effect). Some might argue that including case reports with a lower degree of certainty compensates for underreporting. However, this may not be an appropriate remedy for underreporting because the nonconfirmed cases may not be representative of the unreported true cases. 27.6.3 LIMITATIONS RELATED TO SENSITIVITY No surveillance system succeeds in identifying every event of interest. Most surveillance systems capture from 5 to 80% of cases that occur (Cates and Williamson, 1994). It should be recalled that even surveillance system data without high sensitivity can be useful for monitoring trends, as long as the sensitivity remains relatively constant. The likelihood that a case of pesticide-related illness will be reported may vary with occupation, social status, and the circumstances of exposure, and even the individual pesticide. Surveillance systems that rely on a variety of sources for case ascertainment are likely to be more representative of the universe of cases. Physician reporting is one of the most common mechanisms for surveillance. This method is the mainstay of many communicable-disease reporting systems, but it is not necessarily the most effective method for surveillance of pesticide poisoning. Physician reporting requires that the affected individual seeks medical care, that a diagnosis of pesticide-related illness or injury is made or suspected, and that the physician is aware of the need to report the suspected case. Barriers exist at each of these steps that can hamper physician reporting. For example, some populations at greatest risk for pesticide exposure are less likely to seek medical attention except for more severe illness. Those ill individuals who do not seek health care may be detected only after an investigation is conducted into a sentinel case involving a severely ill co-worker or family member who sought medical care. Furthermore, pesticide-related illness is not routinely encountered by a majority of primarycare providers in the United States and most receive minimal training on recognition of environmental or occupational illness (Institute of Medicine, 1988; Pope and Rall, 1995). In addition, the ability to make the diagnosis is complicated by the fact that symptoms are often nonspecific and by the lack of readily available specific laboratory tests to measure the pesticide, its metabolites, and the effect of the pesticide. Even when tests are available, they are frequently not performed or are not performed sufficiently promptly to detect the abnormality. This problem of recognition is found in industrialized and developing countries alike (Keifer et aI., 1996). Once the diagnosis of pesticide poisoning is made, there are many reasons for failing to report it. Despite broadly worded reporting guidelines, physicians are often reluctant to report cases that they feel are unconfirmed clinically. Additional barriers to
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physician reporting include protection of a patient who fears loss of a job and ignorance regarding the reporting requirement. Cultural pressures to downplay the hazards of pesticides may prevent a physician residing in an agricultural community from reporting cases. 27.6.4 LEGITIMATE USES FOR SURVEILLANCE DATA With so many difficulties and limitations, attempting surveillance of pesticide-related conditions may seem futile. However, even problematic surveillance data can advance public health when used with appropriate caution. Surveillance data can be useful for identifying emerging pesticide hazards and new populations at risk. When these emerging problems are identified, they present an opportunity to implement interventions that will prevent subsequent illness. For example, the identification of several California grape harvesters who became ill after exposure to phosalone led directly to the withdrawal of this pesticide. Although phosalone had been in use for nearly 20 years on crops that require minimal to moderate hand labor activity, it was eliminated only after it began to be used more widely on grapes, a crop requiring more extensive hand labor activity. This problem was detected when the ill grape harvesters were identified using surveillance data (O'Malley and McCurdy, 1990). A similar scenario was repeated in 1993 in Washington when 26 workers at 19 orchards became ill during a period of several months. The outbreak and the ensuing investigation resulted in the suspension, and eventual withdrawal, of mevinphos use in Washington apple and pear orchards (CDC, 1994; Washington State Department of Agriculture, 1994). Another example involves surveillance data from a Nicaraguan regional health center that resulted in identification of an epidemic of acute poisonings linked to cholinesterase-inhibiting insecticides. The identification of the epidemic resulted in a prompt education campaign to reduce poisonings from a powdered formulation of carbofuran. In addition, policy recommendations were made to encourage importation of less hazardous pesticides (McConnell and Hruska, 1993). These situations exemplify the public health importance of prompt health care provider reporting and appropriate public health agency follow-up. Although surveillance that includes active case follow-up is resource intensive, it provides the opportunity to gather information that can be used to develop strategies for prevention. The information obtained through case follow-up is often not available when cases are identified retrospectively through surveys of existing data sources. One is left to wonder whether better surveillance would have reduced the health and financial costs associated with the indoor use of methyl parathion. Since 1984, homes and businesses in at least five different states have been illegally sprayed with methyl parathion. However, corrective action was not enacted until 1997. These events occurred in New York, Ohio, Michigan, Mississippi, and Illinois, resulting in expensive relocation and remediation activities ( see Fig. 27.7). Relocations have involved more than 1500 individuals. The cleanup costs for these
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tems to provide a comprehensive estimate of the magnitude of pesticide-related illness and injury. 27.6.5 MECHANISMS TO STRENGTHEN THE SURVEILLANCE OF ACUTE PESTICIDE-RELATED ILLNESS
Figure 27.7 Worker conducting methyl parathion remediation to a dwelling in Mississippi. (Courtesy of the V.S. Environmental Protection Agency.)
incidents are estimated at more than $90 million (EPA, 1997a). Little information is available on the health effects associated with these incidents. However, one published report describes methyl parathion-related illness among seven siblings, two of whom had a fatal outcome (CDC, 1984). In addition, another government report summarizing the 1995 Ohio investigations found that 20% or more of respondents reported symptoms during the 2 weeks following methyl parathion application (NCEH, 1996). These investigations also found that 20 out of 50 (40%) indoor pets present in these homes died within 2 weeks of methyl parathion application (pointing out the value of pets as sentinels of human exposure and illness). To prevent additional exposure incidents, a memorandum of agreement between the EPA and the manufacturers of methyl parathion emulsifiable concentrate became effective in January 1997. The agreement included recall of particular products, changes in packaging and labeling, as well as the addition of an odor-producing agent (EPA, 1997b). As the problem with methyl parathion has existed since at least 1984 in several different states, a national surveillance system with active case follow-up may have resulted in a more timely identification of the magnitude of this problem and earlier adoption of preventive and regulatory measures. Surveillance data can also be a source of cases for formal epidemiologic studies. Important morbidity and mortality studies can be designed that involve a cohort of individuals poisoned by a specific pesticide or group of pesticides. For example, Steenland et al. (1994) examined a group of workers who had a history of acute poisoning by organophosphate insecticides to determine if these workers had chronic neurologic sequelae. The workers in this study were identified using surveillance data collected by the California Environmental Protection Agency. As noted earlier, surveillance data can also be used to examine the magnitude of pesticide-related illness and to assess trends. More work is needed to determine the best approach for combining data from the many different surveillance sys-
Some very specific actions can be taken to enhance existing surveillance systems. Some of the changes are already underway and an evaluation of their efficacy should be possible in the near future. One important action is the need to improve training of primary health care professionals. The EPA has launched a new initiative to target health care professionals with educational and training opportunities on pesticide-related health issues (EPA, 1998). This initiative involves strategies to ensure that primary health care professionals can recognize health effects from pesticide exposure. It includes mechanisms to enhance their abilities to diagnose illness and manage exposures; engage in preventive management (at the case and community level), appropriately report exposures and illnesses, and access appropriate resources when necessary. To foster success, these activities should be coupled with the education of workers and consumers on many of these same topics. Increasing the quality and availability of biological monitoring tools would aid surveillance by assisting with confirmation of cases. The development of new biomarkers of exposure and health effects is also an extremely important area that would enhance surveillance data. Reliable and affordable screening methods for field and clinical settings must be available if they are to be used routinely in developing countries and under the constraints of managed health care systems. Most biological markers of exposure and health effects of pesticides are still primarily research tools. Even cholinesterase monitoring, the most commonly used measure of biological effects from exposure to organophosphate and carbamate insecticides, suffers from lack of standardization. Both the handling of specimens and the assay method require standardization to obtain valid test results (Wilson et aI., 1996). Although much progress has been made in delineating these problems, they have not been satisfactorily resolved (Wilson et al., 1997). The ability to provide summary data and direct feedback to the medical community, agricultural workers, pesticide manufacturers, commercial pest control firms, and policymakers is a critical aspect of surveillance. Although existing surveillance systems communicate their findings to some degree, this is an area where significant improvements can be made that will strengthen surveillance. The ability to aggregate data across states, combined with increased dissemination of information, will result in a better understanding of the nature of acute pesticide-related illness and injury. The costs of pesticide-related illness and injury are relatively unknown. Most surveillance systems provide minimal information on the severity of illness that can be used in estimating the cost of illness. Some measures of severity are included in the
27.7 Fundamentals of Epidemiology core standardized variables developed for the SENSOR surveillance system on pesticide-related illness and injury. A matrix for ranking illness severity based on signs, symptoms, time loss from work or regular life activities, and days of hospitalization is currently under development by NIOSH. Much can be learned about the costs of pesticide-related illness and injury if surveillance systems are enhanced to collect measures of illness severity.
27.7 FUNDAMENTALS OF EPIDEMIOLOGY When conducting surveillance of pesticide-related disease and injury, a decision must be made as to whether the pesticide exposure caused the documented illness. Epidemiologic studies are often the source of information used to make these decisions. Therefore, although the emphasis of this chapter is on surveillance, we think it is important to describe the basic principles of epidemiology and the role of epidemiology in identifying health effects related to pesticide exposure. We will begin by defining epidemiology. Epidemiology is the study of the distribution and determinants of disease in human populations. Epidemiologic studies compare the rates of disease in populations exposed to various risk factors to populations that are not exposed and, based on these comparisons, evaluate the factors that may cause or influence disease. 27.7.1 PRINCIPLES OF EPIDEMIOLOGY
A major premise of epidemiology is that disease is not simply a random occurrence, but is the result of various causal factors (Checkoway et aI., 1989). These causal, or risk, factors influence the distribution of disease in a popUlation. Differences in disease patterns may be explained by the differential distribution of risk factors between populations. These causal, or risk, factors include age; sex; race or ethnicity; genetic susceptibility; personal lifestyle factors such as smoking habit, diet, exercise, drug use, and weight; and occupational or environmental exposure to various chemical and physical agents. Diseases may be either acute or chronic. Acute diseases occur soon after an exposure, whereas chronic diseases develop many years after exposure. Examples of acute disease are respiratory infections caused by bacteria and viruses and eye and upper respiratory irritation caused by sulfur dioxide or ozone. Examples of chronic disease include pneumoconiosis caused by crystalline silica or coal dust, and leukemia caused by benzene and radiation. The time period between initial exposure to a causal agent and disease detection can be divided into the induction period-time between causal action and disease initiation-and the latency period-time between disease initiation and detection. The longer the induction-latency period, the more difficult it is to link causal factors to the disease outcome. Pesticides are known to cause both acute diseases such as systemic poisonings and skin rashes and chronic diseases such as cancer, lung disease, and reproductive problems.
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27.7.2 EPIDEMIOLOGIC STUDY DESIGNS
Epidemiologists are very rarely able to control the risk factors of study subjects, such as in a randomized controlled trial. Therefore, epidemiology is largely an observational science, relegated to documenting the past and present risk factors and evaluating the association between these risk factors and disease status (Kleinbaum et aI., 1982). Epidemiological research begins with a hypothesis to be tested. For example, it may be hypothesized that farmers with a history of applying herbicides to their crops have a greater risk of developing cancer than the general population. The epidemiologist then designs a study to test this hypothesis. The epidemiologist may choose from among the following types of observational (nonrandomized) study designs. For a fuller discussion on these study designs and their interpretations, the reader is referred to standard epidemiologic textbooks (Kleinbaum et aI., 1982; Mausner and Kramer, 1985). Cohort Study Cohort studies begin with enumeration of a population (the cohort) that shares common characteristics or risk factors (e.g., exposures). The cohort's health experience is then evaluated over a defined time period. The basic question addressed by a cohort study is: Are those with exposure more (or less) likely to develop disease compared to those who are unexposed? The control or reference populations are generally national or regional (i.e., state or province) populations, which provide generally stable disease rates. Cohorts can be enumerated currently and followed forward in time. This is termed a prospective cohort study. As the cohort is followed through time, individual exposures and diseases are documented. The rate of disease among those individuals with particular exposures are compared to the rate of disease among the unexposed reference group. Another type of cohort study is the retrospective cohort study. In this type of study, a cohort is enumerated in the past (i.e., a cohort that has been exposed some time in the past) and followed up to the present to identify those individuals who develop disease. The disease rates in the
Figure 27.8 Farmer applying insecticide with a handheld sprayer to control flies on cattle. (Courtesy of Wayne T. Sanderson.)
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cohort are compared to those occurring in an unexposed comparison population. This type of study has also been termed a historical cohort study. An example of how a cohort study was used to evaluate the association between pesticide exposures and disease is provided by a study of 20,245 agricultural pesticide applicators in Sweden (Wiklund et aI., 1989). The cohort, consisting of all applicators who had been licensed between 1965 and 1976, was followed in the Swedish Cancer Register from the date individuals received their license until 1982 or until their death if it occurred before 1982. The number of cancer cases in the applicator cohort was compared to the number of cases occurring in 5-year age and sex groups during the same time period in the whole Swedish population. The number of cancer cases in the Swedish population provided the number of cases that would be expected in the applicator population. A total of 558 malignant tumors was found compared with 649.8 expected, resulting in a statistically significantly decreased standardized incidence ratio (SIR) of 0.86. A finding that reaches statistical significance implies that the finding is unlikely to be due to chance. No cancer rates were found to be significantly increased, although the pesticide applicators had higher risk rates for testicular cancer, tumors of the nervous system and endocrine glands, and Hodgkin's disease. The Swedish pesticide applicator study was a retrospective cohort study; the cohort was enumerated as of some time in the past and then followed over time to estimate cancer rates compared to rates that occurred in the general population. The Agricultural Health Study, which is being conducted in Iowa and North Carolina by the National Cancer Institute, is an example of a prospective cohort study (Alavanja et aI., 1996). This is a large study of farmers, commercial pesticide applicators, and their families, who are enrolled and asked to answer questionnaires about their lifestyles, work practices, and exposures at 5-year intervals over a 20-year period. The study includes over 75,000 adult subjects. Pesticide usage information and exposures are being collected in an attempt to relate exposure to disease outcomes. Disease rates are determined at regular intervals for the cohort, and potential risk factors are assessed by the information collected on the questionnaires and contained in disease registry records. Comparisons will be made to data from the National Health and Nutrition Examination Survey (NHANES). In addition, an internal reference population will be used consisting of the unexposed segment of the Agricultural Health Study cohort. As this study is currently ongoing, few results are available. Cohort studies have several strengths. They provide information on the time lag between the first known exposure and disease detection, and they can be used to evaluate risk for many different diseases. They also measure exposure before disease occurs, resulting in less recall error by study subjects. However, cohort studies are costly, requiring long-term commitment of time and resources, as well as a large sample size. A cohort study may not be possible if data for constructing a retrospective cohort are incomplete. Also, retrospective cohort studies are usually restricted to investigating fatal diseases
because nonfatal diseases are often not recorded historically. In contrast, prospective cohort studies may document nonfatal diseases as they occur. The primary cost for conducting cohort studies is in obtaining exposure data on a large number of subjects of which only a small proportion develop the disease of interest. Case-Control Study The basic question addressed by the case-control study is: Are the people with existing disease more or less likely to have been exposed than those without the disease? The distinguishing feature of case-control studies is that subjects are selected based on their disease status, reducing cost by limiting exposure assessment to only cases of disease and a control group. Cases may be identified from disease registries, hospital or clinical records, or volunteers. Controls are selected to be similar to the cases with the exception of disease status. Exposure information is developed from existing records or a detailed self-reported questionnaire. This information is used to compare the exposure prevalence between the cases and the controls. An example of how a case-control study was used to evaluate the association between pesticide exposure and disease is provided by a study of soft-tissue sarcoma, Hodgkin's disease, and nonHodgkin's lymphoma in Kansas (Hoar et aI., 1986). All cases of these diseases diagnosed from 1976 to 1982 among white male Kansas residents aged 21 years or older were identified in a state-based cancer registry. Tissue specimens were obtained from the cases to confirm their diagnosis. Each case was matched to three controls who were also white, lived in Kansas, and were within 2 years of the age of the case. Controls were selected either from Medicare files or through telephone calls using random-digit dialing (Waksberg, 1978). The cases, controls, or their next of kin if they were deceased were interviewed by telephone to determine their work practices and use of pesticides and other agricultural chemicals. Herbicide use was associated with non-Hodgkin's lymphoma [odds ratio (OR), 1.6,95% confidence interval, 0.9-2.6]. Neither soft-tissue sarcoma nor Hodgkin's disease was associated with pesticide exposure, but non-Hodgkin's lymphoma increased significantly with estimated number of days of herbicide exposure per year. Case-control studies are particularly useful for studying rare diseases or diseases with a long latency period since first exposure. They are relatively inexpensive because they involve fewer subjects than cohort studies and can be completed in a relatively short time. However, the information collected on exposures occurs after the disease has been diagnosed, which may make diseased people more (or less) likely to remember previous exposures (recall bias). Also, diseased individuals may be more motivated to participate in a case-control study than a healthy control (selection bias). Cross Sectional Study In a cross-sectional study, exposure and disease are evaluated at the same time. The prevalence of disease is measured in a defined population at a particular point in time, while the exposures of the individuals are also
27.7 Fundamentals of Epidemiology
measured at that time. For example, the rate of occurrence of symptoms in workers exposed to a particular pesticide would be compared to the rate in workers who were unexposed. Crosssectional studies are suitable for evaluating nonfatal diseases or measuring physiologic responses to workplace exposures. Data are collected using clinical examinations, symptom surveys, or direct biological or physical measurements. A critical problem with the cross-sectional design is that it may be difficult to determine whether the onset of disease began before or after the exposure. Also, cross-sectional studies may miss many diseases of short duration. Because cross-sectional studies typically only include currently employed workers, retirees or other workers who terminated employment because of ill health, possibly attributable to their exposures, are not studied. It is these noncurrently employed individuals who may be the most relevant subjects for investigating delayed or progressive health outcomes. An example of a cross-sectional study is one that was used to evaluate the association between fumigant exposure and disease among structural fumigation workers in Florida (Calvert et al., 1998). In this study, 123 structural fumigation workers and 120 unexposed controls were interviewed and examined. Nerve conduction, vibration, neurobehavioral, olfactory, visual, and renal function testing was conducted. The median lifetime duration of methyl bromide and sulfuryl fluoride exposure among workers was 1.20 and 2.85 years, respectively. Sulfuryl fluoride exposure over the year preceding examination was found to be associated with significantly reduced performance on one cognitive test and on olfactory testing. In addition, fumigation workers had significantly reduced dexterity of the dominant hand. A nonsignificantly higher prevalence of carpal tunnel syndrome was also observed among the fumigation workers. The authors concluded that occupational sulfuryl fluoride exposures may be associated with subclinical effects on the central nervous system, including effects on olfactory and some cognitive functions. However, no widespread pattern of cognitive deficits was observed. The peripheral nerve effects (reduced hand dexterity and carpal tunnel syndrome) were likely due to ergonomic stresses experienced by the fumigation workers. Ecologic Study Perhaps the crudest approach to evaluating exposure-disease associations is the ecologic study. In this type of study, disease rates are compared between geographic areas rated according to their estimated extent of exposure. The units of exposure correspond to geographical areas rather than individuals. An example of an ecologic study is ~ne used to evaluate the association between cancer and dibromochlorpropane (DBCP) contamination in Fresno County, California, drinking water (Wong et al., 1989). All cases of gastric cancer and leukemia occurring between 1960 and 1983 in Fresno County were identified by the California Vital Statistics office. The cancer rates were calculated using the 1960, 1970, and 1980 census data stratified by age, sex, and race. The cancer rates were compared by areas in the county stratified by the concentration of
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DBCP found in drinking water. No correlation was found between gastric cancer and leukemia mortality rates and DBCP concentrations. Ecologic studies use readily available data that have been collected for other purposes, and they can be done relatively quickly. However, ecologic studies are severely limited because they do not associate exposure to individuals. They do not control for other exposures (confounding) that may be associated with the disease of interest, and they can be significantly influenced by the migration of individuals into or out of the geographic area (selective migration). Although ecologic studies do not provide firm conclusions about the association between exposures and disease, they are used to guide future, more indepth research studies. 27.7.3 EVALUATING PESTICIDE HEALTH INFORMATION Information on the toxicity of pesticides can be found in textbooks, in journal articles, and from information provided by pesticide producers. Each of these sources may be valuable, but each may also have particular bias. When relevant information is identified, one must decide whether to trust the information and whether the information can be generalized to the case at hand. In many medical disciplines, information relevant to a particular question may be summarized in systematic reviews that provide valid conclusions. Unfortunately, such reviews are uncommon with respect to the health effects associated with specific pesticides. As such, skills are needed both to efficiently search the literature for relevant information and to interpret the validity of any information that is discovered. There are several approaches for identifying relevant literature. These include asking knowledgeable colleagues (and pesticide producers), reviewing references cited in textbooks, and using an electronic bibliographic database such as PUBMED. The limitation of asking colleagues and using textbooks is that these sources may not be up to date. The most up to date source of relevant information is obtained by searching PUB MED. Accessing this database has become a basic and easily acquired skill. There is no charge for accessing this database through the National Library of Medicine (http://www.nlm.nih.gov/). Assessing the validity of a study can be more complicated (Levine et al., 1994). In cohort and cross-sectional studies, it is important that the exposed and control groups are similar with respect to all factors that may affect the outcome, save the exception that the exposed group has the exposure of interest. Practically, this means that both groups should be similar in age, gender, race, socioeconomic status, cultural background, and social habits (e.g., smoking and alcohol consumption). The study investigators should demonstrate that these characteristics are comparable or use statistical techniques to adjust for differences. In a case-control study, it is important that the cases and controls are similar with respect to important determinants of the disease of interest. In addition, controls should be randomly drawn from the same population from which cases were drawn
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(i.e., controls should have similar opportunity for exposure). For example, in a case-control study examining the association between pesticide exposure and cancer, it would be inappropriate to have the control group consist entirely of white-collar professionals because of their minimal opportunity for pesticide exposure. Even when investigators take appropriate precaution to ensure comparability for known risk factors, there may be a pronounced imbalance in the distribution of risk factors that are unknown to the investigators. It may be these unknown and unmeasured risk factors that are responsible for any observed findings. Other factors to consider when assessing the conclusions of an observational study include assessing the temporal sequence of events. It is important that the investigators document that the exposure preceded the disease or outcome of interest. It is also important to observe a dose-response relationship. Attributing a particular outcome to a pesticide exposure can be made with more confidence if risk of the outcome increases with increasing amount or duration of exposure. It is also important to ensure that investigators minimized bias in the collection of data (blinding data collectors and study subjects to the study hypotheses). The magnitude of the risk can also be helpful in assessing the validity of a study. With very large values of risk, one may be more confident that bias or uncontrolled risk factors are not responsible for the outcome. It should be noted that none of the conditions described in this section is either necessary (with the exception of temporal sequence) or sufficient to prove that a causal association exists between an exposure and an illness. The topic of causality is complex and will not be discussed further. An excellent review of this topic can be found elsewhere (Rothman and Greenland, 1998).
27.8 INTERNET AND TELEPHONE RESOURCES FOR PESTICIDE INFORMATION With the advent of the Internet, important sources of pesticide information are now readily available. Some Internet sites that are useful for pesticide-related illness and injury surveillance follow. Bureau of Labor Statistics This site contains pesticide illness tabulations that supplement those provided in the BLS annual reports. Information is available only for illnesses that result in lost work time. The Web site address is http://stats.bls. gov/oshhome.htm. California Environmental Protection Agency This site is a source for consumer fact sheets and provides access to several useful databases (pesticide product database, chemical ingredient database, and a company information database). The Web site address is http://www.cdpr.ca.gov.
EXTOXNET This is a nice resource for information on specific pesticides. In includes a useful search engine. This site is a cooperative effort of the University of California-Davis, Oregon State University, Michigan State University, Cornell University, and the University of Idaho. The Web site address is http://ace.orst.edu/info/extoxnetighindex.html. National Agricultural Statistics Service, U.S. Department of Agriculture This site provides information on agricultural pesticide usage. The Web site address is http://www.usda. gov/nass/. National Agricultural Workers Survey This site provides data on U.S. farm workers. Data available at this site may provide useful estimates of the number of crop and livestock workers. These data are available at the county level. The Web site address for NAWS is http://www.dol.gov/dol/asp/publiC/ programs/agworker/naws.htm. National Pesticide Telecommunications Network A cooperative effort of Oregon State University and the EPA, this site provides a toll-free telephone service that provides pesticide information (1-800-858-7378). Information can also be obtained at its Web site (http://ace.orst.edulinfo/nptn). Pubmed The medical literature can be searched for information on specific pesticides. Searches can be conducted free-ofcharge. The Web site address for PUB MED is http://www.nlm. nih.gov. U.S. Environmental Protection Agency This site contains a large amount of information, including consumer fact sheets, information on pesticide regulations, and pesticide product information. The Web site address is http://www.epa.gov/pesticides/. Pesticide product information can be found at http:// www.epa.gov/opppmsdIIPPISdataiindex.html. The manual titled Recognition and Management of Pesticide Poisonings is available at http://www.epa.gov/pesticides/safety/healthcare. University of Nebraska-Lincoln Institute of Agriculture and Natural Resources This Web site contains a comprehensive listing of bookmarks for pesticide-related topics. It includes links to sites on all aspects of pesticides (education, databases, health and safety, pesticide manufacturers, laws and regulations, newsletters, and environmental protection) and includes links to other sites mentioned in this section. The bookmarks are verified and updated at least twice a year. This site is a good place to begin a search on a pesticide-related topic. The address for these pesticide-related bookmarks is http://pested.unl.edu/pestbkmk.htm.
27.9 CONCLUSIONS A comprehensive, national surveillance system for acute pesticide-related illness and injury does not currently exist. However, this chapter describes several surveillance systems for
References pesticide-related illness and injury, each having strengths and weaknesses. Some systems, such as TESS, are most useful for assessing magnitude and trends. Others (e.g., state-based surveillance systems) are more useful for timely identification of outbreaks and emerging problems. Efforts are being made to standardize data collection. Standardization of data collection will facilitate linkage of data across surveillance systems to create a fuller understanding of the acute pesticide-related illness and injury problem. A comprehensive, national surveillance system may be attainable through standardization and information sharing across surveillance systems.
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Arizona Department of Health Services (1996). "Pesticide Poisoning Surveillance Annual Reports 1996." Bureau of Epidemiology and Disease Control Services, Office of Environmental Health Investigation and Surveillance Section, Arizona Department of Heath Services, Phoenix. Blondell, J. (1997). Epidemiology of pesticide poisonings in the United States, with special reference to occupational cases. Occup. Med. 12,209-220. Bureau of Labor Statistics (BLS) (1998). "Occupational Injuries and Illnesses: Counts, Rates, and Characteristics, 1995." Bulletin 2493, U.S. Department of Labor, Washington, DC. Burger, M. (1998). Exposure on pesticides in Uruguay: Impact in human health. In "International Conference on Pesticide Use in Developing Countries: Impact on Health and Environment, February 23-28, 1998, San Jose, Costa Rica," Book of Abstracts, p. 215. Calvert, G. M., Mueller, C. A., Fajen, J. M., Chrislip, D. w., Russo, J., Briggle, T., FJeming, L. E., Suruda, A. J., and Steenland, K. (1998). Health effects associated with sulfuryl fluoride and methyl bromide exposure among structural fumigation workers. Am. J. Public Health 88, 1774-1780. Caldwell, S. T., and Watson, M. T. (1975). Hospital survey of acute pesticide poisoning in South Carolina, 1971-1973. J. S.C. Med. Assoc. 71,249-252. Caldwell, S. T., Barker, M, Schuman, S. H., and Simpson, W. M. (1997). Hospitalized pesticide poisonings decline in South Carolina, 1992-1996. J. S. C. Med. Assoc. 93, 448-452. Castro, R., Morera, N., and Jarquin, C. (1998). Epidemiological surveillance system for pesticide intoxications: The experience in Costa Rica, 1994-1996. In "International Conference on Pesticide Use in Developing Countries: Impact on Health and Environment, February 23-28, 1998, San Jose, Costa Rica," Book of Abstracts, p. 219. Cates, W., Jr., and Williamson, G. D. (1994). Descriptive epidemiology: analyzing and interpreting surveillance data. In "Principles and Practice of Public Health Surveillance" (S. M. Teutsch and R. E. Churchill, eds.), pp. 96-135. Oxford Univ. Press, New York. Centers for Disease Control (CDC) (1984). Organophosphate insecticide poisoning among siblings-Mississippi. Morbility and Mortality Weakly Report 33, 592-594. Centers for Disease Control (CDC) (1988). Guidelines for evaluating surveillance systems. Morbility and Mortality Weakly Report 37, suppl. No. S-5: 1-18. Centers for Disease Control and Prevention (CDC) (1994). Occupati onal pesticide poisoning in apple orchards-Washington State, 1993. Morbility and Mortality Weakly Report 42, 993-995. Centers for Disease Control and Prevention (CDC) (1999). Farm worker illness following exposure to carbofuran and other pesticides-Fresno County, California, 1998. Morbility and Mortality Weakly Report 48, I 13- I 16. Chan, T. Y. K., Critchley, J. A. J. H., and Chan, A. Y. W. (1996). An estimate of the incidence of pesticide poisoning in Hong Kong. Vet. Hum. Toxieol. 38, 362-364. Checkoway, H. Pearce, N. E., and Crawford-Brown, D. J. (1989). "Research Methods in Occupational Epidemiology." Oxford Univ. Press, New York, NY. Cilovtsev, V. E., Ivanov, A. A., Perevalov, A., and Shakin, V. V. (1998). PESTOTEST: A system of the pesticide monitoring in Russia. In "International Conference on Pesticide Use in Developing Countries: Impact on Health and Environment, February 23-28, 1998, San Jose, Costa Rica," Book of Abstracts, p. 223. Cole, D. C., McConnell, R., Murray, D. L., and Pacheco, A. E (1988). Pesticide illness surveillance: the Nicaraguan experience. Bull. Pan Am. Health Org. 22, 119-132. de Alwis, L. B. L., and Salgado, M. S. L. (1988). Agrochemical poisoning in Sri Lanka. Forensic Sci. Int. 36, 81-89. de Oliveira, L. E (1998). The use of a "sentinel case" to improve the epidemiological surveillance of pesticides poisoning. In "International Conference on Pesticide Use in Developing Countries: Impact on Health and Environment, February 23-28, 1998, San Jose, Costa Rica," Book of Abstracts, p.214. Fabritius, K., and Balasescu, M. (1996). Acute nonoccupational intoxications with pesticides in Romania: A comparative study from 1988 to 1993. Toxicol. Lelt. 88,211-214.
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Surveillance of Pesticide-Related Illness and Injury in Humans
Felberg, L., Litovitz, T. L., Soloway, K A., and Morgan, J. (1996). State of the nation's poison centers: 1994 American Association of Poison Control Centers Survey of US Poison Centers. Vet. Hum. Toxicol. 38,214--219. Freund, E., Seligman, P. J., Chorba, T. L., Safford, S. K., Drachman, J. G., and Hull, H. F. (1989). Mandatory reporting of occupational diseases by clinicians. 1. Am. Med. Assoc. 262,3041-3044. Garcia-Repetto, K, Soria, M. L., Giminez, M. P., Menendez, M., and Repetto, M. (1998). Deaths from pesticide poisoning in Spain from 1991 to 1996. Vet. Hum. Toxico!. 40, 166-168 Giesecke, J. (1999). Choosing diseases for surveillance. Lancet 353,344. Hayes, W J., Jr. (1976). Mortality in 1969 from pesticides, including aerosols. Arch. Environ. Health 31, 61-72. Hayes, W J., Jr, and Vaughn, W K. (1977). Mortality from pesticides in the United States in 1973 and 1974. Toxicol. Appl. Pharmacol. 42, 235-252. Health and Safety Executive (1997). "Pesticide Incidents Report 1996/97." INTS03 9/97 C20. Health and Safety Executive, Sudbury, Suffolk, UK. Hoar, S. K., B1air, A., Holmes, F. F., Boysen, C. D., Robel, R. J., Hoover, R., and Fraumeni, J. F., Jr. (1986). Agricultural herbicide use and risk ofIymphoma and soft-tissue sarcoma. 1. Am. Med. Assoc. 256, 1141-1147. Institute of Medicine (1988). "Role of the Primary Care Physician in Occupational and Environmental Medicine." Institute of Medicine, National Academy Press, Washington, DC. Jeyaratnam, J., de Alwis Seneviratne, R. S., and Copplestone, J. F. (1982). Survey of pesticide poisoning in Sri Lanka. Bull. World Health Org. 60, 615-619. Jeyaratnam, J., Lun, K. c., and Phoon, W. O. (1987). Survey of acute pesticide poisoning among agricultural workers in four Asian countries. Bull. World Health Org. 65, 521-527. Jones, J., and Hunter, D. (1995). Consensus methods for medical and health services research. Br. Med. 1. 311, 376-380. Keefe, T. J., Savage, E. P., Munn, S., and Wheeler, H. W (1985). "Evaluation of Epidemiologic Factors from Two National Studies of Hospitalized Pesticide Poisonings, USA" Colorado State University, Fort CoIIins. Keefe, T. J., Savage, E. P., and Wheeler, H. W (1990). "Third National Study of Hospitalized Pesticide Poisonings in the United States, 1977-1982." Colorado State University, Fort Collins. Keifer, M., McConnell, R., Pacheco, A. F., Daniel, W., and Rosenstock, L. (1996). Estimating underreported pesticide poisonings in Nicaragua. Am. 1. Ind. Med. 30, 195-201. Kimani, V. W., and MacDermott, J. (1998). Problem in initiating a programme on health surveillance of factory and farm pesticide workers. A Kenyan experience. In "International Conference on Pesticide Use in Developing Countries: Impact on Health and Environment" February 23-28, 1998, San Jose, Costa Rica," Book of Abstracts, p. 216. Kleinbaum, D. G., Kupper, L. L., and Morgenstern, H. (1982). "Epidemiologic Research." Van Nostrand-Reinhold, New York. Lamminpaa, A, and Riihimiiki, V. (1992). Pesticide-related incidents treated in Finnish hospitals-A review of cases registered over a 5-year period. Hum. Exp. Toxicol. 11,473-479. Leveridge, Y. R. (1998). Pesticide poisoning in Costa Rica during 1996. Vet. Hum. Toxico!. 40, 42-44. Levine, M., Waiter, S., Lee, H., Haines, T., Holbrook, A, and Moyer, V. (1994). Users' guides to the medical literature. IV. How to use an article about harm. 1. Am. Med. Assoc. 271, 1615-1619. Levine, K S., and Doull, J. (1992). Global estimates of acute pesticide morbidity and mortality. Rev. Environ. Contam. Toxico!. 129,29-50. Litovitz, T. (1998). The TESS database: Use in product safety assessment. Drug Sa! 18,9-19. Litovitz, T. L., Clark, L. K, and Soloway, R. A (1994). 1993 Annual report of the American Association of Poison Control Centers Toxic Exposure Surveillance System. Am. 1. Emerg. Med. 12,546-584. Litovitz, T. L., Felberg, L., Soloway, K A., Ford, M., and Geller, K (1995). 1994 Annual report of the American Association of Poison Control Centers Toxic Exposure Surveillance System. Am. 1. Emerg. Med. 13,551-597. Litovitz, T. L., Felberg, L., White, S., and Klein-Schwartz, W (1996). 1995 Annual report of the American Association of Poison Control Centers Toxic Exposure Surveillance System. Am. 1. Emerg. Med. 14,487-537.
Litovitz, T. L., Smilkstein, M., Felberg, L., Klein-Schwartz, W., Berlin, R .. and Morgan, J. L. (1997). 1996 Annual report of the American Association of Poison Control Centers Toxic Exposure Surveillance System. Am. 1. Emerg. Med. 15,447-500. London, L., Ehrlich, K 1., Rafudien, S., Krige, F., and Vurgarellis, P. (1994). Notification of pesticide poisoning in the western Cape, 1987-1991. S. Afr. Med. 1. 84, 269-272. Mausner, J. S., and Kramer, S. (1985). "Mausner and Bahn Epidemiology-An Introductory Text." Saunders, Philadelphia. McConnell, R., and Hruska, A. J. (1993). An epidemic of pesticide poisoning in Nicaragua: Implications for prevention in developing countries. Am. Public Health 83,1559-1562. Meriwether, R. A. (1996). Blueprint for a national public health surveillance system for the 21st century. 1. Public Health Management Practice 2, 1623. National Agricultural Statistics Service (NASS) (1998). "Farm Labor" [Online]. Available at http://www.usda.gov/nass/. National Center for Environmental Health (NCEH) (1996). "NCEH Activities during Lorain County Methyl Parathion Decontamination Project." Final Report to ATSDR, National Center for Environmental Health, Centers for Disease Control and Prevention, Atlanta. New York State Department of Health (1995). "New York Pesticide Poisoning Registry Report: 1992, 1993, and 1994." Bureau of Occupational Health, New York State Department of Health, Albany. New York State Department of Health (1997). "New York Pesticide Poisoning Registry Report: 1995 and 1996." Bureau of Occupational Health, New York State Department of Health, Albany. O'MaUey, M. A., and McCurdy, S. A. (1990). Subacute poisoning with phosalone, an organophosphate insecticide. Western 1. Med. 153,619-624. Persson, H., Palmborg, M., Irestedt, B., and Westberg, U. (1997). Pesticide poisoning in Sweden-Actual situation and changes over a 10 year period. Przeglad Lekarski 54, 657-661. Pesticide Analytical and Response Center (PARC) (1991-1992). "Annual Report: Pesticide Analytical and Response Center." Health Division, State of Oregon, Portland. Pesticide Analytical and Response Center (PARC) (1993). "Annual Report: Pesticide Analytical and Response Center." Health Division, State of Oregon, Portland. Pesticide Analytical and Response Center (PARC) (1994). "Annual Report: Pesticide Analytical and Response Center." Health Division, State of Oregon, Portland. Pesticide Analytical and Response Center (PARC) (1995). "Annual Report: Pesticide Analytical and Response Center." Health Division, State of Oregon, Portland. Pesticide Analytical and Response Center (PARC) (1996). "Annual Report: Pesticide Analytical and Response Center." Health Division, State of Oregon, Portland. Pesticide Analytical and Response Center (PARC) (1997). "Annual Report: Pesticide Analytical and Response Center." Health Division, State of Oregon, Portland. Pope, A M., and RaIl, D. P. (eds.) (1995). "Environmental Medicine: Integrating a Missing Element into Medical Education." Committee on Curriculum Development in Environmental Medicine, Institute of Medicine, National Academy Press, Washington, DC. Rosenstock, L., Daniell, W., Bamhart, S., Schwartz, D., and Demers, P. A (1990). Chronic neuropsychological sequelae of occupational exposure to organophosphate insecticides. Am. 1. Ind. Med. 18,321-325. Rothman, K. J., and Greenland, S. (1998). Causation and causal inference. In "Modem Epidemiology" (K. J. Rothman and S. Greenland, eds.), 2nd ed., pp. 7-28. Lippincott-Raven, Philadelphia. Savage, E. P., Keefe, T. J., Mounce, L. M., Heaton, R. K., Lewis, J. A., and Burcar, P. J. (1988). Chronic neurological sequelae of acute organophosphate pesticide poisoning. Arch. Environ. Health 43, 38-45. Savage, E. P., Keefe, T. J., Wheeler, H. W, and Helwic, L. J. (1980). "National Study of Hospitalized Pesticide Poisonings, 1974--1976." EPA Publication 540/9-80-001, U.S. Environmental Protection Agency, Washington, DC.
References
Steenland, K., Jenkins, B., Ames, R. G., O'Malley, M., Chrislip, D., and Russo, J. (1994). Chronic neurological sequelae to organophosphate pesticide poisoning. Am. J. Public Health 84, 73 I -736. Thacker, S. B., and Berkelman, R. L. (1988). Public health surveillance in the Vnited States. Epidemio!. Rev. 10, 164-190. Thompson, J. P., Casey, P. B., and Vale, J. A. (1995). Deaths from pesticide poisoning in England and Wales 1990-1991. Hum. Exp. Taxieo!. 14,437445. Trape, A., and Zambrone, F. (1991). Epidemiological surveillance program on a population exposed to pesticides in Sao Paulo, Brazil [abstract]. Arch. Environ. Health 46, 123. V.S. Bureau of the Census (1995). "Statistical Abstract of the Vnited States: 1995," I 15th edn. V.S. Govt. Printing Office, Washington, De. V.S. Department of Agriculture (VSDA) (1998). "Agricultural Chemical Vsage: 1997 Restricted Vse Pesticides Summary Reports." V.S. Department of Agriculture, Washington, DC. Available at http://jan.mannlib.comell.edu/reports/nassr/other/pcu-bb/. V.S. Environmental Protection Agency (EPA) (1997a). "Interim Guidance on Maximizing Insurers' Contributions to Responses at Residences Contaminated with Methyl Parathion." Memorandum from Barry Breen, Director, Office of Site Remediation Enforcement. V.S. Environmental Protection Agency, Washington, De. V.S. Environmental Protection Agency (EPA) (1997b). "Illegal Indoor Vse of Methyl Parathion." Office of Pesticide Programs, V.S. Environmental Protection Agency, Washington, De. Available at http://www.epa.gov/oppOOOOl/citizens/methyl.htm. V.S. Environmental Protection Agency (EPA) (1998). "Pesticides and National Strategies for Health Care Providers: Workshop Proceedings, April 23-24, 1998." EPA 735-R-98-001, V.S. Environmental Protection Agency, Washington, DC. V.S. General Accounting Office (GAO) (1993). "Pesticides on Farms: Limited Capability Exists to Monitor Occupational Illnesses and Injuries." GAOIPEMD-94-6. V.S. General Accounting Office, Washington, De. Veltri, J. e., McElwee, N. E., and Schumacher, M. e. (1987). Interpretation and uses of data collected in poison control centers in the Vnited States. Med. Toxieol. Adverse Drug Exp. 2, 389-397. Vlachos, P., Zeis, P. M., Poulos, L., and Papadatos, e. (1982). Agricultural poisons and children. Pediatrieian 11, 197-204. Vorhaus, L. J., and Kark, R. M. (1953). Serum cholinesterase in health and disease. Am. J. Med. 14,707-719. Waksberg, J. (1978). Sampling methods for random digit dialing. J. Am. Stat. Assoe. 73, 40-46. Washington State Department of Agriculture (1994). "Rules Restricting the Vse ofPhosdrin Finalized." Press Release April 18, 1994, Washington State Department of Agriculture, Olympia.
641
Washington State Department of Health (1994). "Annual Report 1993: Pesticide Incident Reporting and Tracking Review Panel." Washington State Department of Health, Olympia. Washington State Department of Health (1995). "Annual Report 1994: Pesticide Incident Reporting and Tracking Review Panel." Washington State Department of Health, Olympia. Washington State Department of Health (1996). "Annual Report 1995: Pesticide Incident Reporting and Tracking Review Panel." Washington State Department of Health, Olympia. Washington State Department of Health (1997). "Annual Report 1996: Pesticide Incident Reporting and Tracking Review Panel." Washington State Department of Health, Olympia. Washington State Department of Health (1998). "Annual Report 1997: Pesticide Incident Reporting and Tracking Review Panel." Washington State Department of Health, Olympia. Weinbaum, Z., Schenker, M. B., Gold, E. B., Samuels, S. J., and O'Malley, M. A. (1997). Risk factors for systemic illnesses following agricultural exposures to restricted organophosphates in California, 1984-1988. Am. J. Ind. Med. 31, 572-579. Weldon, M., Methner, M. M., Willis, T., and Salzman, D. (1996). Acute pesticide poisoning associated with use of a sulfotepp fumigant in a greenhouse. App!. Oeeup. Enviran. Hyg. 11, !l05-1107. Wesseling, e., Castillo, L., and Elinder, e. (1993). Pesticide poisonings in Costa Rica. Seand. 1. Work Environ. Health 19, 227-235. Whitmore, R. W, Kelly, J. E., and Reading, P. L. (1992). "National Home and Garden Pesticide Vse Survey Final Report." RTII5100/17-01F, Research Triangle Institute, Research Triangle Park, Ne. Wiklund, K., Dich, J., Holm, L. E., and Eklund, G. (1989). Risk of cancer in pesticide applicators in Swedish agriculture. Br. J. Ind. Med. 46, 809-814. Wilson, B. W, Padilla, S., Henderson, J. D., Brimijoin, S." Dass, P. D., EIIiot, G., Jaeger, B., Lanz, D., Pearson, R., and Spies, R. (1996). Factors in standardizing automated cholinesterase assays. J. Toxieol. Environ. Health 48, 187-195. Wilson, B. W, Sanborn, J. R., O'Malley, M. A., Henderson, J. D., and BiIIitti, J. R. (1997). Monitoring the pesticide exposed worker. Oeeup. Med. 12, 347-363. Wong, 0., Morgan, R. W., Whorton, M. D., Gordon, N., and Kheifets, L. (1989). Ecological analyses and case-control studies of gastric cancer and leukaemia in relation to DBCP in drinking water in Fresno County, California. Br. J. Ind. Med. 46,521-528. World Health Organization (WHO) (1977). "International Classification of Diseases," 1975 revision. WHO, Geneva. Zeitz, P. A., MacDonald, S. e., and Yoon, S. S. (1998). "1997 CSTE-CDCASPH Survey of Statewide Surveillance Systems of Sentinel Environmental Diseases: Status and Trends." Council of State and Territorial Epidemiologists, Atlanta.
CHAPTER
28 Environmental Transport and Fate James N. Seiber Western Regional Research Center USDA-ARS
The approaches taken from 1970 to 2000 toward understanding the principles of pesticide transport and fate in the environment led to development of a prospective, predictive capability for evaluating environmental behavior before widespread use or release. Progress has been made in defining and understanding dissipation pathways, the relationship between physicochemical properties and dissipation, structure-activity relationships, and environmental activation and deactivation. Improvements in analytical methodology, which provide much lower detection limits for following the fate of breakdown products as well as parent chemicals, have been central to the development of a principle-based approach to pesticide processing in the environment.
28.1 INTRODUCTION Assessing the transport and fate of pesticides in the environment is complicated. There are many environmental pathways available at the local, regional, and global levels. Pesticides vary greatly in physical and chemical properties and use patterns, plus the environment itself is complex and varies from one location to another and from one time to another. It is a goal of environmental sciences to understand and deal with the complexities in nature by defining and sorting out underlying principles. These can serve as a basis for developing an assessment of chemical processing and its relationship to the health of the environment. In the past, including the first decades after the widespread introduction of synthetic organic chemicals for pest control, knowledge of environmental behavior and fate was determined by analysis for these chemicals in environmental samples after they had been used/released for many years. By analyzing soil, water, sediment, air, plants, and animals, environmental scientists were able to piece together profiles of each chemical's environmental behavior. Dibromochloropropane, ethylene dibromide, and chemicals with similar uses as soil nematicides and similar physicalJchemical properties were recognized for their potential to contaminate groundwater in general use areas. DDT and other chlorinated insecticides and organic comHandbook of Pesticide Toxicology Volume 1. Principles
pounds of similar low polarity and water solubility and high stability threatened some aquatic and terresterial organisms because of their potential for undergoing bioaccumulation and their chronic toxicities. Like the chlorofluorocarbons, methyl bromide was found to be quite stable in the atmosphere and able to diffuse to the stratosphere, where it entered into reaction sequences contributing to the thinning of the ozone layer. However, the retrospective approach is fraught with difficulty:
1. Adverse chemical behavior might be discovered too late, after considerable environmental damage was already done. An example is the decline of raptorial bird species after widespread use of DDT or substantial loss of stratospheric ozone from long-term use and release of chlorofluorocarbons and methyl bromide. 2. By analyzing for the wrong chemical or the wrong target media, the problem may be misdefined or completely overlooked. For example, parent pesticides such as aldicarb or aldrin appear to have low persistence in the environment, but they can be converted into breakdown products (aldicarb sulfoxide and sulfone; dieldrin and, eventually, photodieldrin), which may be the primary offenders. Targeting only the parents rather than the products in the analysis scheme may overlook the more hazardous products. The trend from roughly the 1970s to the present has been to develop an understanding of the underlying principles of environmental fate to find ways to predict environmental behavior before the chemical is released. For economic materials such as pesticides, premarket environmental fate/effects testing is now built into the requirements for regulatory approval. The Environmental Fate Guidelines of the U.S. Environmental Protection Agency (EPA) (Kovacs, 1983; U.S. EPA, 1982) specify the tests and acceptable behavior required for registration of candidate pesticides in the United States. Similar guidelines and test protocols exist in Europe (Thomas, 1991), Canada (Agriculture Canada, 1987), Australia (Holland, 1999), and other nations and economic organizations.
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Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
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Environmental Transport and Fate
Another stimulus for better analytical and predictive tests was the development of risk assessment tools for evaluating risks of chemicals in the environment. Risk assessment and risk science in general are relatively new to the evaluation of health impacts of chemicals, dating from the late 1970s and early 1980s for human health risk assessment (National Research Council, 1983) and even more recently for ecological risk assessment (Suter et aI., 1993). For use in both the hazard identification component, which includes measuring/estimating emissions to the environment, and particularly the exposure assessment component of risk assessment, which involves measuring or modeling exposures via food, water, air, etc., predictive tools (models) are undergoing rapid development for use in regulatory actions, both for premarket screening and for reaching decisions on continuing use. Many commercial pesticides, as well as hazardous air pollutants (National Research Council, 1994) and other chemicals of environmental concern, have undergone or are now in the process of undergoing risk assessment review (National Research Council, 1993). Much of the spotlight has been focused on chemicals already in use or present in the environment to guide societal decisions on what steps might need to be taken to reduce adverse impacts. However, the methodology can also be applied to simulate hazards and risks for new candidate pesticides. Both the regulatory agencies and industry have played important roles in encouraging the development of predictive methods. It is clearly in the best interests of companies to screen out potential environmental problems early in the development process and to focus resources on chemicals with prospects for long-term environmental compatibility. For example, environmental scientists at Dow Chemical in the early 1970s developed a "benchmark approach" for evaluating the environmental characteristics of candidate pesticides (Goring and Hamaker, 1972). The benchmark approach and other early developments in screening/predicting environmental behavior including modeling became formalized in the field of environmental chemodynamics, which may be generally defined as "the subject dealing with the transport of chemicals (intraand interphase) in the environment, the relationship of their physical-chemical properties to transport, their persistence in the biosphere, their partitioning in the biota, and toxicological and epidemiological forecasting based on physicochemical properties" (Haque and Freed, 1974; Thibodeaux, 1979). Another factor in developing a predictive capability for environmental behavior and fate is the shift in the types of pesticide chemicals over the past 30 years or so. The highly stable, lipophilic organochlorines, organophosphates of relatively high mammalian toxicity, and environmentally persistent triazine and phenoxy herbicides that dominated pesticide chemistry to the 1970s either are gone entirely from pesticide markets or being replaced. In their place are synthetic pyrethroids, sulfonylureas, aminophosphonic acid derivatives, biopesticides, and many other classes and types whose environmental fate and ecotoxicological effects are less straightforward. Some of the new pesticides are attractive because they degrade relatively rapidly and extensively in the environment, but this can
multiply the number of discrete chemicals that need to be evaluated. Relying solely on experimentation under environmental conditions could significantly slow regulatory approval, arguing again for using predictive screening assessment tools as an integral component of the overall research and development effort. Pressure is increasing to develop tests for subtle environmental effects that go beyond the presistence, leaching, bioaccumulation, and acute/chronic toxicity testing prominent in environmental fate tests in the past. A current concern is environmental endocrine disruption effects caused by trace levels of chemicals and chemical mixtures (Jobling et aI., 1995; Inter-Organization Programme for the Sound Management of Chemicals, 1997; National Research Council, 1999). Ideally, environmental chemists would be able to detect interactions of endocrine disrupting chemicals (EDCs) with mammalian tissues and ecosystems by biobased testing for the chemicals themselves or biomarkers which indicate that exposures to EDCs had occurred. The methods of and approaches to screening for EDCs, under intense development from the stimulus of the Food Quality Protection Act (FQPA, 1996; Johnson and Bailey, 1999), have the potential for adding significantly to the already complex business of premarket predictive testing. Much of our current capability for determining transport and fate of pesticides and other chemicals may be traced to the tremendous developments in analytical chemistry from the 1950s and 1960s to the present. Detection limits oflow parts per billion (ppb) and even parts per trillion (ppt) are now achievable by better methods of extracting, preparing, and, particularly, determining residues of pesticides and breakdown products in a variety of matrices (e.g., Fong et aI., 1999). Developments in gas and liquid chromatography, mass spectrometry, and immunoassay have been among those most useful to environmental scientists, but computer data handling capabilities have also enabled the routine use of these sophisticated techniques in both research and regulatory laboratories. Biosensor-based methods provide promise for superior detectability in miniaturized equipment that can be used in the field.
28.2 PRINCIPLES 28.2.1 THE DISSIPATION PROCESS
Once a substrate (agricultural commodity, body of water, wildlife, soil, etc.) has been exposed to a chemical intentionally or accidentally, dissipation processes begin immediately. The initial residue dissipates at an overall rate that is a composite of the rates of volatilization, washing off, leaching, hydrolysis, microbial degradation, and other individual processes (Seiber, 1985). Concentrations of the total residue (parent chemical plus breakdown products) typically decrease with time after exposure or treatment ends. For chemicals that are converted to products which are more stable than the parent, product formation can slow the overall dissipation process, well exemplified
28.2 Principles
645
0.00 2.0 ·0.5
le.,
... -=C
1.5
t
•
=·0.23 day·1 112 =3.0 days
c:l
C3 -1.0
.!
cS 1.0
0.5
0L.._ _...._ _..&._ _....I1o..._
o
5
10
15
Time, Days (a)
o
5
10
Time, Days (b)
Figure 28.1 Dissipation rate of molinate from a rice field at 26°C as (a) a dissipation curve and (b) as a first order plot. Co is the initial concentration and C is the concentration at time t (see Soderquist et aI., 1977 for original data; Crosby, 1998).
by significant concentrations of DDD and DDE in field soils exposed to DDT many years previously. When low-level exposure results in the accumulation of residues over time, as in the case of bioconcentration of residues from water by aquatic organisms, the overall environmental process includes both the accumulation and dissipation phases. However for simple dissipation, such as that occurring in the application of pesticides and resulting exposure from residues in food or water or air, the typical results are that concentrations of overall residue (parents plus products) decrease with time after exposure of treatment (Fig. 28.1). Because most individual dissipation processes follow apparent first-order kinetics, overall dissipation or decline is also often observed to be first-order. Because first-order decline processes are logarithmic, that is, a plot of remaining residue concentration versus time is asymptotic with respect to the time axis, residues will approach zero with time but never cease to exist entirely (Fig. 28.1a). Thus all environmental exposures lead to residues which have, theoretically, unlimited residue longevity. But our ability to detect those residues is limited, by the operational abilities of analytical methods of gas chromatography, high-performance liquid chromatography, mass spectrometry, immunoassay, and other approaches. The goal is to have sufficient detectability in the methods used to be able to follow residues to the point where their amounts are well below any plausible threshold for adverse biological effects. This presents an inherent dilemma, because biological effects testing is subject to continuing reevaluation (e.g., with environmental endocrine disruptors). Thus, more sensitive analytical techniques are constantly needed so that dissipation processes
can be followed to lower concentration levels, with more chemical breakdown product detail. 28.2.2 ENVIRONMENTAL COMPARTMENTS
Once a pesticide gains entry to the environment, by intentional or accidental release, it may enter one or more compartments, illustrated in Fig. 28.2. The initial compartment contacted by the bulk of the pesticide will be determined largely by the manner of use or release. Pesticides applied to flooded rice fields, for example, enter the aquatic compartment initially. In time, however, residues will tend to redistribute and favor one or more compartments or media over others, in accord with the chemicals' physical properties, reactivity, and stability characteristics and the availability and composition of compartments in the general environment where the use or release has occurred (Biggar and Seiber, 1987). Figure 28.2 tabulates the compartments, the transfer/transformation process, and the environmental characteristics that are involved in transport and fate. The nature of the chemical of interest will dictate the pathways to be favored, so that environmental dissipation and fate must be evaluated on a chemical by chemical basis, as well as an environment-specific basis. This is illustrated in Fig. 28.3 for chemical behavior in a pond environment, for which the properties of the chemical of interest must be taken into account along with, and as influenced by, the properties of the pond environment. Analogous schematics have been developed for chemical behavior in soil (Cheng, 1990) as well as other environments, including plants and animals.
646
CHAPTER 28
Environmental Transport and Fate
ENVIRONMENTAL COMPARTMENTS
CHE<)
Air
Volatilization
Oxidation
Soil
Sorption
Reduction
Water
Diffusion
Hydrolysis
Biota
Partitioning
Conjugation
~ Figure 28.2
TRANSFORMATION PROCESSES
TRANSFER PROCESSES
..}Q#'
$
-#
FATE
~ ##'
~
A schematic diagram of the components of the fate of a chemical in the environment (Seiber, 1985). PROPERTIES PHYSICAL CHEMICAL
MW
khyd kphoto koxidn
H
BCF Sol.
kmetab kmicro
V.P.
Kd Air
Temp pH Susp. Sed.
Water
Windspeed Sunlight Intensity Humidity
Biomass Dissolved 02
Sediment Figure 28.3 Intrinsic and extrinsic properties governing the distribution and fate of a chemical in a pond environment (Seiber, 1987).
Some chemicals inherently favor and thus will migrate when the opportunity arises to water. These are primarily chemicals of high water solubility and high stability in water, such as salts of carboxylic acid herbicides [2,4-dichlorophenoxyacetic acid (2,4-D), 4-chloro-2-methylphenoxyacetic acid, and trichloroacetic acid)]. Others favor the soil or sediment compartment because they are preferentially sorbed to soil and they may lack other characteristics (volatility or water solubility) which favor removal from soil. Examples include paraquat, which is strongly sorbed to the clay mineral fraction of soil, and DDT, toxaphene, and the cyclodiene pesticides, which sorb to and are stabilized in soil organic matter. Others, particularly fat-soluble substances, favor storage in fatty animal tissue when the opportunity arises. Volatile chemicals such as the fumigants methyl bromide and telone (1 ,3-dichloropropene) and chemicals with high Henry's law constants favor the air compartment. The elements of environmental fate prediction, based on properties of the chemical of interest that may be measured or estimated based upon structures, become apparent through these wellestablished "benchmark" chemicals (Boethling and Mackey, 2000; Lyman et aI., 1982; Mackay et aI., 1992).
28.2.3 STRUCTURE The key to how a chemical will behave is contained in the chemical's molecular makeup or structure. The field of structure-
activity relationships is an important tool in pesticide environmental chemistry that is still undergoing development. An example of the importance of even small structural changes is provided by contrasting the behavior of the two closely related chemicals, DDT and dicofol (Table 28.1). The subtle structural change, by substituting OH for H at the central carbon, has major environmental fate ramifications (as well as a strong influence on biological activity). DDT degrades slowly in the environment, and its primary breakdown products DDE and DDD are also very stable. Dicofol degrades rather rapidly in the environment, and its principal breakdown product, dichlorobenzophenone (DCBP), is also degraded further rather rapidly. DDT (and DDEIDDD) is highly lipophilic, showing strong tendencies to bioconcentrate in aquatic organisms and, through food chain accumulation, in terrestrial animals and man. Dicofol has lower lipophi1icity and more ready breakdown, both owing to the presence of the hydroxy substituents, and does not significantly bioconcentrate or bioaccumulate. Its primary breakdown products do not exhibit these negative characteristics either. Even though there has been much experience with both DDT and dicofo1, new information continues to surface for the parent chemicals as well as for the degradation products. Because of these differences in toxicity and environmental behavior, DDT was banned for use in the United States in the 1970s and dicofol is still registered for use. Thus the impor-
28.3 Summary
64"
Table 28.1 Influence of Structure on Biological Activity, Environmental Behavior, and Regulatory Status ofDDT and Dicofol CCh I C6H4 CI-C-C6H4Cl I H DDT
CCh I C6H.Cl -C-C6H4 Cl I OH Dicofol
Activity as pesticide
Insecticide
Acaricide
Environmental reactivity
Stable; breakdown products
Breaks down; primary
(DDE and DDD) also stable
breakdown product
Bioconcentration potential
High in aquatic and terrestrial
Low
Regulatory status (U.S.)
Banned
(DCBP) also unstable food chains
tance of subtle structural features cannot be overemphasized for closely related structures such as DDT and dicofol and certainly so for more structurally diverse chemicals. If methylchlor and/or methiochlor, which are good insecticides but biodegrade rapidly in the environment (Metcalf, 1977), had been developed rather than DDT, we might still be using "DDT-1ike" insecticides in U.S. agriculture today. 28.2.4 ACTIVATION-DEACTIVATION
Environmental transformations generally lead to products that are less of a threat to biota and the environment than the parent chemicals; that is, they result in deactivation of the parent. The products may be less toxic than the parent or have lower mobility and persistence relative to the parent. They may, in short, be simply transient intermediates on the path to complete breakdown or "mineralization" of the parent. Thus, 2,4-D may degrade to oxalic acid and 2,4-dichlorophenol. The latter is of some concern, but it lacks the herbicidal activity of 2,4-D and appears to be further degraded in most environments by sunlight, microbes, etc. Organophosphates can be hydrolyzed in the environment to phosphoric or thiophosphoric acid derivatives and a substituted phenol or alcohol. These products, in the case of most organophosphates, are much less of a threat to man and the environment than the parent chemicals. Environmental activation represents the relative minority of transformations, which lead to products that are more of a threat due to one or more of the following characteristics: • Enhanced toxicity to target and/or nontarget organisms. • Enhanced stability, leading to greater persistence. • Enhanced mobility, leading to greater potential for contamination of ground water or other sensitive environmental media. • Enhanced lipophilicity, leading to bioconcentration and bioaccumulation. Examples of activations (Coats, 1991; Wolfe and Seiber, 1993) include the formation of DDE, which is apparently the
S till registered
form most responsible for causing thin egg shells in birds that have accumulated DDT or DDE from their prey, and DDD, which can persist for years in some soil and water systems; formation of dieldrin and eventually photodieldrin from aldrin; oxidation of organophosphate thions to the more toxic "oxon" form; S-oxidation of aldicarb (and some other N-methylcarbamates) to the more water soluble and, in some cases, more persistent sulfoxide and sulfone forms; formation of the volatile fumigant methylisothiocyanate (MITC) from metam sodium, the commercial precursor of MITC, when the parent is applied to moist soil; and formation of ethylenethiourea, a carcinogen, from ethylenebisdithiocarbamate fungicides. In part because of the concern over environmental activation, the U.S. EPA requires extensive information on the occurrence, toxicity, and fate of transformation products of candidate pesticides submitted for registration (Kovacs, 1983). The tests must include significant products of hydrolysis, photolysis, oxidation, and microbial metabolism, in both laboratory and field tests, but increasingly, regulations are also geared to products that might be formed during illegal use or during fires, explosions, spills, disinfection, and other situations that expose chemicals to conditions for which they were not intended (Bourke et aI., 1992). Unfortunately, not all such situations can be anticipated, requiring continual vigilance by the registrant and regulatory agencies as a part of product stewardship and environmental protection.
28.3 SUMMARY Significant advances have been made in defining and understanding the principles that underlie pesticide dissipation in the environment. This has led to better testing procedures and better environmental fate models, which in turn has given rise to methods for predicting environmental behavior or fate before use or release occurs. The development of a predictive capability has helped focus efforts in industry and agencies toward marketing and regulating safer chemicals and to restrict or elim-
648
CHAPTER 28
Environmental Transport and Fate
inate those that are likely to pose significant risks. As a result of this and parallel developments in understanding the mechanisms of toxic action, pesticides and pest control practices at the beginning of the 21st century are considerably safer than in the past 60 years of the synthetic pesticide era. These accomplishments provide greater protection to users of pesticides, to nearby residents, to consumers, and to wildlife.
REFERENCES Agriculture Canada, Environment Canada, Department of Fisheries and Oceans. (1987). "Environmental Chemistry and Fate Guidelines of Pesticides in Canada." Ottawa, July 15. Biggar, J. w., and Seiber, J. N. (eds., and Technical Coordinators). (1987). "Fate of Pesticides in the Environment, Proceedings of a Technical Seminar." Publication no. 3320, University of California, Division of Agriculture and Natural Resources. Boethling, R. S., and Mackey, D. (2000). "Handbook of Property Estimation Methods for Chemicals." Lewis, Boca Raton, FL. Bourke, J. B., Felsot, A. S., Gilding, T. J., Jensen, J. K., and Seiber, J. N. (eds.) (1992). "Pesticide Waste Management: Technology and Regulation," ACS Symposium Series, Vol. 510. Am. Chem. Soc., Washington, DC. Cheng, H. H. (1990). "Pesticides in the Soil Environment: Processes, Impacts, and Modeling," Book Series no. 2. Soil Sci. Soc. Am., Madison, WI. Coats, J. R. (1991). Pesticide degradation mechanisms and environmental activation. In "Pesticide Transformation Products: Fate and Significance in the Environment" (L. Somasundaram and J. R. Coats, eds.), ACS Symposium Series, Vol. 459, pp. 10-31. Am. Chem. Soc., Washington, DC. Crosby, D. G. (1998). "Environmental Toxicology and Chemistry." Oxford University Press, New York. Fong, W. G., Moye, H. A., Seiber, J. N., and Toth, J. P. (1999). "Pesticide Residues in Foods: Methods, Techniques, and Regulations." Wiley, New York. Food Quality Protection Act. (1996). U.S. Congress, Washington, DC. Goring, c.1., and Hamaker, J. N. (1972). "Organic Chemicals in the Soil Environment," Vols. I and 2. Dekker, New York. Haque, R., and Freed, V. H. (eds.). (1974). "Environmental Dynamics of Pesticides." Plenum, New York. Holland, J. (1999). Environmental fate: A down under perspective. In "Pesticide Chemistry and Bioscience. The Food-Environment Challenge" (G. T. Brooks and T. R. Roberts, eds.). Royal Society of Chemistry, Cambridge, UK. Inter-Organization Programme for the Sound Management of Chemicals. (1997). "International Workshop on Endocrine Disruptors." Report, UNEP Chemicals, Geneva. Jobling, S., Reynolds, T., White, R., Parker, M. G., and Sumpter, J. P. (1995). A variety of environmentally persistent chemicals, including some phthalate
plasticizers, are weakly estrogenic. Environ. Health Perspect. 103, 582587. Johnson, S. L., and Bailey, J. E. (1999). Pesticide risk management and the United States Food Quality Protection Act of 1996. In "Pesticide Chemistry and Bioscience. The Food-Environment Challenge" (G. T. Brooks and T. R. Roberts, eds.). Royal Society of Chemistry, Cambridge UK. Kovacs, Jr., M. E (1983). EPA guidelines on environmental fate. Residue Rev. 85,3-16. Lyman, W. J., Reehl, W. E, and Rosenblatt, D. H. (1982). "Handbook of Chemical Property Estimation Methods." McGraw-Hill, New York. Mackay, D., Shiu, W. Y., and Ma, K. C. (1992). "Illustrated Handbook of Physical-Chemical Properties and Environmental Fate for Organic Chemicals." Vols. I-V. Lewis, Boca Raton, FL. Metcalf, R. (1977). Model ecosystem studies of bioconcentration and biodegradation of pesticides. In "Pesticides in Aquatic Environments" (M. A. Q. Khan, ed.), pp. 127-144. Plenum, New York. National Research Council. (1983). "Risk Assessment in the Federal Government: Managing the Process." National Academy Press, Washington, DC. National Research Council. (1993). "Pesticides in the Diets ofInfants and Children." National Academy Press, Washington, DC. National Research Council. (1994). "Science and Judgment in Risk Assessment." National Academy Press, Washington, DC. National Research Council. (1999). "Hormonally Active Agents in the Environment." National Academy Press, Washington, DC. Seiber, J. N. (1985). General principles governing the fate of chemicals in the environment. In "Agricultural Chemicals of the Future" (J. L. Hilton, ed.), pp. 389-402, Beltsville Symposia in Agricultural Research No. 8. Rowan and Allanheld, Totowa, NJ. Seiber, J. N. (1987). Principles governing environmental mobility and fate. In "Pesticides: Minimizing the Risks" (N. N. Ragsdale and R. J. Kuhr, eds.), ACS Symposium Series, Vol. 336, pp. 88-105. Am. Chem. Soc., Washington,DC. Soderquist, C. J., Bowers, J. B., and Grosby, D. G. (1977). Dissipation of molinate in a rice field. 1. Agr. Food Chem. 25, 940-946. Suter, G. w., Barnthouse, L. w., Bartell, S. M., Mill, T., Mackay, D., and Paterson, S. (1993). "Ecological Risk Assessment." Lewis, Boca Raton, FL. Thibodeaux, L. J. (1979). "Chemodynamics: Environmental Movement of Chemicals in Air, Water, and Soil." Wiley-Interscience, New York. Thomas, B. (1991). Pesticide registration in Europe. In "Regulation of Agrochemicals" (G. J. Marco, R. M. Hollingworth, and J. R. Plimmer, eds.), pp. 73-79. Am. Chem. Soc., Washington, DC. U.S. Environmental Protection Agency (U.S. EPA). (1982). "Pesticide Assessment Guidelines. Subdivision N. Chemistry: Environmental Fate." EPA540/9-82-021, Office of Pesticides and Toxic Substances, Washington, DC. Wolfe, M. E, and Seiber, J. N. (1993). Environmental activation of pesticides. In "De Novo Toxicants: Combustion Toxicology, Mixing Incompatibles, and Environmental Activation of Toxic Agents" (D. J. Shusterman and J. E. Peterson, eds.), Occupation Medicine: State of the Art Reviews, Vol. 8, pp. 561-573. Han1ey and Belfus, Philadelphia.
CHAPTER
29 Hydrophohicity as a Key Physicochemical Parameter of Environmental Toxicology of Pesticides Toshio Fujita EMIL Project
Keiichiro Nishimura Osaka Prefecture University
Chiyozo Takayama Sumitomo Chemical Company
Masanori Yoshida and Matazaemon Uchida Nihon N ohyaku Company
29.1 INTRODUCTION The environmental behaviors of pesticides, such as accumulation in soil, contamination of aquasphere, residue levels in crops, and bioaccumulation through food chains as well as nondietary routes, are dependent on their distribution properties among various environmental phases. These distribution features are modeled by phase-distribution equilibrium constants such as soil absorption coefficient, water solubility, and bioconcentration factors in biota, including crops (Briggs, 1981a). As early as 3 decades ago, the logarithm of these constants was recognized to be related with a reference parameter representing the molecular "hydrophobicity" (Briggs, 1969, 1981a; Hance, 1967; Hansch et aI., 1968; Neely et aI., 1974; Valvani and Yalkowsky, 1980). The most frequently used hydrophobicity parameter is the log P [or log k(ojw)], P [or k(ojw)] being the l-octanoVwater partition coefficient (Fujita et aI., 1964; Leo, 1993; Noble, 1993). Moreover, the log P value of organic compounds has been shown to be the most decisive parameter for their toxicity brought about by nonspecific perturbation of biomembra-
Handbook of Pesticide Toxicology Volume 1. Principles
neous and enzyme systems (Hansch and Leo, 1995). Specific toxicities to target pests are governed by specific mechanisms defined by various physicochemical properties of the molecule, including the hydrophobicity (Hansch and Leo, 1995). Thus, the environmental toxicology of pesticides covering distribution patterns, persistence, and toxicity could be quantitatively analyzable in terms of physicochemical molecular descriptors including hydrophobic, electronic, steric, and others with the use of regression analyses, in which the log P value is regarded as playing a central role (Hansch and Leo, 1995). That is, the QSAR (quantitative structure-activity relationship) procedure initiated and developed by Hansch and co-workers (Hansch and Fujita, 1964, 1995; Hansch and Leo, 1995) could apply to various environmental aspects of pesticides in spite of the tremendous complexity of the processes involved. In this chapter, we review the measurement and estimation procedures of the log P for a wide range of organic compounds and the significance of the log P value in elucidating and predicting the environmental behavior of pesticides in terms of the QSAR.
649
Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
650
CHAPTER 29
Hydrophobicity as a Key Parameter of Environmental Toxicology
29.2 MEASUREMENTS AND EXPERIMENTAL ESTIMATIONS OFlogP There have been a number of experimental procedures either to measure directly or to estimate indirectly the log P value of various types of compounds (Sangster, 1997a). Among them, the shake-flask and slow-stirring procedures as direct methods and the potentiometric method for ionizable compounds and the high-performance liquid chromatographic procedure as an indirect method are described in this section. The shake-flask and liquid chromatographic procedures are now standardized by the Organization for Economic Co-operation and Development (OECD) (1981, 1989) for the assessment of the environmental effects of organic compounds, including pesticides. 29.2.1 DIRECT PARTITIONING (SHAKE-FLASK AND SLOW-STIRRING) METHOD
In the measurements of the P value in the l-octanoVwater system, the ratio of concentrations in the two phases should be calculated after the partitioning equilibrium is established as far as possible (Fujita et aI., 1964). Analytically pure l-octanol is commercially available. Impurities, if any, can be removed by consecutive washings with 2 N H2S04, 2 N NaOH, and water. Sufficiently pure l-octanol is obtained by distillation after desiccation. Distilled water, under decarbonated or buffer conditions if necessary, should be used for the partitioning. The two solvent phases should be mutually saturated before the partitioning (Smith et aI., 1975). In most cases except for highly hydrophobic and hydrophilic solutes, the equilibrated concentration is measured only in the water phase. The concentration in the l-octanol phase is calculated after subtraction of the amount in the water phase from the total. The partitioning experiment is usually repeated several times, varying the amount of solute and volume ratios of l-octanol and water. To minimize the calculation error, the volume ratio should be chosen so that the amount of the solute in the two phases after equilibrium is equal, or nearly so. Depending on its polarity, the test compound is dissolved in either the pretreated octanol or the pretreated water phase. The amount is selected so that the initial concentration is on the order of 10- 3 M or less (to discourage dimer or micelle formation), but sufficient so that the equilibrated concentration in the water phase can be accurately measured directly, whenever possible (i.e., without concentration via reextraction or evaporation). For moderately functionalized and polar compounds, the log P of which is neither too high nor too low, a consistent log P value can sometimes be obtained by 100 inversions of the flask in 5 min or less (Leo et al., 1971). However, if the solute does not have surfactant properties, "vigorous" agitation for 1 h seems to be better, in general, for hydrophobic pesticides. If the agitation is mild, it takes the partition equilibrium a longer time to be achieved.
After the partitioning agitation, the two phases are usually separated with centrifugation at 2000-3000 rpm for 15 min. Note that a shorter time at a higher rpm can still leave the aqueous phase supersaturated with l-octanol. In cases in which the partitioning procedure yields emulsion in both phases (or between two turbid solvent layers even after centrifugation), filtration of the aqueous phase through a minimum amount of "inert filtration aids" under a slightly reduced pressure is effective to make the phases clear. Celite powders (Highflo-super-cel, 0: 7 J.!m, Celite Corp., Lompoc, CA) are conveniently used after being packed tightly with a minimum amount of the turbid octanol on a cotton ball placed on a funnel neck. The adsorption of solutes from the water phase to filtration aids has been shown to be negligible as far as compounds with a log P value between -1 and 2.5 are concerned (Nishikawa, 1989). Although the P values are not affected much by temperature variations (Fujita et aI., 1964; Leo et aI., 1971), it is preferable to keep the temperature within a few degrees of 25°C during the entire operation (Dearden and Bresnen, 1988). Vigorous agitation with a "complete" mixing of the two phases is not applicable to highly hydrophobic compounds with a log P value greater than 4-5 because of a persistent emulsification of the system. The emulsified phases are difficult to clear even with a prolonged centrifugation. The celite filtration would not be favorable in this case because a stronger adsorption of compounds from the aqueous phase is probable. For such compounds, usually with a high molecular weight and only poorly functionalized, the slow-stirring procedure is recommended. In this procedure, the two phases are equilibrated under conditions of slow stirring (about 200 rpm for 1000 ml of the system, including 20-50 ml octanol phase) (Brooke et aI., 1986; de Bruijn and Hermens, 1990; de Bruijn et aI., 1989). It takes much time for the equilibrium to be established because any mixing between the two phases should be avoided as far as possible. For compounds with a log P value lower than 4, the equilibrium occurrs within 1-2 days, but, for compounds of log P > 5, it takes 2-3 days. For p,p' -DDT of which log P value was expected to be higher than 6 (6.29, as estimated from that of methoxychlor, 4.83; Nishimura and Fujita, 1983), 3-4 days of stirring were required with a resultant log P of 6.91 (de Bruijn et aI., 1989). The reproducibility of the procedure is reported to be very good, the standard deviation among five measurements being ±0.03 for p,p' -DDT (de Bruijn et aI., 1989). For highly hydrophobic compounds of log P > 5, the amount of solute in the octanol phase is only negligibly decreased ( <0.1 %) at the equilibrium from the total in the initial state with a volume ratio of 1 : 100 (octanoVH20). If only the water phase is analyzed, it is very important to measure the very low concentration with a high accuracy. For nonvolatile solutes, this could be done by extracting the solute from the water phase with hexane followed by evaporation to give a concentration range in which the solute can be analyzed with a sufficient precision. For highly hydrophilic compounds with log P < -5, the reverse situation occurrs. A very low concentration in the octanol phase must be measured accurately. Condensation of the
29.2 Measurements and Experimental Estimations of log P
octanol phase under a reduced pressure and an appropriate dilution with methanol or acetone-hexane (1 : 1, v/v) may make it possible to analyze the amount of solute fairly precisely. An example for the highly hydrophilic glyphosate is published recently (Chamberlain et aI., 1996). To measure the log P value of ionizable compounds, 0.10.01 N HCl and NaOH solutions are used as the aqueous phase for acidic and basic solutes, respectively. Buffer solutions can also be used under conditions in which their buffer capacity should be enough to make the solutes in the aqueous phase exist entirely as the nonionized neutral form. The buffer should not be extractable into the octanol phase so that phosphate buffers are best recommended (Wang and Lien, 1980). The log P value of organic ion pairs, in which either or both the cationic and the anionic species are organic, is sensitive to variations in the ionic strength of the aqueous phase (Takayama et aI., 1985; Terada et aI., 1981). The partition behavior of multiprotic compounds such as peptides, including insect neurohormones (Menn and Borkovec, 1989), is also modified by (inorganic) counterions as well as ionic strength in the aqueous phase (Akamatsu et aI., 1989). Therefore, experimental conditions for the measurement of the "true" log P value of noncharged species of ionizable pesticides should be carefully controlled. It should be mentioned that charged molecular species or ions have no unique hydrophobic index in terms of log P. It varies depending on the counterionic species as well as ionic strength in the aqueous phase. For gaseous compounds used as fumigants and spraypropellants, as well as environmental contaminants such as halomethanes, fluoroalkanes, and nitrogen oxides, specially designed partitioning apparata and gas chromatographic systems are required to measure the accurate log P value (Hansch et aI., 1975). The concentration or the amount of solute in the water phase and, if necessary, also in the octanol phase is analyzed by the appropriate spectrometric, chromatographic, radiometric, and gravimetric (Masutani et aI., 1981) procedures. This analysis is performed directly after centrifugation or with an appropriate combination of extraction, condensation, and solvent evaporation, depending on the situation. The ultraviolet spectrometric analysis is most widely used, particularly for aromatic compounds. From the spectra, it is possible to confirm whether or not the structure of the solute molecule is transformed during the partitioning process (Fujita et aI., 1964). The gas (Sotomatsu et aI., 1987) and high-performance liquid chromatographic analyses (Akamatsu et aI., 1989) are also convenient for aliphatic as well as volatile aromatic compounds. In these cases, the confirmation can be made by comparison of retention time before and after partitioning. It is the most important prerequisite in the log P measurement that no structural change occurs during the partitioning procedure. Sometimes, the procedure should be done under nitrogen atmosphere or with an addition of a minimum amount of sodium thiosulfate to the aqueous phase to prevent air oxidation (Fujita et al., 1964). Radiometric analysis are sometimes used for solutes at very low concentration in the aqueous or octanol phase, that is,
651
highly hydrophobic in the water (Uchida et aI., 1974) or highly hydrophilic in the octanol (Chamberlain et aI., 1996). Serious error can occur in this procedure if radiocolloids adsorb on the inner wall of the vessels. This is in addition to any errors introduced from lack of purity of the labeled compounds. To reduce the effect of the radiocolloid adsorption, an appropriate amount of the corresponding nonlabeled compounds (carriers) should be added to the system to lower the radiocolloid formation by the chemical dilution of the radioactivity (Keller, 1993). 29.2.2 POTENTIOMETRIC TITRATION METHOD FOR IONIZABLE PESTICIDES A potentiometric titration method for measuring the pKa of ionizable pesticides has been reported, which gives log P values at the same time (Chamberlain et aI., 1996). The method typically involves two titrations: The first is to measure the pKa value in the water phase and the other is for an apparent "pKa" value, pK~, after addition of an appropriate volume of l-octanol. The difference, llpKa = pK~ - pKa(> 0), is a function of the P value of the neutral form as shown in the following equation, where V is the volume of the solvent phase (Kaufman et aI., 1975): (1)
The entire operation of titration and calculation of pKa and log P values has been automated (Avdeef, 1991; Clarke, 1984). It has been developed for measurement of stepwise pKa values as well as the true log P values of noncharged species and apparent log P values of ion-paired counterparts for multiprotic compounds (Avdeef, 1992). The true log P value more negative than - 2 for neutral species cannot be measured using this method because the llpKa value is too small to estimate accurately. Also the volume ratio of l-octanol/water should be high, making it difficult to titrate the octanol/water system including the test compound. The true log P value of amitrole (-0.97) was measurable, but that of glyphosate (-3.39 by the shakeflask procedure) was not (Chamberlain et aI., 1996). 29.2.3 HIGH-PERFORMANCE LIQUID CHROMATOGRAPHIC METHOD There are review articles dealing with reversed-phase highperformance liquid chromatography (RP-HPLC or HPLC) applied to the experimental estimation of log P values (Braumann, 1986; Terada, 1986). In HPLC, the affinity of a solute for the stationary phase is characterized by the retention factor k, which is defined as
k
=
(tR - to)/to,
(2)
where tR is the retention time of the solute and to is that of a nonretained reference compound. Because the chromatographic retention can be regarded as an equilibrium partition process between two "immiscible" phases, the log P of solutes in the
652
CHAPTER 29
Hydrophobicity as a Key Parameter of Environmental Toxicology
l-octanoVwater system is regarded as being linearly related to the log k empirically as log P = a log k
+ b.
(3)
The coefficients a and b in Eq. 3 are calculated by the regression analysis from the log k values measured for a series of related compounds whose log P values have been determined by the shake-flask procedure. If the statistical quality of Eq. 3 is sufficiently high, then the log P values of other compounds belonging to the same series can be estimated from their log k values and Eq. 3. It is unfortunate, however, that the stationary phase is not entirely able to simulate the octanol phase. The mobile-phase eluent is a solvent/water (solventlbuffer) mixture of certain proportions, the most commonly used solvent being methanol. Thus, the retention (partition) behavior of compounds in the HPLC system differs from that in the 1octanol/water system according to physicochemical differences of each of the two phases between the two experimental systems. As the stationary phase, various types of ODS (octadecyl silica trapping some amount of organic solvent from the mobile phase) columns have been most widely used. According to the difference in the hydrogen-bonding characteristics with the l-octanoVwater system, the solutes are generally grouped into non-hydrogen bonders, hydrogen acceptors, and amphiprotics. For each of the groups, a separate linear relationship of the type of Eq. 3 can be drawn in such a way that the a value is almost equivalent but the constant b differs among groups. Thus, hydrogen-bonding indicator variable terms (Fujita et aI., 1977) for groups can be added to Eq. 3 (Terada, 1986). With an ODS column and an acetonitrile/water eluent, Takahashi et al. (1988) measured log k values for a set of fungicidal substituted N -pheny1carbamates having various substituents on the benzene ring. From a significant correlation of the type of Eq. 3, observed for 31 analogs with log P values measured by the shake-flask procedure, the log P values for 38 new compounds were estimated from their log k values and used for the QSAR studies. It has been proposed that the log k value for a given compound increases linearly, over a moderate range, with decrease in proportion of the miscible organic solvent in the eluent (Karger et aI., 1976; Miyake and Terada, 1982). Recently, Yamagami et al. (1994) have observed that the linearity is not always valid. They also demonstrated that the log ko value defined by the extrapolation of log k values to the zero organic-solvent concentration in the mobile phase is not necessarily a useful predictor of the log P value. The use of eluents containing 50% (v/v) methanol gives the most reliable and straightforward correlation between log k and log P for a series of related compounds having non-hydrogen-bonding and hydrogen-accepting substituents. For those having amphiprotic substituents, the situation is less straightforward, requiring additional parameters (Fujita et aI., 1977). Yamagami and Fujita (1995) and Yamagami et al. (1990, 1994, 1995) have further indicated that the correlation of the type of Eq. 3 should generally be dealt with so as to include, besides the previously mentioned indicator
variable terms, parameter terms representing mutual electronic effects among substituents or functional groups in a number of substituted heteroaromatic compound series. These electronic effects are thought to govern variations in the hydrogenbonding interactions of solutes with the mobile phase relative to that with the stationary phase. With improvements in the chromatographic system as well as e1aborations in the combinations of solid support and mobile phase, the HPLC method is reported to work very well for estimation ofthe log P ranging from 0 to 6 (Sangster, 1997a). The most convenient aspect of the HPLC method is that the analytical measurements are not required. This advantage can only be guaranteed, however, with highly qualified model correlations of the type of Eq. 3.
29.3 NONEXPERIMENTAL ESTIMATIONS OF log P Experimental log P values are correctly measurable/estimable under careful conditioning. However, experimental procedures are sometimes time consuming, especially for highly hydrophobic and hydrophylic compounds. Moreover, the log P values of compounds are required before synthesis in order to assess their environmental behaviors as well as biological activity. Using recently developed methods of combinatorial synthesis, an enormous number of possible pesticide candidates are synthesized almost simultaneos1y. To measure the log P value of each compounds experimentally is nearly impossible within a given period of time. Thus, nonexperimental procedures, preferably computerized, for the log P estimation are greatly needed. It should be mentioned that the construction of any estimation procedure is based largely on experimental log P values of existing compounds analyzable in terms of their physicochemical nature and composition. There is always the need for a reliable training set of standard experimental log P values to judge the performance of the estimation procedure. After an introductory section about the additive-constitutive nature of log P values, some empirical and computerized procedures will be described in this section. For other procedures, especially computerized systems, review articles (Leo, 1993; Sangster, 1997b) should be consulted.
29.3.1 SCOPE AND LIMITATION OF THE ADDITIVE NATURE OF log P VALUES The log P value of a molecule had been recognized to be roughly expressible by the sum of the hydrophobicity indices attributable to submolecular components such as a parent skeleton and substituents (Fujita et aI., 1964; Iwasa et aI., 1965). For a monosubstituted aromatic compound PhX, the hydrophobic constant :rr of the substituent X is defined as follows, where (X/PhH) denotes that X is to be introduced into unsubstituted benzene PhH: :rr(X/PhH) = log P(PhX) - log P(PhH).
(4)
29.3 Nonexperimental Estimations of log P
653
Table 29.1 Hydrophobicity Parameter n of Common Substituents n(XjRH)b
n(XjRH)b
Substituent X
n(XjPhH)a
F
0.14
-0.73
OH
-0.67
-1.80
Cl
0.71
-0.13
OMe
-0.02
-0.98
Br
0.86
Substituent X
n(XjPhH)a
1.12
0.04 0.22d
OEt
0.38
_c
OPr
1.05
_c
0.53 e
OCF3
1.04
-
Me
0.56
Et
1.02
_c
Pr
1.55 1.53
i-Pr
C
OCONHMe
-0.97
C
c
OCH2COOH
-0.79
_c
C
NH2
-1.23
-1.85
_c
NMe2
0.18
-0.95
c
NHAc
-0.97
_c
NHCONH2
-1.30
_c
CONH2
-1.49
-2.28
-
CH20H
-1.03
CH2COOH
-0.72
CF3 CHO
0.88 -0.65
_c
Ac
-0.55
-1.26
N02
-0.28
COOMe
-0.01
-0.91
SMe
0.61
COOH
-0.32
-1.26
SCF3
CN
-0.57
-1.47
S02 Me
-1.63
-
_c
S02 NH 2
-1.82
_c
Ph
1.96
0.161
-1.07 8 _c c
1.44
C
aFrom Hansch and Leo (1995).
bUnless noted, R = PhCH2CH2CH2 from Iwasa et al. (1965). The n(XjRH) varies depending on the structural features of R. See Section 29.3.5. CNot available, but calculable by the CLOGP procedure described in Section 29.3.5. dThe log P value of PhCH2CH2CH21 is from Leo et al. (1971). eR = PhCH2CH2. fR = PhCH2CH2, taken from Takayama et al. (1985). 8R = PhCH2CH2, taken from Hansch et al. (1995).
On certain occasions, there is a simple additivity so that the log P of disubstituted benzene PhXY can be expressed as log P(PhXY)
= log P(PhH) + JT(X/PhH) + JT(Y /PhH).
(5)
However, such a simple additivity is severely limited and does not hold in general. As a minimum requirement for Eq. 5, X and Y should be neither capable of hydrogen bonding nor located close together (vicinally). In certain PhXY series, X and Y usually denote variable and fixed substituents, respectively. In general, the JT (X/Ph Y) value of a certain X substituent is not a constant among various Ph Y systems, but varies depending on the nature of, as well as the location relative to, the Y substituent (Fujita et aI., 1964). The hydrogen-bonding solvation of the Y substituent, when it is capable of hydrogen bonding, with octanol relative to that with water varies depending on the electron-withdrawing character of the X substituent. The susceptibility of variations in the relative solvation of the electronic effect of X differs among various Ph Y series according to the structural features of the Y substituent. The situation will be illustrated taking the X-substituted phenols (Y = OH) and the corresponding X-substituted benzoic acids (Y = COOH) as examples. The JT value of a certain X substituent in these disubstituted benzenes, being defined as log P(PhXY) -log P(Ph Y), differs between the two systems. Besides the "intrinsic" hydrophobicity of the X substituent, the difference in the pattern of the solvation with partitioning solvents between OH and
COOH is involved by definition, leading to the variations in the JT(X/PhY) value (see Section 29.3.4 for further references). For the aliphatic substituents, the JT value also varies depending on the stereoelectronic situations in which the substituents are located. Thus, the JT(X) value in the aliphatic system, JT(X/RH), where R is the alkyl, differs from that in the aromatic system, JT(X/PhH). The only exception is the JT value of eletronically nearly neutral alkyl groups. JT(H) is always 0 by definition. Structural factors such as branching, cyclization, and multiple bonding, as well as inductive interaction between polar groups through single bonds and conjugation of JT -electron systems are known to contribute to variations in the molecular log P value. Table 29.1, lists the JT(X/PhH) and JT(X/RH) of some common substituents. The structural factors participating in the log P values will be described in Section 29.3.5. 29.3.2 EMPIRICAL (MANUAL) PROCEDURE
With the preceding limitations, one should be very careful to use the "simple" additivity principle "manually." Some appropriate examples follow. Equation 6 is for a synthetic pyrethroid, phenothrin. The first three terms together, 5.57 (= 3.76+ 1.990.18), represent the log P value estimated for benzyl chrysanthemate (I: X = H). The experimentally measured log P value, 5.49, agrees well with the estimated value, considering the dif-
654
CHAPTER 29 Hydrophobicity as a Key Parameter of Environmental Toxicology
ficulties that were supposed to have occurred in the shake-flask procedure (Nakagawa et aI., 1982). CH,oC~ COOCH,--oX 11
In
log P(Phenothrin, I: X = m-OPh)
XCH2CONH2 structure, which is actually equivalent to ~n = n[OPhjH(CH2CONH2)] - n[CH2PhjH(CH2CONH2)]. This difference is considered to simulate the difference in the log P value between molecules with the XCH2CONHPh structure. As the previous two examples show, the stereoelectronic environments of the reference compounds or reference substituents should be carefully selected so as to be as close as possible to those of compounds the log P value of which is to be estimated with this procedure.
= log P(Methyl1R-trans-chrysanthemate)
3.76 (measured; Nakagawa et aI., 1982)
29.3.3 EMPIRICAL PROCEDURE USING RELATIONSHIPS WITH log P VALUES OF SIMPLER COMPOUNDS
+ log P(CH3COOCH2Ph) 1.99 (measured; Nakagawa et a!., 1982) -log P(CH3COOMe)
The following equation was derived from experimentally measured log P values for a series of N-acyl-N'-alkyl- and N-acylN'-phenylureas, including insecticidal benzoylphenylureas (IV) in which Xl, X2, and Y were variously changed (Sotomatsu
0.18 (measured; Hansch et aI., 1995) +n(OPhjPhH) 2.08 ("measured"; Hansch et a!., 1995) = 7.65.
(6)
The second and third terms in Eq. 6 represent the n[Phj (CH3COOCH2)H]. The stereoelectronic environment of the Ph group in benzyl acetate and that of the Me group in methyl acetate can be simulated well by that of the Ph group in benzyl chrysanthemate and that of the Me group in methyl chrysanthemate, respectively. The addition of the n (OPhlPhH) value does not seem to be correct enough because it is from the nonsubstituted benzene system. The aromatic system, to which the OPh substituent is introduced, has, however, a substituted "toluene" (benzyl) substructure, so that the electronic effect of the OPh on the solvation of the ester ( -OCO-) grouping is attenuated by the methylene unit. In fact, the n[Xj(CH3COOCH2C6H4)H] values were experimentally shown to be almost equivalent to the corresponding n(X/PhH) values (Nakagawa et a!., 1982). The log P values of other substituted benzyl chrysanthemates (I) were estimated by Eq. 6 using n(XjPhH) in place of n (OPhjPhH) (Nakagawa et aI., 1982). Similar procedures were used for substituted benzyl pyrethrates (11) (Nishimura et a!., 1987) and kadethrates (Ill) (Matsuda et a!., 1989) in which the log P value of the methyl esters of skeletal acids was experimentally measured. The log P value of an experimental amide-type Hill reaction inhibitor (Shimizu et aI., 1988) was derived as follows: log P(3,4-Cl2C6H3NHCOCH20Ph)
= log P(3,4-Cl2C6H3NHCOCH2CH2Ph) 4.85 (measured; Mitsutake et a!., 1986) + log P(PhOCH2CONH2) 0.76 (measured; Hansch et aI., 1995) -log P(PhCH2CH2CONH2) 0.91 (measured; Iwasa et a!., 1965)
= 4.70.
(7)
The second and third terms together express the ~ log P value between the OPh and CH2Ph compounds having the
et aI., 1987). X,
~CONHCONH~Y X2
IV
log P(RCONHCONHR')
= 0.955(±0.031) [log P(RCONH2) + log P(NH2CONHR')] + 1.938(±0.054), n
= 10,
s
= 0.072,
(8)
r = 0.999.
In this and the following equations, n is the number of compounds, s is the standard deviation, and r is the correlation coefficient, whenever applicable. The numbers in parentheses represent the 95% confidence intervals. Rand R' are either simple alkyl or (un)substituted phenyl existing in structure IV. Equation 8 indicates that the molecular log P value of the Nacyl-N'-substituted ureas can be linearly related to the sum of the log P values of the component-substituted amide and urea. The effect of the R group on the CONH2 in amides is almost linear (with a slope of nearly unity) to that on the (CONHh bridge in the molecule. The effect of the R' group on the NHCONH2 substituent in R' -substituted urea is also almost linear (with a slope of nearly unity) to that on the (NHCOh moiety. In this simple combination, the NH between two carbonyl groups is counted twice. The total effect of the Rand R' substituents on this NH group is incorporated into the summation of the log P values. The difference between the observed value and the summation term would be, however, almost constant and included in the constant term in Eq. 8. Equation 8 also shows that the n(XjPhCONH2) as well as the n(XjPhNHCONH2) values can be used to estimate the log P values of substituted benzoylphenylureas with the value of the experimental log P(PhCONHCONHPh) (= 3.39) (Sotomatsu et a!., 1987).
29.3 Nonexperimental Estimations of log P
655
Table 29.2 Hydrophobic Fragment Constants (f) and Correction Factors (F) Fragment
aC
pd
J(ar.)b
0.23
_e
Fragment -0-
J(al.)Q
_e
-1.82
0.37
0.28
0
-N=f
-2.37
-0.61 -1.12
0.94
0.28
0
-S-
-0.79
0.03
0.20
1.09
0.28
0
-SOl
-3.13
-2.17
0.59
1.35
0.28
0
-NH-
-2.15
-1.03
0
1.08
-l.l6
-0.03
0.60
0
-CO-
-1.84
-1.09
0.51
0.27
-1.64
-0.44
0
1.06
-COl
-1.45
-0.56
0.51
0
0.50
-CH=N-
-1.20
-1.03
0
0.61
1.08
-CONH-
-2.71
-1.81
0.32
0.72
J(al.)Q
H
0.23
F
-0.38
Cl
0.06
Br N02 OH
J(ar.)b
pd
aC 0.17
0
0.50 0.61
0
0
0.70
0.45
SH
-0.23
0.62
NH2
-1.54
-1.00
0
NHCONH2
-2.18
-1.07
0
1.08
-OCONH-
-1.79
-1.46
0.17
0.50
S02NH2
-2.37
-1.61
0.35
0.88
-2.18
-1.57
CF3
_e
-NHCONH-CON(A)-g
-3.14
-2.80
0 0.51
0.27
CN
-1.27
0
1.08
l.ll
0.49
0
-0.34
0.65
0
-OPO(OAh 8
-2.29
-1.71
0.17
0.80
0.35
-N=h
e
-1.14
0.90
0.30
0.60
=C=i
0.20
_e
- e
COOH
-1.07
-0.03
CONH2
-1.99
-1.26
Aliphatic
0.32 0.32
0.13
Aliphatic
features
F
Aromatic features
-0.13
features X-C-X
F
Chain branch
0.6-2.81
Internal H bonding
Group branch
-0.22
X-C-C-X
0.28 k
Ortha effect
-0.281
Double bondm
-0.09
X-C-Y
0.9-2.7 n
Triple bondm Chain bond P Ring bondr
-0.50
X-C-C-y
-0.12
Y-C-Y y-C-C-y
0.35-0.45 Q -0.26--0.42Q -0.15--0.26Q
Electronic interaction among fragments
Lpa s
-0.09
F
0.63
a Aliphatic.
bAromatic. cElectron-withdrawing effect of aromatic fragments on others regardless of their positions. d Susceptibility to the L a of other aromatic fragments. eNot applicable. fTrivalent nitrogen. 8"jI;' means to be connected to the aliphatic IC. h Nitrogen fused in heteroaromatic rings such as pyridine. iJsolating carbon. ilncreases with the number of geminal halogens (X) from 2 to 4. kMultiplied by the total number of halogens (on both sides of C-C) minus unity. I At least one of the artha pair is capable of H bonding. Multiplied by a number defined for each combination of ortho pairs. m For isolated bonds. nVaries depending on the type of Y (H-bonding fragment) and the number of geminal X. 00.45 when X = F. PMultiplied by the number of bonds connecting fragments minus unity. qMultiplied by L J(Y), which is usually negative to give positive corrections; varies according to structural features of Y. rMultiplied by the number of bonds. sSee Table 29.4.
The log P value of 10 dibenzoylhydrazines (V) acting as ecdysone agonists to various extents was nicely analyzed to give the following with the use of the corresponding submolecular log P value (Oikawa et aI., 1994).
Q- l'fU
CONNHCOU Xn
Cl
V
log P(Compounds V)
= 1.008(±0.040) log P(XnPhCONH2) -0. 158(±0.030) L E~rtho + 1.952(±0.046), s = 0.015, n = 10, r = 0.999.
(9)
Equation 9 clearly indicates that the log P of compounds V, in which the substituents X(n) are varied, corresponds nicely with that of the X(n)-substituted benzamides. L E~rtho is the summation of the Es value of the X substituent(s) located at
656
CHAPTER 29
Hydrophobicity as a Key Parameter of Environmental Toxicology
substituents on the Y substituent but also on the "backward" effect of the Y substituent on the X substituents. This situation is represented in general as (Fujita, 1983; Fujita et al., 1964) /';.1T
= 1T(X/PhY) -
1T(X/PhH)
= pya~ + pxa~ + c.
(11)
a O is one of the variations of the Hammett a value used in situFragment types, correction factors
Components in CLOG
Estimated values
Polar fragment
Thioiminocarbamate
-1.850 (f)
Isolating Cs
3 x aliphatic ICs
0.585 (3 x 0.195f)
Ex-fragment Hs
9 x H on ICs
2.043 (9 x 0.227 f)
Ex-fragment bonds
(3-1) x bond factor
-0.240 (2 x -0.12F)
Measured = 0.60a a
La! + LbF =
0.538 (CLOGP)
Drabek and Bachmann (1983).
ations where any through-resonance interaction does not occur directly between the X and Y substituents (Yukawa and Tsuno, 1959). For the log P values of meta- and para-disubstituted benzenes relative to that of unsubstituted benzene, Eqs. 10 and 11 are modified as follows, respectively, in which the slash within the parentheses following the /';.log P notation means just "relative to" (Nakagawa et al., 1992): /';.log P (X -C6H4 - Y /PhH) - L1T(X, Y/PhH)
the artha position(s) as defined by Taft (1956) (for aliphatic substituents) and extended by Kutter and Hansch (1969). Note that the reference point has been shifted to that of hydrogen so that Es(H) = O. The Es value is defined so that the bulkier the substituents, the more negative the value. Thus, the negative L E~rtho term means that, by increasing the sum of the size of the arrha substituents (x~rtho, n = 1 or 2) in compounds V, the log P value increases relative to that of the corresponding benzamides. This equation was used to the estimate log P values of some 60 analogs ranging from 2.0 to 5.5 (Oikawa et al., 1994).
= pya~ + c,
(12)
/';.log P(X -C6H4 - Y /PhH) - L
1T(X, Y /PhH) = pya~
+ pxa~ + c.
(13)
For meta- and para-substituted benzoic acid (X -C6H4 COOH), the preceding situations are represented as follows. For derivatives in which X is not capable of hydrogen bonding, Eq. 12 is enough without considering the backward effect of the COOH function on the non-hydrogen-bonding X substituents (px = 0) to give (Fujita, 1983) /';.log P = 0.964(±0.07l) L1T(X, COOH/PhH) +0.499(±0.140)a~(m, p)
29.3.4 EMPIRICAL PROCEDURE USING FREE-ENERGY RELATED SUBSTITUENT PARAMETERS
n = 12,
The preceding procedures can be used in a limited series of compounds under conditions where the log P values for the reference series of (simpler) compounds are either available or measured. A more general procedure is certainly needed. The 1T (X/Ph Y) value of the substituents in a series of metaand para-substituted phenols and ani lines (X -C6H4 - Y: Y = OH or NH2) was shown to be related to the 1T(XlPhH) value in monosubstituted benzenes in the following manner (Fujita etal.,1964): /';.1T
= 1T(X/PhY) -
1T(X/PhH)
= pyaX + c.
(14)
-0.021 (±0.048),
(10)
In Eq. 10, ax is the Hammett constant of each non-hydrogenbonding substituent, X, representing its electronic effect on the relative hydrogen-bonding solvation of the fixed substituent Y (OH or NH2). py is the susceptibility constant of the Y substituent to the variations in the electronic effect of X. The intercept c should be as close as to O. This type of procedure for analyzing the hydrophobicity parameters of substituted benzene systems using free-energy related substituted parameters has been continued. For meta- and para-disubstituted benzenes, X -C6H4 -Y, where X and Y are both capable of hydrogen bonding, the /';.1T value in Eq. 10 depends not only on the "forward" electronic effect, ax, of the X
s = 0.035,
r = 0.997.
In Eq. 14, the L 1T term corresponding to the second term on the left-hand side of Eq. 12 is moved to the right-hand side. Although it should be close to unity, the slope of this L 1T term is not necessarily equal to unity. As a counterpart of Eq. 13 for cases in which the set of variable X includes hydrogen-bonding substituents, the following was derived (Sotomatsu et aI., 1993): /';.log P = 0.999(±0.042) L1T(X, COOH/PhH)
+0.419(±0.095)a~(m, p)
(15)
+0.383(±0.1l5)px(m, p) - 0.019(±0.044), n =22,
s = 0.049,
r =
0.998.
In Eqs. 13 and 15, the Px value is the susceptibility parameter of the X substituents capable of hydrogen bonding to the (backward) electronic effect of the fixed Y (COOH) substituent. It is used as an independent variable. It is estimated as being equivalent to the py value in the correlation of the type of Eq. 14 for a series of compounds in which X is now fixed and Y is varied only within non-hydrogen-bonding substituents (X is replaced by Y in Eq. 12). The Px value of non-hydrogen-bonding substituents is O. The regression coefficient of the Px term corresponds to the a~ value of the Y substituent. Thus, according
657
29.3 Nonexperimental Estimations of log P Table 29.4 CLOGP Example Calculation of Terbutryn
Fragment types, Components in CLOGP
Estimated values
Polar fragments
2 x 2° amine
-2.060 (2 x -1.030f)
Polar fragments
3 x aromatic (fu sed) N
-3.420 (3 x -1.1401)
Polar fragment
sulfide, -$-
Isolating Cs
7 x aliphatic rcs
1.365 (7 x 0.195f)
Isolating Cs
3 x aromatic rcs
0.390 (3 x 0. 130f)
correction factors
Ex-fragment Hs
17 x H on ICs
Ex-fragment bonds
7 x bond factor
0.030 (f)
3.859 (17 x 0.227 f) -0.840 (7 x -0.12F)
Ex-fragment branch
I x chain branch
-0.130 (F)
Ex-fragment branch
I x group branch
-0.220 (F)
Electronic interactions
L
Ortho correctiona
Between -$- and =N-
3.755 (F)
pa, see below
0.400 (F)
Measuredb = 3.38
La! + LbF =
3.129 (CLOGP)
Composition of electronic interactions
$ubstituent (X)
OX
px
Composition of effects
Fading factor of
towards XC
each pa product<
Total C
I
-$-
0.00
0.00
2
=N-
0.90
0.30
a(4-N,6-N) x p(2-N)
1.0,0.3
3
-NH-
0.00
1.08
a(2-N,4-N,6-N) x p(3-NH)
1.0, 0.3, 0.09
1.351
4
=N-
0.90
0.30
a(2-N,6-N) x p(4-N)
1.0,0.3
0.351
5
-NH-
0.00
1.08
a(2-N,4-N,6-N) x p(5-NH)
1.0, 0.3, 0.09
1.351
6
=N-
0.90
0.30
a(4-N,6-N) x p(6-N)
1.0,0.3
0.351
0 0.351
L (Aromatic electronic interactions) = 3.755 aThe positive correction may be due to a I : l hydration in the octanol phase (Fujita, 1983). bFrom Liu and Qian ( 1995). C After the a value of each of the "substituents" parenthesized is multiplied by the p value of X, each of the ap products is weighted by each of the "fading factors " respectively in the descending order and the weighted ap products are summed up,leading to "total."
to the relative position of the X and Y substituents, two Px [px(rn) and px(p)] terms should "theoretically" be considered. Because the a~(rneta) and a~(para) values as the coefficients of the two Px terms were close and no statistical difference was observed, the two independent terms are combined into the single pxa~ term in Eq. 15. For a set of n values, including those of the ortho-X substituents (Fujita, 1983; Fujita and Nishioka, 1976), the "regular" electronic effect of the artha substituents is taken as being equivalent to that of the corresponding para substituents, so that a~(artha) = a~(para), whereas the "proximity" electronic
effect is expressible by such an inductive effect parameter as aI, defined by Charton (1981). In addition, the proximity steric effect of artha substituents is represented by the Taft-Kutter-
Hansch Es parameter mentioned in Section 29.3 .3 (Kutter and Hansch, 1969). Because the proximity effects also work bidirectionally, Eq. 13 can be modified as
~logP(X-C6H4-Y/PhH) -
I> (X, Y/PhH)
= pya~(a, rn , p) + pxa~(a, rn, p) + p[ ajX(o) X Y +pfal (0) + 8 E;-(a) + 8 E; (a) + c. (16)
658
CHAPTER 29
Hydrophobicity as a Key Parameter of Environmental Toxicology
Table 29.5 CLOGP Calculation of Triazole Fungicides
(CH 2)
0--R--H
X
-
HO
N\\
~,~N\
X
n
Measureda
CLOGpb
2-CI
4
2.68
2.63
2,5-CI2
4
3.24
3.34
4-CI
5
3.03
3.19
QKataoka et al. (1989). bLeo (1998).
pt,
In Eq. 16, a~, at, E~, Px, and 8x are used as independent variables. The first three variables are for the "forward" effect of variable X substituents on the Y substituent, whereas the next three are for the susceptibility to the "backward" effect of the Y substituent on the hydrogen-bonding X substituents. To extend the procedure toward a large number of multisubstituted compound series in which various substituents capable of hydrogen-bonding are involved, differentiation between the forward and backward effects is not straightforward. Instead, Eq. 16 can be generalized so that only the forward effects of every other substituent are considered on each of the hydrogenbonding substituents. Because the forward effects are generally composed of three (regular and proximity electronic and steric) components, the extended correlation equation to analyze as well as to predict the !1log P [= log P - log P (PhH)] value takes the form !110g P(Xi-benzene/benzene) = a I>r(Xi/PhH)
+Pr
+ l]p L
L ar(o) + 8 L
aO(o, m,
Es(o)] + c.
p) (17)
L IT covers all substituents, Xi, on the benzene ring. The second term on the right-hand side takes care of the three components of the "forward" effect. Lao, Lar, and L Es are made for substituents at positions indicated in parentheses relative to individual hydrogen-bonding substituents. The L sign outside the brackets means to sum up the forward effects on every hydrogen-bonding substituent. Each of the a, P, Pr, 8, and c values is calculated by regression analysis (Nakagawa et aI., 1992). The preceding procedure, counting substituent effects toward every hydrogen-bonding substituent "forwardly" and "multiply," has been nicely used to analyze some 200 log P values of multi substituted benzenes (n = 210, s = 0.118, r = 0.994) with some corrections for intramolecular hydrogen bonding and a buttressing effect of vicinally located substituents on the solvation of hydrogen-bonding substituents (Nakagawa et aI., 1992). Recently, a procedure similar to that
described previously has been applied to analyze and predict the log P values of substituted pyridines and diazines, in which the fused N atom in the ring is dealt with as being a substituent (Yamagami et aI., 1995). The procedure works well with certain approximations, but indicates that further elaboration is required for heteroaromatic compounds. The log P value of a number of zwitterionized di- to pentapeptides at the isoelectric point has been analyzed with the use of free-energy related parameters under defined conditions (Akamatsu and Fujita, 1992). The IT parameter of the side chain of the amino acid units is that for aliphatic substituents (Iwasa et aI., 1965; Leo et aI., 1971). Along with the term for the steric effect of the side chain in terms of a variation of Es, correction terms for polar side chains interacting with the backbone -CONH - structure and the ,B-turn formation are included (Akamatsu and Fujita, 1992). A similar set of aliphatic parameters is used in the analyses of the log P values of primary, secondary, and tertiary amines and quaternary ammonium ions pairing with picrate (Takayama et aI., 1985). 29.3.5 COMPUTER-AIDED PROCEDURES To make a more comprehensive procedure that covers a wide variety of compounds as aliphatic, alicyclic, (hetero )aromatic, and various combinations, it should be computerized with an access to a large database of reliably measured log P values. Various "rules" governing the structural contribution to log P, and correction terms for various intramolecular interaction features specific to each series should be incorporated into the program software, and calculated values should be immediately compared with the measured values in the database whenever available. The software could be constructed so as to recognize structural features of compounds when it is input into the computer and to output the calculated log P value "automatically" after data processing according to the rules and corrections covering any types of structural features. There are quite a few programs based on this type of concept (Sangster, 1997b). One of them, the "fragmental method," was first proposed by Nys and Rekker (1973). They have been developing software for their own fragmental method (Rekker and Mannhold, 1992). With a procedure different from that used in the Rekker method, Leo and Hansch developed another fragmental method, the software of which is called CLOGP (Hansch and Leo, 1979, 1995; Leo, 1991,1993; Leo et aI., 1975). In this section, the CLOGP procedure is briefly described. In principle, the fragmental method is based on a "correlation" equation such as the following to elucidate the measured log P value: (18) Here, a is the number of occurrences of molecular fragment f of type n, and b is the number of occurrences of correction factor F of type m. The fragmental hydrophobicity index f differs from the IT value for a certain substituent X. For instance, the log P of chlorobenzene is conceptually expressible by either of
29.3 Nonexperimental Estimations of log P
the following equations, although the f(C6HS) value is algorithmically dealt with as being divisible into smaller fragments in the CLOGP: log P(PhCI) = log P(PhH) log P(PhCI) = f(C6HS)
+ n(CI/PhH),
+ f(CI).
(19) (20)
Because log P(PhH) in Eq. 19 is expressible as the sum of f(C6HS) and f(H), the relationship between the nand f values can be represented as n(CI/PhH)
=
f(CI, aromatic) - f(H, aromatic).
(21)
Table 29.2 lists the f values of some representative substituents as well as the F values of typical correction features (Leo, 1998). As indicated, an "isolating carbon" (IC) atom is defined as the one not doubly or triply bonded to a heteroatom. An IC can be bonded to heteroatom inside an aromatic ring, and one IC can be multiply bonded to another. The IC and hydrogens attached to it (ICHs) are considered hydrophobic fragments. All atoms and groups of covalently bonded atoms, which are left after removing ICs and ICHs, are considered polar fragments. The polar fragments do not contain ICs, but each is connected to ICs with one or more bonds. The f value is assigned first to each fragment. Depending on the bond environment, such as aliphatic and aromatic as well as benzyl and vinyl, the f value of polar monovalent fragments is found to vary, with the aromatic value highest and the aliphatic value is lowest. Vinyl and benzyl values are intermediate. This difference is probably due to variations in the degree of delocalization of electron pairs of the polar fragment. The correction factors, F, for six types of fragment interactions are noted in the aliphatic systems. When halogens, X, and hydrogen-bonding fragments, Y, are located either geminally or vicinally in such arrangements as X -C- X, X -CC- X, X-C-Y, X-CC-Y, Y-C-Y, and Y-CC-Y, the F value for "proximity polar" interactions is assigned to each of them as positive corrections with detailed rules depending on their structural features. For aromatic systems, electronic interactions among substituents modifying their hydrogen-bonding capability, such as those described in the preceding section, are incorporated after modifications/simplifications. The "ordinary" electronic effect of a certain substituent designated by the 0'°(0, rn, p) in Eqs. 16 and 17 is simplified/approximated here as being expressible as a single a value, some of which are listed in Table 29.2. The p value of substituents/fragments, which is not easily accessible directly by the analyses according to such correlation equation as Eqs. 13-17, is estimated/calculated so that each of the 0'/ P combinations (or pyaX products) makes up the difference between the measured value and the simple fragment sum. The p values of some fragments are also shown in Table 29.2. Provisions are also made for multiply substituted compounds, including heteroaromatic systems. Thus, the susceptibility p of each substituent to multiple interaction is not additive but attenuated starting with the greatest 0'/ P combination of fragments (substituents) regardless of whether they are present as substituents or as fragments fused in a heteroaromatic ring.
659
For ortho disubstitutions, besides the "ordinary" pO' interaction, a negative correction factor is assigned when the effect is regarded as a twisting of one of the substituents out of the ring plane, whereas a positive correction factor is defined for the internal hydrogen bond formation. In addition, the CLOGP program uses correction factors for the bond flexibility in aliphatic systems. For chain compounds, the total correction is made by a negative unit F value multiplied by the bond number connecting (or outside of) fragments minus unity (not counting those to H). For alicyclic compounds, the bond flexibility correction is made by a less negative factor. The branching structures in the alkane chain at ICs and at polar fragments are also considered as negative correction factors. Table 29.3 shows the CLOGP components of methomyl, an oxime carbamate insecticide (Leo, 1998). The polar fragment, including the amide H, is enclosed by the broken line, and three CH3 groups, each including an IC and three ICHs, connected to the polar fragment, are circled in the structural formula of methomyl. The CLOGP estimation is in a good agreement with that measured by Drabek and Bachmann (1983). Table 29.4 is that for the more complicated triazine herbicide, terbutryn (Leo, 1998). The six polar fragments numbered from 1 to 6 are shaded circles, whereas three aromatic ring ICs and seven aliphatic ICs with respective ICHs (five CH3, one CH2, and a quaternary C) are just circled. In Table 29.4, the f value of secondary (2°) amino (-1.030) and thio (0.030) fragments is that for the aromatic substituents. The f value of aromatic ring IC (0.130) is defined as being lower than that of aliphatic IC (0.195). In this molecule, the correction for the bond flexibility is made for each of the "side chains." The NH-tertiary butyl substituent is considered double-branched, whereas the NH-ethyl is a straight chain. Electronic interactions are considered between the three substituents and the three - N = fragments fused in the triazine ring as mentioned before. In terbutryn, the electronic interactions are assumed to be "faded with use" in terms of the ay px product. For instance, each of the fused "aza" nitrogens and secondary amino groups (X) undergoes the electronic effect ay of other fragments according to the susceptibility Px and its fading factor. The situation is illustrated in the lower half of Table 29.4. The ortho correction factor of 0.400 is empirically assigned because the measured log P value of 2-methylthiopyridine (Yamagami and Fujita, 1995) deviates positively from the CLOGP similar to this value. The value of 3.129 here for terbutryn is calculated using the newest version of the program (Leo, 1998) and differs from that in earlier publications, 3.73 (Hansch and Leo, 1995; Leo, 1991), but it is closer to the more recently measured value of 3.38 (Liu and Qian, 1995). With the preceding types of improvements for the estimation of the log P values of complicated compounds, the latest version of the CLOGP program seems to work very well. The correction factors are physicochemically as well as empirically reasonable enough. The elaboration in the treatment of the electronic interactions among aromatic substituents to assume the fading effect has been observed to also work well for the calculation of the log P of such candidate azole fungicides as shown
660
CHAPTER 29
Hydrophobicity as a Key Parameter of Environmental Toxicology
in Table 29.5 with the five-membered heteroaromatic system (Kataoka et al., 1989). Although a number of correction factors are required, this is a matter of course if one can understand that there are various types of interactions between solutes and solvents as well as between substructures within single molecules that can affect the log P value. Further elaboration is needed because a number of factors have yet to be taken into account (Leo, 1998). It is recommended not to evaluate the computer-aided system in terms of its computational efficiency/simplicity but to judge it in terms of the physical organic background of the program.
29.4 PHYSICOCHEMICAL SIGNIFICANCE OF log P IN ENVIRONMENTAL QUANTITATIVE STRUCTUREACTIVITY RELATIONSHIPS 29.4.1 BEHAVIOR IN SOIL
The behavior of pesticides in soil is governed by adsorption, movement, vaporization, and degradation (Arnold and Briggs, 1990). The adsorption of organic chemicals to soil and sediment plays a very important role in their transport and mobility in the environment. The adsorption is an exothermic process, and a strong adsorption is observed in a low temperature range. The soil adsorption coefficient, Kd, in a soil/slurry-water system is expressed by the ratio of the amount of compounds adsorbed in soil (!-lg/g soil) to the concentration in water (J.Lg/ml) at an equilibrium state under defined conditions. The soil adsorption of organic un-ionizable pesticides has been shown to be expressible by the hydrophobic/hydrophilic balance parameterized by the partition between organic solvent and water or the chromatographic retention (Hance, 1967). Thus, for a certain soil sample, the log Kd value is observed to be almost linearly correlated with the log P value for a related series of compounds. For instance, Uchida and Kasai (1980), using a local soil sample in Japan, obtained the following linear relationship between the log Kd values of a rice blast fungicidal isoprothiolane (VI: RI = R2 = i-Pr) and related compounds and their log P values: log Kd
= 0.41 (±0.06) log P
n = 12,
s = 0.150,
= 0.980.
, "-8 0N
(22)
Uchida et al. (1982b) also found that the log Kd values calculated from Eq. 22 for structurally unrelated buprofezin (VII), an insect growth regulator, and flutoranil (VIII), a rice sheath blight fungicide, agree well with experimentally estimated Kd values using the same soil sample as before. The basic mechanism of the soil adsorption of organic compounds from the water phase had been recognized earlier to be that involved in the distribution between soil organic matter and water (Goring, 1962). Thus, the "soil organic-carbon adsorption coefficient" or Koe should be normalized by the proportion of
N)=N-tett_Su
VI
VII
I CF~
CX;° N'V0'~
I
iso-Pr
~
VIIJ
organic components in soil samples, Foe, as determined in a separate experiment and defined by
= Kd/ Foe.
Koe
(23)
The definition assumes that pesticide adsorption by soils is entirely due to organic matter, even though the organic matter is a complex mixture of carbon, hydrogen, and nitrogen compounds, which acts as a nonpolar film coating soil particles. The Koe value is relatively constant for a particular compound among soil samples from different origins. Briggs (1981 b) reported that the Koe value of herbicidal phenylureas measured with English soils agrees well with that measured with Australian soils. Dzombak and Luthy (1984) showed that the Koe value of chlorobenzenes hardly depends on soil samples having various fractional organic matter ranging from 0.15 to 33%. Oliver (1987) determined Koe values for chlorinated hydrocarbons, including polychlorinated biphenyls with suspended sediments, collected from various districts in the United States, yielding log Koe
= O. 76 log P + 1.66,
n = 19,
s = 0.27,
r =
(24) 0.93.
These relationships for a number of individual series of compounds were comprehensively examined by Sabljic et al. (1995). They derived the following equation for a number of organic compounds belonging to such classes as acetanilides, carbamates, phenylureas, phosphates, triazines, triazoles, and uracils, including practically used agrochemicals: log Koe = 0.47 log P
- 0.40(±0.24),
r
0)-- ,iso-Pr
n = 216,
s
+ 1.09,
= 0.425,
(25) r =
0.826.
In Eq. 25, log Koe varies from 0 to 5 and log P from -I to 8. They proposed that this equation can apply to the prediction of log Koe values of nonmeasured agrochemicals irrespective of origin of the soil samples. In fact, Eq. 25 is very similar to the following equation, which was derived for phenylurea-type herbicides independently (Liu and Qian, 1995) using soil samples from China: log Koe n =9,
= 0.59(±0.15) log P + 1.18(±0.43), s = 0.152, r = 0.961.
(26)
29.4 Physicochemical Significance of log P in Environmental Quantitative Structure-Activity Relationships
The mobility of pesticides in soil is important in governing their persistency as well as downward movement to pollute the ground water and lateral movement to pollute surface water. Mobility can be estimated using soil column chromatography or soil thin-layer chromatography. Uchida and Kasai (1980) studied the mobility of isoprothiolane and its analogs (VI) using a soil column chromatography and expressed it as log f.l. , where f.l. is the ratio of the volume of soil packed in the column to the volume of the aqueous phase required to elute the solute i.e., f.l. = Vsoil/ Veluent. They derived the following equation: log f.l.
= -1.44(±0.17) 10gKd - 0.45(±0.33),
n = 11,
s
= 0.193,
r
= 0.986.
(27)
Equation 27, together with Eq. 22, indicates that the greater the log P, the lower the mobility of the compounds in soil. Briggs (1973) obtained similar results for herbicidal phenylureas with soil thin-layer chromatography. Their mobility in tenns of the RM (Boyce and Milborrow, 1965) nonnalized by the content of organic matter was related to the log Koc value irrespective of the source of soils. Helling (1971) measured the mobility of ionic/ionizable herbicides such as dicamba, 2,4-dichlorophenoxyacetic acid (2,4-D), fenac, picloram, and diquat by thin-layer chromatography with 14 kinds of soil samples. The mobility of ionic!ionizable compounds was not simply related to the log Koc value. Their soil adsorption mechanism is not a simple hydrophobic partitioning into the soil organic matter from the aqueous phase, but includes various interactions. According to Wauchope et al. (1992), these interactions include (a) binding of cations to negatively charged sites on clay surfaces (a very strong interaction), (b) binding of anions to soil anion-exchange sites (a very weak interaction), and (c) specific chemical binding mechanisms such as the phosphate-fixation-like binding of glyphosate and the arsenicals to soil metal oxides. In many cases, anionic and cationic pesticides, which give very low and very high Kd values, respectively, have no reported soil adsorption values, probably because the extreme values involved are difficult to measure. The vapor losses of volatile pesticides from soil to air depend primarily on the air/soil distribution constant, KAS. Vaporization from water is conveniently estimated by the Henry's law constant, KAW (air/water). In the soil/water/air system, the soil adsorption lowers the concentration of pesticides in the water phase and, in turn, in air. The air/wet soil distribution could be approximated by the ratio of KAW and a soil/water distribution constant such as Koc. The KAS value is thus a function of water solubility, vapor pressure, and soil adsorption. According to Arnold and Briggs (1990), the KAS constant is roughly expressible as a function of boiling point and log P value under defined conditions for such factors as the air/soil ratio, distribution of the pesticide in the soil sample, and climate. To a first approximation, pesticides in soil exhibit an exponential degradation according to the first-order kinetics. The degradation half-life, dTso, can be estimated from the reciprocal of the first-order rate constant. Degradation in soil occurrs,
661
however, as a combination of mechanistically complex processes. It generally includes abiotic and biotic processes. The abiotic degradation is due to chemical reactions such as hydrolysis, photolysis, air oxidation, and others. Depending on the structural feature, pesticides could suffer from various types of reactions. No common parameter such as log P alone is capable of describing a variety of reactivities. The reaction rate can be evaluated if experimentally established model systems are available. For biotic degradation of organic compounds, there have been quite a few efforts to establish the QSAR model correlation equations. With the use of acclimated mixed microbial cultures, Babeu and Vaishnav (1987) measured the 5-day BOD (biological oxygen demand in mmol/mmol chemical) of a wide variety of organic compounds, including aIcohols, acids, esters, ketones, and aromatics. They examined the correlation of 10g(BOD) with various physicochemical parameters for 45 compounds. Their analysis, showing that the 10g(BOD) values fit well a correlation equation with quadratic tenns of log(theoretical BOD), seems to be reasonable (n = 45, r = 0.862) among others. There would be an optimum theoretical BOD value for compounds to be most biodegradable. Molecules in which the number of carbon atoms is high necessarily have a high theoretical BOD value. They are often highly hydrophobic. The highly hydrophobic compounds, being trapped by cell membrane lipids, would not be easily incorporated into microbial cells. In fact, using the BOD data of Babeu and Vaishnav, Zakarya et al. (1993) fonnulated a correlation equation showing that the experimental BOD values are parabolically related to log P and linearly related to molecular volume (n = 43, s = 0.575, r = 0.906). In contrast to the preceding studies, Dearden and Nicholson (1986) discussed the significance of the electronic structure of the molecule. The 5-day BOD value for various types of compounds, including amines, phenols, aldehydes, acids, halogenated hydrocarbons, and amino acids, supplied by the U.S. Environmental Protection Agency, Duluth, Minnesota, was shown to be highly dependent on a new electronic parameter that is expressible as the difference in the modulus of atomic charge across a key bond in the molecule, which could be attacked by microbials (n = 79, s = 3.459, r = 0.993). In spite of the fact that the biodegradability is estimated under simplified conditions without soil, the QSAR model building needs much improvement. In reality, assignment of a single dTso value to each pesticide is impossible under field conditions. Although it could be related to dTso values from model systems, it is also highly sensitive to the type of soil with varying mineralogy, carbon and water content, and pH, and the distribution and activity of soil microbials as well as climate. Thus, it is almost impossible to build a comprehensive single QSAR model for the degradation of a variety of pesticide classes. Careful studies should be made under defined conditions perhaps on a series-to-series basis.
662
CHAPTER 29 Hydrophobicity as a Key Parameter of Environmental Toxicology
29.4.2 BIOACCUMULATION
Processes involved in the accumulation of environmental chemicals in various organisms through aquatic phases are generally classified into two types: bioconcentration and biomagnification (Connell, 1988). Bioconcentration is the accumulation of chemicals dissolved in water in fish and aquatic organisms through the gills and body surface directly. The bioconcentration factor (BCF) is defined as the ratio of the concentration of a chemical in an aquatic organism to that in the aqueous phase under steady-state conditions. The measurement of the BCF has been made with the average concentration of the chemical in the whole body absorbed through the gills, skin, and digestive tract of fish of small to moderate size reared in a sublethal aqueous concentration. Sometimes, the BCF is estimated as that for the lipid content of fishes. Following the work of Neely et al. (1974) showing that the log value of the BCF of nonpolar compounds can be correlated with their log P value, a number of examples have been accumulated. Mackay (1982) critically reviewed the BCF values and proposed that the fundamental process observed in bioconcentration is such that P and BCF are proportional; that is, the slope of the log BCF versus log P correlation should "theoretically" be close to unity. In the earlier publications (Neely et al., 1974; Veith et aI., 1979), the slope was often lower than unity, even when dealing with nonpolar compounds. Mackay suggested that the lower slope was the result of overestimation (by calculation) of the log P of compounds of high molecular weight. After omitting compounds, the log P value of which are above 6 or are unreliable, compounds that are ionizable, and compounds that can act as surfactants, Mackay (1982) proposed the following equation for 43 compounds. These are mostly nonpolar such as lindane, DDT, and (polyhalogenated) aromatic hydrocarbons with log P values ranging from unity to 6. 10gBCF -log P n
= 43,
r
= -1.32(±0.25),
= 0.974.
(28)
Equation 28 is valid as far as inert compounds having log P values lower than 6 are concerned. It also indicates that variations in fish species, with which BCF is measured, are insignificant in the general relationship between log P and log BCF values. The lipid content of the fish used in developing Eq. 28 has been estimated as being 5% (Connell and Hawker, 1988) and does not vary significantly among fish species. There are quite a few examples conforming to Eq. 28. Oliver and Niimi (1983) determined the BCF of chlorobenzenes in rainbow traut (Salmo gairdneri) as log BCF = 1.022(±0.057) log P - 0.632, (29) n =
11,
r = 0.993.
In this and other correlation equations (Connell, 1988; Davies and Dobbs, 1984; Isnard and Lambert, 1988; Oliver and Niimi, 1985; Opperhuizen et aI., 1985), the slope of the log P term
is indeed close to unity and the intercept ranges from -0.5 to -1.3. Considerable deviations from the linear relationship represented by Eqs. 28 and 29 have been observed, however, for highly hydrophobic compounds with a log P value > 6 (Bruggeman et aI., 1984; Opperhuizen et aI., 1985). Reduced membrane permeation (Opperhuizen et aI., 1985), lowered lipid solubility (Banerjee and Baughman, 1991), and other possible reasonings have been proposed. To simulate this situation better, Bintein and Devillers (1993) proposed the following equation using the bilinear model developed by Kubinyi (1977) for a number of compounds, roughly one-third of which belongs to pesticides: 10gBCF
= 0.91OlogP -1.97510g(6.8 x 10-7 P n = 154,
r
= 0.950,
+ 1) s
(30)
0.784,
= 0.347,
F
= 463.5.
In Eq. 30, F is the ratio of regression and residual variances. The compounds were selected so that they are mostly inert and cover a wide range of log P values, ranging from unity to 9. In the second term on the right-hand side, conventionally written as b[log(,8 P + 1)], the ,8 value is supposed to correspond to the volume ratio between lipid and aqueous phases involved in the entire system for the manifestation of biological "activity." For small P values, (,8 P + 1) is close to unity, so 10g(,8 P + 1) is 0, and Eq. 30 takes the form of Eqs. 28 and 29. For large P values, (,8 P + 1) is almost equal to ,8 P, so that 10g(,8 P + 1) is linear with log P. The value of -log,8 nearly corresponds to the log P value where the log BCF is maximum. Biphasic functions with linear ascending and descending sides and a rounded apical part are represented by this model. In Eq. 30, the positive slope for the ascending side is 0.910, and the negative descending slope is (0.9lO - 1.975 = -1.065). Among the compounds included in Eq. 30, 24 compounds have a log P higher than 6; that is, they are covering a part of the apical region and the descending phase. In measuring the BCF values of compounds having a large log P value, one needs to carefully set the test period. Oliver and Niimi (1983) showed that 120 days are needed to attain the equilibrated steady state for polychlorobenzenes in fish. Hawker and Connell (1985a, b) suggested that half a year may be required to obtain the steady state for compounds with a log P of around 6 and about 10 years for those with a log P of about 8, and formulated a QSAR similar to Eqs. 28 and 29 for the BCF value under nonequilibrium conditions with a certain exposure period. Devillers et al. (1996) compared in detail the versatility of the bilinear model expressed by Eq. 30 with that of linear, quadratic, and polynominal correlation models published from different organizations. They selected 342 log BCF values for 181 compounds, which are mostly inert. Some values were independently measured in duplicate for a single compound. The selection criteria for these log BCF values required that the BCF data are obtained only after a steady state was established and
29.4 Physicochemical Significance of log P in Environmental Quantitative Structure-Activity Relationships
that, if one or more values appeared out of line in a publication, all the data contained in that publication are not used. They concluded that the bilinear model represented by Eq. 30 is among the best in predictive performance in terms of the root mean square (rms) value for residuals between log BCF values experimentally measured and calculated from model equations. Banerjee and Baughman (1991) proposed another model for the nonlinearity of log BCF versus log P. They tried to rationalize the breakdown of the linear relationship not only for the highly hydrophobic compounds, but also for many azo dyes (multifunctionalized azobenzenes), the log P value of which is below 6 (mostly between 3.5 and 4.5) (Anliker and Moser, 1987). Their model considered the fact that large compounds of low lipid solubility such as azo dyes and polyhaloaromatic hydrocarbons, have lower than expected BCF values because of the difficulty in cavity formation in lipids. Thus, with an approximation in which the lipid solubility of compounds (generally unavailable) is replaced with the solubility in octanol, Soct (M), they proposed the following equation for a set of compounds. These include inert pesticides, aromatic and aliphatic (halogeno)hydrocarbons, and polar but mostly nonionized azo dyes, the log P values of which ranged from 1.5 to 8.3. log BCF
= -1.13 + 1.02 log P +0.84 log Soct + 0.0004(mp - 25), n
= 36,
(31)
r = 0.95.
In Eq. 31, mp is the melting point (in 0c), which was intended to allow octanol solubilities for both liquids and solids to be included in a single equation (Valvani and Yalkowsky, 1980). For liquids, mp is regarded as 25 to remove the entire term. For small compounds, the log Soct and (mp - 25) terms in Eq. 31 tend to be constant and Eq. 31 takes the same form as Eq. 29. Most of the compounds included in Eq. 31 are solids. Without the log Soct and (mp - 25) terms, the correlation was much poorer (r = 0.73). In the region of log P > 6, even though the log P increases, the BCF value could decrease because of a decrease in the lipid (octanol) solubility, resulting in the descending phase in the log BCF/ log P relationship. The previous relationships seem to be in accord with the low fish toxicity of an insecticide of nonester-type pyrethroids, silaftuofen (IX) (Sieburth et aI., 1990). The log P value of this compound has been estimated as being about 10 (Okimoto et aI., 1994). Its uptake in fish should well be very low.
663
1989; Oliver and Niimi, 1985; Opperhuizen and Voors, 1987). Such deviations from Mackay's postulate (Mackay, 1982) have been attributed to relatively high rates of biotransformation (de Bruijn and Hermens, 1991; Opperhuizen and Voors, 1987; Southworth et aI., 1980). Uchida et al. (1982a) measured the BCF value of isoprothiolane and its analogs (VI) with the killifish (Orizias latipes) and proposed the following: 10gBCF = 0.65(±0.17)logP -1.17(±0.62),
s
n =9,
= 0.197,
r
= 0.962,
F
(32)
= 85.5.
The reason that the coefficient of the log P term is smaller than unity is attributable to the fact that the test compounds have two hydrolyzable ester groups. Uchida et al. also measured the rate of disappearance of the compounds in the entire system. The rate followed approximately the zero-order kinetics, and the rate constant, logk, tended to increase with decreasing bulkiness of the ester substituents in terms of the STERIMOL width (Verloop, 1983) and with increasing log P value. This was taken to indicate that the hydrolytic degradation would occur under conditions "reacting" with fish. The "apparent" log BCF term in Eq. 32 should be compensated for the degradation effect corresponding to the log k value so that the size of the log P term under the "real" conditions would be closer to unity. de Bruijn et al. (1993) examined the effect ofbiotransformation of a set of organophosphorus insecticides on bioconcentration. The compounds belonged to O,O-dimethyl-O-phenyl phosphorothioates in which various substituents such as CN, N02, SMe, and halogens are located at the 2-, 4-, and 5positions of the benzene ring either singly or multiply. Their log BCF values were measured using guppies, from which the following equation was derived: 10gBCF = 0.80(±0.12)logP +0.45, (33) n =
12,
s = 0.35,
r = 0.910.
One of the most important metabolic pathways of dimethyl phosphorothioates has been shown to be the demethylation of one of the two methyl groups by glutathione (Fukami and Shishido, 1966). Thus, the rate of demethylation under pseudofirst-order conditions, k (min- l mg protein-I), was measured using a glutathione S-methyltransferase preparation. Introduction of the log k term into Eq. 33 yielded the following: log BCF = 0.94(±0.08) log P -0.63(±0.14) logk - 2.31,
n = 12,
For less inert compounds, including pesticides in which certain reactive/vulnerable functions are required for their biological activity, the slope has been observed to be significantly lower than unity. Moreover, the correlation is often of a lower statistical quality (de Wolf et aI., 1992; Niimi et aI.,
s = 0.21,
(34) r =
0.971.
The improvement in the correlation quality is significant and the slope of the log P term is close to unity in Eq. 34. The preceding two examples indicate that, if an appropriate term for the vulnerability is incorporated, then Mackay's postulate is valid even for the unstable series of pesticides. Mackay's postulate has also been observed for bioconcentration in mollusks, daphnias, and aquatic microbes (Baughman and Paris, 1981; Geyer et al., 1982; Hawker and Connell, 1986).
664
CHAPTER 29
Hydrophobicity as a Key Parameter of Environmental Toxicology
The structure-activity relationships for biomagnification are not as well established as those for bioconcentration. A number of mechanisms, which are not well understood, are involved in the entire process of biomagnification, which occurs through the food chain. However, it is highly probable that the biomagnification is also able to be significantly related to the octanol/water partition coefficient (Esser, 1986). Davies and Dobbs (1984) indicated that the uptake of chemicals from both food and water results in tissue concentrations comparable to those resulting from water alone. It should be mentioned that the uptake of pesticides from food is far less important than the uptake from water and that only a part of the residue present in the lower level biota is transferred to the higher level of the food chain (Ellgehausen et aI., 1980).
(halogen-substituted) a1cohols, esters, ketones, phthalates, substituted benzenes, and several pesticides (mostly rather stable herbicides). For inert organic compounds, the log P value of which is very high (>5-6), the pLC50 value has been observed to be lower than expected by such linear correlations as Eqs. 35 and 36, probably because of the limit in the water solubility (solubility cutoff) to be partitioned into the fish body required for the symptom of toxicity. This situation was confinned by Veith et al. (1983), who formulated a bilinear correlation of excellent quality such as the following for the 4-day toxicity to the fathead minnow of five classes of unreactive compounds similar to those included in Eq. 36: pLC 50 (M) = 0.94 log P
-0.94log(6.8 x 10-5 P
29.4.3 AQUATIC TOXICITY
n = 65,
The aquatic toxicity of "simple" organic compounds has been recognized as being closely related to their lipophilicity (hydrophobicity). Overton and Meyer independently proposed the "lipoid theory of narcosis" about 100 years ago (Lipnick, 1989), narcosis being considered as a toxic effect that can be lethal. The modern fonnulation of the aquatic toxicity in terms of the QSAR was, however, first published by Hansch and Dunn (1972). They found that the earlier toxicity data of various sets of homologous a1cohols or miscellaneous inert compounds, in terms of the minimum narcotic and the minimum lethal concentrations, can mostly be explained by a single parameter log P with a slope close to unity. Subsequently, a number of hypotheses of aquatic toxicology were combined with the QSAR concept. Thus, K6nemann (1981) investigated the QSAR of "environmental pollutants" of structurally heterogeneous organic compounds such as methyl- and chlorobenzenes, aliphatic chlorohydrocarbons, and a1cohols, among others. For the 14-day LCso (M) value in the guppy (Poecilia reticulata), the following equation was derived: pLCso n =
= 0.871 log P + 1.13,
50,
r
= 0.988,
(35) s = 0.237.
In this set of compounds, the pLC50 (M) value varies from 6.15 for pentachlorobenzene (log P = 5.69) to 0.24 for diethyleneglycol (log P = -1.30). The selection of compounds included in Eq. 35 was intended so that they are rather stable and unreactive (inert) as well as un-ionizable. Thus, pesticides that are "reactive" to "specific" targets in vivo are not included. With a larger (4 day) toxicity data set against juvenile fathead minnows (Pimephales promelus) from the V.S. Environmental Protection Agency, Duluth, Minnesota, the following equation was derived (McCarty et aI., 1992): pLC50 (M) = 0.90(±0.04) log P n = 150,
r = 0.959.
+ 1.29(±0.12), (36)
The compounds included in Eq. 36 are inert and cover the range of log P similar to that in Eq. 35. They are rather stable, belonging to halogenated aliphatic hydrocarbons, ethers,
+ 1) + 1.25,
(37)
r = 0.999.
Although the detailed mechanism of the acute fish toxicity is not completely clear, the observed symptoms caused by inert and unreactive compounds strongly suggest a mechanism categorized as general narcosis (Lipnick, 1990). In Eqs. 3537, the coefficient of the log P term is close to unity and the intercepts are almost equivalent to one another irrespective of the test organism and the range of test compounds. A number of similar correlations have been accumulated for sets of inert compounds (Cronin and Dearden, 1995; Ikemoto et aI., 1992; Lipnick, 1990). Thus, the general narcotic (analgesic) potency is dependent only on the overall hydrophobicity and not on the specificity of the chemical structure. Because every compound is supposed to exert at least this type of narcotic activity, the QSAR correlations similar to Eqs. 35-37 are considered to predict the minimum toxicity or the "baseline toxicity" (Lipnick, 1990) of any organic compound to aquatic biota, unless the compound is biodegradable and therefore less toxic. The toxic effects of mixtures of compounds from a single group in terms of the mechanism (type) of toxicity should be concentration additive. Concentration addition means that the LC50 of a mixture is observed at the sum of concentrations, c, of individual compounds (as the fraction of their own LC50) being 1.0 (L c /LC50 = 1.0). The compounds included in Eq. 35 have been shown to be concentration additive, that is, to exert their effect by an equivalent mechanism (K6nemann, 1980). Veith and Broderius (1987) noted that a number of compounds that appear to produce narcosis are significantly more toxic than the baseline toxicity. They are more polar and often have weakly acidic and/or hydrogen-bond donor groups, such as substituted phenols, mostly existing as the nonionized fonn under experimental conditions, and substituted anilines. This structurally heterogeneous set of compounds has been examined by the concentration additivity test, showing that these phenols and anilines indeed belong to a group exerting a narcotic syndrome [Type 11 (or polar) narcotic syndrome] differing from that of the unreactive inert compounds [Type I (or nonpolar) narcotic syndrome]. For phenols and anilines substituted both singly and multiply, by alkyl, alkoxy, halogen, N02, the
29.4 Physicochemical Significance of log P in Environmental Quantitative Structure-Activity Relationships
following equation was formulated with the 4-day LCso value against juvenile fathead minnows: pLCso(M) = 0.65(±0.07) log P n =39,
r
+ 2.29(±0.22),
= 0.95.
(38)
Prior to this, Saarikoski and Viluksela (1982) analyzed the 4-day LCso value of variously substituted phenols against guppies (P. reticulata) at pH 6, 7, and 8. The substitution pattern of phenols was carefully selected so that the collinearity between the log P and 1'1 log Ka values is as low as possible. 1'1 log Ka, the difference between the log Ka and the reference log Ka of unsubstituted phenol, was used as a parameter for the electron-withdrawing effect of substituents. They also corrected for the effect of ionization on the LCso value in terms of an "effective" concentration of the neutral form defined empirically (Saarikoski and Viluksela, 1981). For phenols substituted with Me, Cl, t-Bu, OMe, OH, and N02 groups singly or multiply, Saarikoski and Viluksela derived the following equations: pLCso (M; corrected for pH 6) = 0.67 log P + 0.191'1 log Ka n =
19,
r = 0.985,
+ 2.33, s
(39)
= 0.17,
pLCSO (M; corrected for pH 7)
= 0.71 log P + 0.191'1 log Ka + 2.23, n = 19,
r = 0.978,
(40)
s = 0.21.
Because Eqs. 39 and 40 are practically identical, they can serve to predict the toxicity at any pH from 6 to 8. It is interesting to note that, for phenols, the 1'1 log Ka value of which is so low that the effects of ionization on the LCso and electron withdrawal of substituents are not great, Eqs. 39 and 40 are very similar to Eq. 38. For phenols with more acidic hydrogen than anilines, the electron-withdrawing effect of substituents on the OH, which may hydrogen-bond with possible taget site(s), seems to be accounted for explicitely in Eqs. 39 and 40. Equations 38-40 probably reflect the feature of Type 11 (or polar) narcotics, in which the slope of the log P term is lower and the intercept is higher than those of the nonpolar narcotics. The higher intercept may result from the hydrogen-bond donor group in phenols and anilines enhancing the interaction with the target site(s). The lower slope could mean that there is an increasing shift toward nonpolar narcosis with increasing hydrophobicity. For nonpolar narcosis, the slope is higher. However, the toxity of non polar narcotics is lower than that of polar narcotics in the region of low log P values. Thus, the shift of the narcosis type could make the slope higher. A possible reason for this may be related to differences in distribution so that compounds with a high log P value will show an increasing tendency to accumulate in the lipid phase (Cronin and Dearden, 1995). In Eq. 38, such phenols as Cls- and 2,4-(N02h-derivatives are not included because these compounds do not share the concentration additivity with those included. In Eqs. 39 and 40, phenols with 2,5- and 2,4-(N02h-substitution patterns are not included because of their much higher toxicity than that
665
predicted. The structural characteristics of these outliers are shared by the pesticidal phenols mostly acting as uncouplers with the mitochondrial oxidative phosphorylation. They are neither Type I nor Type II narcotics, but exert a toxicity about lO-fold higher than that predicted by the Type II correlation equations. A number of compounds with biologically and/or chemically "reactive" sites have toxicities considerably greater than those predicted for either nonpolar or polar narcosis. Various pesticides with specific structural features and/or specific functional moieties to inteact with respective target site(s) are such compounds not categorized into narcotics. These compounds are often defined to show "reactive toxicity" (Hermens, 1990). Ikemoto et al. (1992) examined the excess toxicity exerted by various pesticides against killifish (Oryzias lapipes) over the baseline toxicity which was formulated for nonpolar inert compounds such as homologous alkanols, chlorobenzenes, and alkylbenzenes. Photosynthesis-disrupting herbicides, such as diuron, and chitin-synthesis-disrupting larvicides, such as diflubenzuron (IV: Xl = X2 = F, Y = Cl) and buprofezin (VII) were almost on the line for the nonpolar narcotic compounds. Their targets do not exist in fish. Such neurotoxic insecticides as DDT, dieldrin, fenvalerate, and lindane are more toxic than the baseline by 1.2-2.5 log units, and the respiration-inhibiting rotenone is higher than 3 log units more toxic. The neurotoxicity and respiratory toxicity are probably common between fish and insects. Besides the specific biochemical mechanisms exerted by pesticides, a variety of chemical reactivity mechanisms such as electro- and nucleophilic, redox, and free-radical processes are thought to be involved in the interactions of various types of toxicants with biological systems. Especially important are toxicants expected to work as electrophiles which can react with nucleophilic groups such as NH2, OH, and SH in deoxyribonucleic acid (DNA) and proteins. Hermens (1989) classified the reaction mechanisms of nucleophilic groups in biological systems with electrophilic toxic ants into (a) nucleophilic displacement reaction, (b) addition to carbon-oxygen double bonds, and (c) addition to activated carbon-carbon double bonds (the Michael-type addition). He also surveyed various molecular substructures present in possible toxicants where these types of reactions might be responsible for their unwanted activity (Hermens, 1990). In this type of reactive toxicity, one should not expect simple relationships between toxicity and hydrophobicity even when accompanied by electronic parameters such as those represented by Eqs. 35-40. For instance, Hermens et al. (1985) derived the following equation for the 14-day toxicity to guppies of 15 alkyl, alkenyl, acylmethyl, and benzyl halides: pLCso(M) = 0.224 log P - 1.3210g(2484 + k- l ) + 10.05, n
= 15,
r
= 0.956,
(41)
s = 0.39.
The pLCso value was only very poorly correlated with the log P value alone (r = 0.41), but the correlation was much improved
666
CHAPTER 29
Hydrophobicity as a Key Parameter of Environmental Toxicology
when the pseudo-first-order rate constant, k (in min- 1), with a model nucleophile (4-nitrobenzylpyridine) was included. The second term on the right-hand side of Eq. 41 means that the variation in the toxicity is biphasic with respect to log k. For compounds with very low reactivity, that is, for those with very small k or very large k- 1 values, the value of 2484 within the parenthess can be neglected relative to the k- 1 value. Therefore, the value of the negatively signed second term, being approximated by + 1.3210g k, initially increases nearly linearly with increasing logk, but, in the region above logk > -3.4 (k > 1/2484), follows a plateau-like pattern. Examples of the use of experimentally estimated reactivity indices have been reviewed by Hermens (1990), who also analyzed the 14-day LCso values of epoxides and aldehydes and formulated correlations somewhat similar to Eqs. 38 and 41, respectively. In the reactive toxicants, the term "reactive" encompasses a wide spectrum of chemical processes as mentioned previously. Further classification must be made for hard (charge controlled) and soft (orbital controlled) interactions (Comporti, 1989) because relevant molecular descriptors are different between these interactions. Soft electrophiles cover one of the largest groups of anthropogenic toxic ants exceeded only by narcotics. Soft electrophilic interactions of toxicants with biomolecules are regulated by the ability of toxicants to accept electron density represented by the "nucleophilic" or electron-acceptor superdelocalizability, SN (the term "nucleophilic" is from the side of "biological" reactants), through the orbital interactions. The SN value on the activated unsaturated carbon atoms and the frontier charge of the lowest unoccupied molecular orbital, f(LUMO), along with the log P value, are important in predicting the toxicity of such soft (pro )electrophiles (Mekenyan and Veith, 1993). The "proelectrophilic" mechanism of the toxicity refers to compounds that are not direct-acting electrophiles but are metabolized to electrophiles such as primary and secondary allyl and propargyl alcohols (Lipnick et aI., 1985). Veith and Mekenyan (1993) examined to extend this approach to a large set of n -electron systems such as aromatic compounds, including variously substituted hydrocarbons, phenols and ani lines, and unsaturated aliphatic compounds, including alkenols and alkynols. These were selected to represent a wide variation in n bonding and a variety of modes of toxic actions. The compounds include polar and nonpolar narcotics, uncouplers with oxidative phophorylation, and "reactive" electrophiles and proelectrophiles. Veith and Mekenyan derived the following equation for the 4-day LCso value of 114 compounds against fathead minnow: pLCso(M) = -1.49(±0.53) + 0.56(±0.04) log P +13.7(±1.7)SN(av.), n
= 114,
r
= 0.90,
s
= 0.43,
(42)
toxicity of proelectrophiles, when the SN(av.) value is calculated for their metabolites. The authors (Mekenyan and Veith, 1993) further showed that the soft electrophiles included in Eq. 42 can be clustered according to their SN values by defining "isoelectrophilic windows" along the toxicity response plane. Nonpolar narcotics are located in the lowest SN region where toxicity varies almost only with hydrophobicity. Polar narcotics are more toxic than nonpolar narcotics at similar values of log P and the toxicity increase can be illustrated by higher SN values (by stronger electronic interactions with cellular soft nucleophiles). Highly reactive soft electrophiles, which have dissociable protons, act as uncouplers. Electrophiles without dissociable protons indicate symptoms of reactive toxicity consistent with covalent bonding. Perhaps because of their high specificity, the aquatic toxicology of pesticides has not been studied intensively in terms of the QSAR. There have been studies for sets of miscellaneous pesticides, but their toxicity is usually dealt with as that of QSAR outliers with higher potency than narcotic toxicants (Ikemoto et aI., 1992). Exceptions are the organophosphorus insecticides (Hermens et aI., 1987; Schiiiirmann, 1992) where a series of O,O-dimethyl phenylphosphorothioates are substituted at the 2-, 4-, and 5-positions on the benzene ring by various groups such as H, CN, N02, Me, halogens, and SMe. The 14-day LCso toward guppies were measured and analyzed for 12 analogs to give the following (Verhaar et aI., 1994): pLCso(M) = 0.38(±0.12) log P - 27.9(±7.78)p(PO) + 11.0(±5.43), n = 12,
r
s = 0.371,
F
= 12.9.
In Eq. 43, p(PO) is the bond order between the central phosphorus atom and the phenoxy oxygen calculated by MOPAC 6.0 (Stewart, 1990). This descriptor was believed to contain information about the strength of the corresponding bond reflecting the nature of the phenoxide as the leaving group and thus to have relevance for the phosphorylation step of acetylcholinesterase inhibition. The negative sign of this term suggests that the lower the PO bond order, the more easily the phosphorylation of the serine OH of acetylcholinesterase occurs, leading to the higher toxicity. The preceding study originally analyzed the LCso data with the use of experimentally estimated reactivity indices such as the rate constant k with the model nucleophile, 4nitrobenzylpyridine (Hermens et aI., 1987). The following equation was fotmulated for 10 analogs (two compounds were added later to formulate Eq. 43): pLCso(M)
= 0.34(±0.1O)
L:n
+0.76(±0.19) logk - 3.57,
F = 238.7.
SN(av.) is the superdelocalizability averaged for the values assigned to atoms included in the conjugated n -electron system. It is expected to represent the global electron-acceptor character of the molecule. The log P and SN (av.) parameters are orthogonal for these compounds. Equation 42 accurately predicts the
= 0.861,
(43)
= 0.29. same set of compounds, the use of L CT n
= 10,
r = 0.912,
(44)
s
instead of With the log k yielded a poorer correlation. The significant contribution to toxicity by the demethylation rate, k, can be taken as a possible (but not definite) hint that the in vivo demethylation may be
References
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ACKNOWLEDGMENTS The authors would like to dedicate this chapter to Professor Corwin Hansch of Pomona College, Clare mont, California, on his 80th birthday. They would like to express their sincere thanks to Dr. Albert Leo of Pomona College for his invaluable discussions as well as for his careful review of the manuscript.
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Hansch, C., and Fujita, T. (1995). "Classical and Three-Dimensional QSAR in Agrochemistry," ACS Symposium Series No. 606. Am. Chem. Soc., Washington, DC. Hansch, c., and Leo, A. (1979). "Substituent Constants for Correlation Analysis in Chemistry and Biology." Wiley, New York. Hansch, c., and Leo, A. (1995). "Exploring QSAR: Fundamentals and Applications in Chemistry and Biology." Am. Chem. Soc., Washington, DC. Hansch, C., Leo, A., and Hoekman, D. (1995). "Exploring QSAR: Hydrophobic, Electronic, and Steric Constants." Am. Chem. Soc., Washington, DC. Hansch, c., Quinlan, J. E., and Lawrence, G. L. (1968). The linear free-energy relationship between partition coefficients and the aqueous solubility of organic liquids. J. Org. Chem. 33, 347-350. Hansch, c., Vittoria, A., Silipo, C., and Jow, P. Y. C. (1975). Partition coefficients and the structure-activity relationship of the anesthetic gases. J. Med. Chem. 18,546-548. Hawker, D. W., and Connell, D. W. (1985a). Relationships between partition coefficient, uptake rate constant, clearance rate constant and the time to equilibrium for bioconcentration. Chemosphere 14, 1205-1219. Hawker, D. W., and Connell, D. W. (1985b). Prediction of bioconcentration factors under non-equilibrium conditions. Chemosphere 14,1835-1843. Hawker, D. W., and Connell, D. W. (1986). Bioconcentration oflipophilic compounds by some aquatic organisms. Ecotoxicol. Environ. Sa! 11, 184-197. HeIling, C. S. (1971). Pesticide mobility in soils. Ill. Influence of soil properties. Soil Sci. Soc. Am. Proc. 35, 743-748. Hermens, J. L. M. (1989). Quantitative structure-activity relationships of environmental pollutants. In "Handbook of Environmental Chemistry" (0. Hutzinger, ed.), pp. 111-162. Springer-Verlag, Berlin. Hermens, J. L. M. (1990). Electrophiles and acute toxicity to fish. Environ. Health Perspect. 87, 219-225. Hermens, J., Busser, E, Leeuwanch, P., and Musch, A. (1985). Quantitative correlation studies between the acute lethal toxicity of 15 organic halides to the guppy (Poecillia reticulata) and chemical reactivity towards 4-nitrobenzylpyridine. Toxicol. Environ. Chem. 9, 219-236. Hermens, J., de Bruijn, J., Pauly, J., and Seinen, W. (1987). QSAR studies for fish toxicity data of organophosphorus compounds and other classes of reactive organic compounds. In "QSAR in Environmental Toxicology-II" (K. L. E. Kaiser, ed.), pp. 135-152. Reidel, Dordrecht. Ikemoto, Y., Motoba, K., Suzuki, T., and Uchida, M. (1992). Quantitative structure-activity relationships of nonspecific and specific toxicants in several organism species. Environ. Toxicol. Chem. 11,931-939. Isnard, P., and Lambert, S. (1988). Estimating bioconcentration factors from octanol-water partition coefficient and aqueous solubility. Chemosphere 17, 21-34. Iwasa, J., Fujita, T., and Hansch, C. (1965). Substituent constants for aliphatic functions obtained from partition coefficients. J. Med. Chem. 8, 150-153. Karabunarliev, S., Mekenyan, O. G., Karcher, W., Russom, C. L., and Bradbury, S. P. (1996a). Quantum-chemical descriptors for estimating the acute toxicity of electrophiles to the fathead minnow (Pimephales promelas): An analysis based on molecular mechanisms. Quant. Struct.-Act. Relat. 15, 302-310. Karabunarliev, S., Mekenyan, O. G., Karcher, W., Russom, C. L., and Bradbury, S. P. (1996b). Quantum-chemical descriptors for estimating the acute toxicity of substituted benzenes to the guppy (Poecilia reticulata) and fathead minnow (Pimephales promelas). Quant. Struct.-Act. Relat. 15,311-320. Karger, B. L., Gant, J. R., Hartkopf, A., and Weiner, P. H. (1976). Hydrophobic effects in reversed-phase liquid chromatography. J. Chromatogr. 128, 6578. Kataoka, T., Hayase, Y., Hatta, T., Hayashi, Y., Murabayashi, A., Makisumi, Y., and Fujita, T. (1989). Quantitative structure-activity study of fungicidal I-substituted cis-2-(I H -1 ,2,4-triazol-I-yl)cycloalkanols. Pestic. Biochem. Physiol. 34, 228-239. Kaufman, J. J., Semo, N. M., and Koski, W. S. (1975). Microelectrometric titration measurement of the pKa's and partition and drug distribution coefficients of narcotics and narcotic antagonists and their pH and temperature dependence. J. Med. Chem. 18,647-655. Keller, C. (1993). "Grundlagen der Radiochemie," 3rd ed., p. 247. Otto Sal1e Verlag, Frankfurt am Main; Japanese Translation (1993). "Houshakagaku
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CHAPTER
30 Modern Approaches to Analysis of Pesticide Residues in Foods and the Environment Luis O. Ruzo PTRL West, Inc.
Thomas Class PTRLEurope
30.1 INTRODUCTION The capability to determine the actual residues of pesticides, and increasingly of their metabolites, is at the core of the complex and sometimes bewildering regulatory processes in all industrialized nations. Thus, in the United States we are currently experiencing the replacement of laws based on long-term toxicological effects, such as the Delaney Clause, with those based on quantitative determinations that lead to a different type of risk assessment, as is the Food Quality Protection Act of 1996 (FQPA). To understand the benefits and limitations of modem analytical techniques, it is necessary to examine the regulatory requirements that they aim to satisfy. Residue chemistry data are used by regulatory agencies to estimate the exposure of the general population, as well as discrete subpopulations, to pesticide residues in food and water and for setting and enforcing tolerances for such residues in food crops or animal feed. These data are also used to monitor environmental contamination of soil, air, and water and thus to determine adverse effects that may arise from transport of residues between these compartments. The tolerance values are a key component of the regulatory equation as they represent the amounts of pesticiderelated materials legally allowed to be present in a given matrix. Thus, tolerances are legally enforceable limits that are currently under scrutiny for implementation of the FQPA. The passage of this law by the U.S. Congress in 1996 signaled a fundamental change in the way exposure to pesticides is evaluated by introducing the concept of aggregate exposure to several compounds of a given chemical class and mode of action. These expanded requirements will necessitate the development of new methodHandbook of Pesticide Toxicology Volume 1. Principles
ologies for analysis, specifically to address the lower limits of quantitation (LOQs) (Ragsdale, 1998). A similar situation is observed in the European Community (EC). Here (re-)registration of pesticides and enforcement of tolerances or maximum residue limits (MRLs) requires that existing multiresidue methods be assessed for their applicability toward the determination of active substances and their toxicologically relevant metabolites.
30.2 METHOD VALIDATION Because this chapter will deal primarily with analytical techniques now in use for developing methodologies, it is worth reviewing briefly the criteria that enforcement agencies use to evaluate method performance. Methods used by government laboratories are generally developed by the pesticide registrants. A method is considered acceptable upon validation. This process may entail the examination of various parameters but always includes the establishment of a limit of quantitation (LOQ) and a limit of detection (LOD). The former is usually set at the lowest matrix fortification level for which acceptable, quantifiable recoveries of the analyte(s) are obtained.
30.2.1 CONFIRMATION OF THE ANALYTICAL METHOD This entails the fortification of untreated (control) matrices (crops, soil, water, etc.) with varying concentrations of analyte. Thus, the processed sample is fortified at two or three levels in duplicate or triplicate (United States) or in five replicates (EC). The fortified samples and corresponding untreated controls are
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then subjected to extraction, cleanup, and chromatographic separation and quantitation. The ultimate recovery of analytes must be in the 70-120% (70-110% in Europe) range based on the initial concentration, with repeatability demonstrated by relative standard deviations of Q 20%. The lowest concentration for which this is achieved during validation is generally considered the limit of quantitation or determination (LOQ) for the method. The limit of detection (LOD) may be any value below the LOQ at which an analyte signal is clearly distinguished from background signals present or absent in matrix extracts. At present, methods aim at an LOQ of 0.01 ppm (mgikg) for most foodstuffs, whereas much lower values are considered desirable for water analyses, where the LOQ Q 0.05 I-Lg/I (ppb) is required, or for compounds of known toxicological significance. Important aspects to be considered in the conduct of validation studies are provided by lenke (1996a-c).
30.2.2 INDEPENDENT LABORATORY VALIDATION
For a method to be used in the enforcement of tolerances or MRLs, it must be rugged and straightforward. Therefore, the U.S. Environmental Protection Agency (EPA) and the EC require that such methods be validated independently by laboratories that are not familiar with the procedures or analytes (EC Directorate General for Agriculture, 1998; EPA, 1996). Because the equipment necessary for the enforcement method must be generally accessible and affordable for the enforcement laboratories, some of the most advanced techniques (such as MSIMS, or in the European Community LC-MS) are still not acceptable to regulatory agencies. Measures to demonstrate the validity of results obtained by enforcement methods include round-robin testing using identical samples with fortified or incurred residues, which allow assessment of the reproducibility of commonly employed enforcement methods.
30.2.3 METHOD RADIOVALIDATION (EPA,1996)
A stringent test for an analytical method is its reproducibility when applied to incurred residues as opposed to an externally fortified matrix. Plant and animal metabolism studies utilizing 14C-Iabeled pesticides generate matrices containing incurred residues that can be readily quantified with radiochemical methodology, which is quite different (and simpler) than that generally developed for an analytical residue method. In order for a method to be considered fully validated, the results obtained when the "cold" method is applied to matrices containing 14C-Iabeled incurred residues must agree closely with the results arising from quantitation of the radiocarbon conducted by radiochemical methods such as liquid chromatography and liquid scintillation counting.
30.3 DEVELOPMENT OF THE ANALYTICAL METHOD 30.3.1 MULTIRESIDUE METHODS
Multiresidue, multiclass methods are generally the most cost effective overall approach for pesticide analysis in foods, soil, and water. Regulatory enforcement methods routinely deal with multiresidue determinations under fairly standardized conditions. For example, the widely used DFG S19 method (Thier and Zeumer, 1987) establishes conditions for extraction and quantitation of organochlorine, organophosphorous, and nitrogencontaining pesticides (typically 80-100 compounds per method) in crop plants under standardized conditions. The original method (Specht et al., 1995) uses an acetone/ethyl acetate!cyclohexane mixture for extraction and partition, thus replacing dichloromethane and introducing a "one-beaker" extraction and partition procedure. In addition to use with watery plant matrices, the DFG S19 method has also been successfully employed for dry plant matrices (straw, hops, tobacco, herbal teas), for oily crops (oil seed rape, sunflower seeds, nuts, etc.), for animal matrices (milk, whole egg, muscle, fat), and for soil. Thus, a single solvent (acetone) with fixed amounts of added water (in a ratio of 2: 1 solvent water) or acetonitrile (for oily matrices) is used for extraction of compounds with a wide range of polarity. Cleanup involves gel permeation chromatography (GPC) and fractionation on a small silica gel column followed by separationlquantitation with capillary gas chromatography equipped with selective detectors. The Netherlands' Inspectorate for Health Protection has recently published in its 6th edition (1996) an extended collection of multiresidue methods for a multitude of pesticides: The Netherlands' multiresidue method 1 (MRM-1) covers all pesticides that can be analyzed by capillary gas chromatography using selective detectors [electron-capture detector (ECD), nitrogen-phosphorus detector (NPD), flame photometric detector (FPD)] and increasingly full-scan ion trap mass spectrometric detection. Three submethods described in The Netherlands' MRM-2 use high-performance liquid chromatography (HPLC) and postcolumn derivatization with fluorescence detection for N-methylcarbamate (including metabolites) and phenylurea pesticides or precolumn switching employing a precolumn packed with internal surface reversed-phase material for chlorophenoxy herbicides. The latter class of compounds is also analyzed more selectively by The Netherlands' MRM3 after derivatization with pentafluorobenzyl bromide (PFBBr) as esters. Derivatives of aromatic amines are covered by The Netherlands' MRM-4 (two submethods) and by the German DFG S6 and S6-A methods, all of which use alkaline hydrolysis and steam distillation of the amines, followed by various derivatization procedures, and gas chromatography. The Luke method is a multiresidue method currently employed in the United States for enforcement of tolerances and import tolerances. Its cleanup includes mainly solid-phase extraction (SPE) cartridges of various selectivities with either gas or liquid chromatography of pesticides in separate fractions.
30.3 Development of the Analytical Method
However, because of the current emphasis on metabolites of toxicological concern, target methods are generally developed with an individual or closely related compounds in mind. Frequently, this involves simultaneous determination of the parent pesticide and its metabolites, which are generally of greater polarity (because they typically arise from oxidation or hydrolytic cleavage reactions). 30.3.2 EXTRACTION Generally, extraction methodology must be developed such that nearly quantitative recovery of target analytes is obtained. At this stage of method development, the use of radiolabeled standards is invaluable because it allows for rapid determination of the percentage extractability by direct liquid scintillation counting (LSC). In later steps, the radiotracer is useful in determining efficiencies for each step of the proposed method. Traditional solvent extraction must take into account the chosen solvent's water miscibility, ease of solvent concentration/removal, safety, and disposal costs. DFG S 19 uses water/acetone (100 ml1200 ml) or acetonitrile/acetone for oily matrices and allows a sample size ranging from 109 of very dry material (straw, herbs) to 100 g for watery matrices. The Netherlands' multi methods use mainly acetone or ethyl acetate for extraction; other conventional extraction systems include methanol or acetonitrile. Some methods use combined extractionlhydrolysis steps either for deconjugation or to form common moiety products, which may be separated from the extraction mixture by steam distillation (e.g., The Netherlands' MRM-4 and the DFG S6 for derivatives of aromatic amines). Several new extraction approaches are being developed such as supercritical fluid extraction (SFE) and pressurized liquid extraction (also known as accelerated solvent extraction, ASE). Typically, SFE gives very clean extracts with somewhat low recoveries and ASE gives good recoveries but the samples are exposed to high temperatures and pressures and require more extensive cleanup to remove co-extracted matrix components. The latter techniques allow only the extraction of relatively small amounts of samples (e.g., 5-10 g), which requires an increased emphasis on sample preprocessing and homogenization (to assure homogeneity) if market sample size amounts to 5-10 kg. SFE offers potential advantages for removing trace levels of target agrochemical ana1ytes from various matrices. Of particular interest are the enhanced extraction rates obtained due to the high diffusivity of critical fluids. In addition to rapid extraction, improved penetration of the matrix with subsequent high recovery of bound residues is feasible (Hawthorne et aI., 1992; Lira, 1988; Lopez-Avila and Dodhiwala, 1990). However, the success of commonly used fluid systems such as carbon dioxide or polar-modified carbon dioxide in binary mixtures is limited to, at best, moderately polar analytes. Because many factors can influence SFE efficiency (Erstfeld and Chen, 1998; Fahmy et aI., 1993; Snyder et aI., 1992, 1993), including pressure, temperature, fluid flow rate, extraction time, and modifier, considerable effort must be invested
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during method development before choosing SFE as the extraction technique. However, there is evidence that at least in some cases (as with chlorothalonil) it compares favorably to traditional approaches such as Soxhlet extraction (Erstfeld and Chen, 1998). Thorough descriptions of the instrumentation utilized by SFE are available in the literature (McNally et aI., 1992; Riekkola et aI., 1992). Excellent short reviews of SFE applications are those of Bond (1994) and King (1989). Supercritical fluid extraction has been reported to be successful in multiresidue analysis with pyrethroids (Argauer et al., 1997), organophosphates (Skopec et aI., 1993), and other compound classes (Jones, 1996, 1997; King et aI., 1993). In fact, even the hydrophobic avermectins are amenable to SFE techniques (Brooks and Uden, 1995). There is general agreement among researchers that although the scope of applicability for SFE may be limited, great advantages are provided in selected cases by the cleaner extracts obtained. Accelerated solvent extraction (ASE) is rapidly gaining acceptance as an alternative extraction approach with positive results in magnitude of residue, multiresidue, soil dissipation, and animal health studies (Stanek and Keller, 1998). ASE is based on the use of a variety of solvents under elevated temperatures and pressure. The higher temperatures involved accelerate the kinetics (as with Soxhlet) and the elevated pressure keeps the solvent in the liquid phase. The technique is amenable to automation (Ezzell, 1998). Traditionally, compounds that are difficult to extract from environmental matrices have been subjected to reflux conditions such as Soxhlet extraction. This approach is wasteful in time and solvent use. In fact, ASE compares favorably with Soxhlet and supercritical fluid extraction (David and Seiber, 1996; Frost et aI., 1997; Lou et aI., 1997). ASE is most effective with thermally stable low- or medium-polarity substrates. Extraction results are often better than those obtained with traditional methods (Conte et aI., 1997; Ezzell et aI., 1995). Sample preparation is quite simple (Richter et aI., 1996), involving grinding and mixing of the soil followed by air drying or admixture with drying agents (e.g., sodium su1fate). 30.3.3 CLEANUP OF EXTRACTS The primary extraction methodology can result in significant cleanup of the sample by separating the ana1yte from the bulk of interfering matrix components. However, primary extraction methods are designed to decrease sample bulk rather than to achieve complete purification. Therefore, additional steps are generally needed. The degree of purification ranges from none to very little with binary solvent systems (e.g., aqueous acetonitrile or methanol) to significant with solvent systems that allow homogeneous partition such as in the DFG S19 method, which uses a sequential system consisting of NaCl, water, acetone, ethyl acetate, and cyclohexane and results in rather clean extracts for watery crops, or with SFE or ASE, as discussed previously. Therefore, purification regimes must generally be instituted subsequent to the initial analyte separation.
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Traditional approaches to extract cleanup usually involve liquid/liquid partition. The most common technique focuses on differences in solubility (and polarity) of the matrix constituents. Thus, it is common to find methods in which (a) fatty components in matrices are removed by partition with nonpolar solvents (acetonitrilelhexane fat cleanup, extraction of aqueous extracts, or residues with dichloromethane) or (b) acidic or basic analytes are converted to water-soluble salts and the aqueous phase extracted with organic solvents to remove matrix components. Gel permeation chromatography (GPC) is an established cleanup technique that separates (size exclusion) high-molecular-weight compounds such as proteins, fats, and sugars from relatively low molecular weight compounds such as pesticides in animal and plant matrices. Typically, GPC involves injection of large-volume solutions of analyte ('"'-'5 ml) via autosampler onto a low-pressure glass column containing polystyrene divinylbenzene beads conditioned with the appropriate solvent. Fractions of eluate are then collected and concentrated prior to further chromatographic analysis. A recent modification (Chambers, 1998) involves using a high-pressure steel column, lower injection volumes, and collection of smaller fractions, thus minimizing solvent use and concentration time. GPC is an integral part of multiresidue methods (Thier and Zeumer, 1987). A more efficient, but time-consuming method involves liquid/solid partition. For several decades, silica gel and florisil and size exclusion supports (gel permeation) have been used in liquid chromatography (LC). Thin-layer chromatography is far less effective. However, the advent of reverse-phase (RP) systems and high-efficiency SPE (and microextraction) cartridges has revolutionized the approaches to analyte purification. Whereas normal-phase (silica-based) supports could handle only organic solvents of medium to low polarity, reverse-phase systems can extract hydrophobic pesticide residues directly from aqueous solutions without involving significant amounts of organic solvents. The first report on applications of solid-phase extraction (SPE) with reverse-phase supports (Belardi and Pawliszyn, 1989) involved chemically modified fused-silica fibers. There are now an increasing variety of other solid supports relying on ion exchange, size exclusion, and other physicochemical properties (Zhang et aI., 1994). The technique has rapidly advanced, especially for the analysis of pesticide traces in systems such as surface and groundwater (Balinova, 1993; Eisert and Levsen, 1995a; Field et aI., 1997; Hatrik et aI., 1994; Moore et aI., 1995) and procymidone in wine (Vrruty et aI., 1997). SPE has also become an integral part of multiresidue methods (Analytical Methods for Pesticide Research, 1996; Barnabas et aI., 1995; Nouri et aI., 1995). A great advantage of SPE is that in many cases similar HPLC column supports are available (C8, C18, aminopropyl), which can be used to predict the chromatographic behavior of the analyte relative to potential interferences. Solid-phase extraction Empore disks are an alternative to SPE cartridges. These disks may eliminate certain cleanup procedures and further reduce the use of organic solvents. The
disks may be extracted with small amounts of solvents directly in the autosampler vial (Field and Monahan, 1995, 1996), thus allowing for increased method automation. In general, SPE methodology is increasingly being incorporated into on-line systems (Marce et aI., 1995; Maris et aI., 1985; Nielen et aI., 1987) with an emphasis on polar pesticides such as diuron and bromacil in water samples (Parrilla et aI., 1993; Sancho et aI., 1997; Sennert et aI., 1995).
30.3.4 SEPARATION AND QUANTITATION 30.3.4.1 Gas Chromatography and Mass Spectrometry For the past 40 years, gas chromatography (GC) has been the most widely utilized technique to analyze pesticide residues. Advances in chromatography, in particular, capillary column technology, have provided an increasing variety of thermally stable stationary phases, thus improving selectivity. The development of highly specific carbon, phosphorus, sulfur, and nitrogen detectors based on flame ionization and photometry (NPD, FPD) and on electron capture (ECD) culminated with the direct coupling of capillary GC columns with decreased carrier gas flow to mass spectrometers, in particular, those equipped with quadrupole analyzers (March, 1997). Thus, selectivity and sensitivity were improved tremendously. Large numbers of samples could thus be screened cheaply and efficiently once automation was introduced for sample injection. In typical applications, such a GC-MS system can provide quantitation of 20-40 samples overnight, including full- or selected-ion spectra on several components. In fact, GC-MS is already being used extensively for the V.S. Department of Agriculture Pesticide Data program, as exemplified by the routine simultaneous determination of diphenyl amine, o-phenylphenol, and propargite in apples (Yu et aI., 1997). Other significant developments in GC-MS included the following: 1. The use of high-resolution mass spectrometry (HRMS) with magnetic sector instruments, which led to limits of detection in the femtogram range, especially for analytes that contain heteroatoms with significant mass defects (e.g., chlorine) and that are prone to give simple spectra on negative chemical ionization (NCI). 2. The introduction of the ion trap mass spectrometer (ITD), which resulted in better sensitivities in the full-scan mode, thus providing improved identification power in combination with nontarget multiresidue enforcement methods. 3. The introduction of multiple MS capabilities either by means of a row of quadrupoles (triple-quads), providing MSIMS in space; or by the use of ion trap mass spectrometers, allowing MS n (multiple MS experiments) in time.
30.3 Development of the Analytical Method The MSIMS or MS n techniques result quite often in better selectivity and thus in improved sensitivity compared to the single-quadrupole or early ion trap techniques, whereas the price for the instruments is still much lower than that for the high-resolution MS instruments. Thus, the versatility of these instruments, especially of the ion trap mass spectrometers equipped either with an external ionization source (which allows negative and positive chemical ionization) or with internal ionization (which seems to give better ion yields on electron impact and positive chemical ionization), has resulted in a tremendous improvement in the limit of detection for many analytes that can be chromatographed by capillary Gc. Multiple MS thus serves as an additional "cleanup" technique, allowing for very low detection limits of analytes in matrix. For example, nanogram per liter detection of several pesticides in aquatic matrices has been reported using tandem GC-MS (Boyd-Boland et aI., 1996; Rossi et aI., 1997; Steen et aI., 1997). That GC-MSIMS is rapidly becoming the preferred technique for pesticide analysis (Feigel, 1997) is due in part to the added capability for confirmation of identity, which allows for elimination of the false positives often detected in GC-MS analysis (de Cruz et aI., 1996; Schachterle and Feigel, 1996). Furthermore, the possibility of conducting two or more MS n experiments in the search for ions not present in co-eluting interferences greatly simplifies the extraction and purification process. GC-MS has proven useful in multiresidue methods, sometimes dealing with mixtures containing in excess of 100 compounds, in a very cost-effective manner (Fillion et aI., 1995; Liao et aI., 1991). In accordance with present trends toward automation of pesticide analysis, it may be expected that in-line techniques coupling solid-phase microextraction of pesticide mixtures with GC-MS (Eisert and Levsen, 1995b) will be increasingly reported. 30.3.4.2 Liquid Chromatography and Mass Spectrometry
A major shortcoming of GC-MS techniques is their inability to analyze samples of low volatility, high polarity, or thermal instability. Because regulatory requirements focused only on the parent pesticide, GC-MS could accomplish the required goals for the great majority of commercial products. However, in the past decade the need for quantitative data on degradation products and metabolites has increased and now the complete expression for "toxic residues" routinely includes one or more metabolites in addition to the pesticide itself. A multitude of derivatization techniques (Lunn and Hallwig, 1998) is available to convert hydroxy, carboxy, amino, and other polar derivatives to entities amenable to GC-MS. The derivatization approach is expensive and often unreliable in view of the variety of matrices and co-extractives involved. Furthermore, the analytes of interest are often obtained in aqueous fractions incompatible with the majority of available derivatizing agents. Liquid chromatography (LC) can successfully handle polar, nonvolatile compounds but, as a residue technique, it has
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developed more slowly than GC, and, mainly due to the lack of element-specific detectors, LC has not reached the wide applicability of gas chromatographic methods. Although detectors utilizing ultraviolet absorbance, fluorescence (with/without postcolumn derivatization), electrochemical properties, and refractive index have proven useful, the advent of combined liquid chromatography and mass spectrometry (LC-MS) has provided renewed and strong interest in LC as a residue analytical tool. The union of the two analytical techniques in LC-MS combines an instrument that operates in the condensed phase with one that operates at reduced pressures. Thus, for LC, it is the mobile phase and flow rate that affects separation of analytes; for MS, it is the ionization mode and factors affecting the production and transport of gas-phase ions into the analyzer that are important. In the 1980s, several approaches evolved toward finding a practical application for LC-MS: First were the moving belt interface and direct liquid introduction, which were later replaced by particle beam and thermospray technologies (Cairns and Siegmund, 1990). These methods have been applied to a great variety of analytical problems involving pesticides, their metabolites, and conjugates (Brown, 1990). In particular, it is worth noting that LC-MS proved itself as a viable technique with the successful analysis of thermally labile sulfonylurea herbicides (Shalaby and George, 1990). These compounds are used at very low application rates so the use of thermospray techniques established the increased sensitivity of LC-MS as compared to other methods. In spite of interface improvements such as thermospray, complications in the use of LC-MS were common and based on the introduction of a fluid stream into a vacuum system. At present, the best results are being obtained with ion sources developed over the past decade that operate at or near atmospheric pressure. Atmospheric pressure chemical ionization (APCI) relies on nebulization of the solvent stream followed by thermal evaporation. Thus, the mixture is ionized in the vapor phase and the reactant ions are formed from the components present in the LC eluent. Chemical ionization utilizes solvent molecules as the "reagent gas." APCI can handle high flow rates and high electrolyte concentrations. The second widely utilized atmospheric pressure technique is electrospray ionization (ESI). Here the sample has to be in an ionized form in solution. Neutral samples can be converted to ions by adjustment of pH or by addition of electrolytes (such as ammonium acetate) to form ion-molecule complexes. The influence of the species utilized for ion-pair processes in LC is quite important as evidenced in the analysis of diquat and paraquat (Startin et aI., 1998). The sample solution is then dispersed into an electrically charged aerosol. The target ions are separated from the droplet interface by the action of an electric field at the surface of the charged droplet, which, upon partial solvent removal, develops a substantially smaller cross section. The gas-phase ions are then transported to the mass analyzer. Because electrospray ionization can be accomplished at low temperatures, this tech-
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nique is ideal for thermally labile compounds. ESI generally works best at low flow rates and with lower concentrations of electrolytes. Both APCI and ESI are quite sensitive to eluent composition and to matrix effects. Thus, extensive method development is required in the early stages of analysis to optimize sensitivity, selectivity, and compatibility with matrix components and eluent solvents. In spite of these drawbacks, APCI and ESI are rapidly becoming the LC-MS techniques of choice. As with GC-MS, tandem (MSIMS), or MSn, methods for LC have proliferated (Gilbert et aI., 1995). The great advantage of the MSIMS technique using triple-quadrupole instruments (MSIMS-in-space) and of the MS n (-in-time) technique using ion trap technology lies in the gain of selectivity using the "multiple-reaction mode" (MRM). In the case of APCI, protonated quasi-molecular parent ions are selected either by passing the first quadrupole or by eliminating all other ions in the ion trap, and then fragmented further by collision-induced dissociation in the middle quadrupole or in the trap by applying energy. The daughter ion fragments formed are then filtered by the third quadrupole or by the ion trap and yield highly selective MS signals. The majority of pesticides and metabolites can be detected by LC-MSIMS in concentrations greater than 10 ng/ml with high selectivity, allowing much abbreviated cleanup procedures. However, although the current view is that MSIMS techniques can largely eliminate cleanup steps, the issues and strategies that have historically been important in sample preparation and analysis remain fundamentally important to reliable LC-MSIMS. The "dilute and shoot" approach is valid in some cases where analyte concentrations are high and matrix components do not co-elute or otherwise interfere with ionization. Contamination of HPLC column and MS source by matrix components, however, may cause variation and drift of the LC-MS signals and thus hamper unattended automated analysis of crude extracts. This can be avoided to a certain extent by introducing precolumn-switching techniques or automated on-line SPE cleanup steps. An obvious approach to bypass matrix effects is to prepare and analyze the quantitation standards in solutions containing matrix. However, the EC and U.S. guidelines diverge on this point. Whereas in Europe the use of standard in-matrix calibrants is encouraged, the EPA does not generally allow it. As with all mass spectrometric techniques, the use of isotopically labeled internal standards results in an improved reliability and ruggedness of methods. This approach is very well established in pharmaceutical analysis, but nonradioactivelabeled tracer compounds are available for only a limited number of pesticides and metabolites. Use of internal 13C_ or 2H-labeled standards is of particular usefulness when extraction techniques result in partial losses of analytes, as in the case with chlorinated anilines (Hurlbut et aI., 1998), which react with matrix components, and of pentachlorophenol (Gremaud and Turesky, 1997). Analytical methods using GC or LC-MSIMS and MS n techniques are per se target methods (i.e., tailored to detect with
high selectivity and sensItIvIty one or few analytes). Thus, the use of MSIMS has limitations for enforcement methods that should ideally cover a multitude of pesticides and relevant metabolites. This is especially true for LC-MS due to the limited separation power of HPLC in comparison to the much higher peak capacity obtained by capillary Gc. This disadvantage can be overcome by two approaches, as follows. First, if several rugged automated short LC-Msn methods analyze sample extracts after a general cleanup procedure consecutively for several groups or classes of compounds, this would allow unattended screening of one extract for many analytes. As sample extraction and cleanup are time-consuming steps, this approach allows increased sample throughput for enforcement purposes. Second, a technique called data-dependent full-scan MSIMS can be implemented with ion trap mass spectrometers. This application first screens in the full-scan mode the mass spectral information for compounds eluting from the HPLC column. As soon as ions formed from a potential analyte are detected in a defined retention time window, the ion trap switches to a predetermined MSIMS method for improved selectivity, thus providing additional information and confirmation. A combination of these two approaches could very well result in the future in reliable and affordable multiresidue methods for pesticides and relevant metabolites not covered by GC-based multimethods. 30.3.5 IMMUNOASSAY TECHNIQUES
The use of immunoassays (lAs) as pesticide residue analytical methods represents a radical departure from the more conventional chromatographic approaches. Immunoassays utilize antibodies that have been prepared in animals (commonly rabbits, mice, or sheep) to a particular pesticide or family of pesticides. These molecules are too small to elicit immune responses by themselves, but, upon coupling of a chemical analog of the pesticide to a carrier (usually a protein), the "conjugate" may evoke production of antibodies. For the method to be successful, these antibodies must be able to bind selectively to the free pesticide. The key steps in development of an antibody test are as follows (Gee et aI., 1995; Harris et aI., 1998): 1. Synthesis of a pesticide (or derivative), coupled to a suitable carrier protein for immunization. 2. Immunization of rabbits, mice, and/or other species; preparation and purification of antibodies. 3. Development of initial immunoassay using pesticide standards; checking assay sensitivity and specificity. 4. Assessment of assay performance with water and soil matrices in laboratory-spiked and field samples. 5. Formatting of methods as prototype kits, stabilization and stability trials on components and prototypes. 6. Field trials of kits and training workshops. The advantages associated with lA include low detection limits and high analyte selectivity. Because sample preparation
30.3 Development of the Analytical Method
is minimal, the methods allow for high throughput, thus increasing cost effectiveness. These advantages have been extensively reviewed in the literature (Harris et aI., 1995, 1998). A key limitation in the development of lA methodology is the longer time required when compared to traditional instrumental methods. Specifically, the selection of the target analyte analog (hapten) is critical for the production of the high-affinity antibodies required for high selectivity and sensitivity. The functional group used for protein coupling should not mask the key structural feature(s) of the target analytes so that the immune system of the host animal can recognize it. Hapten design and synthesis have been extensively reviewed (Goodrow et aI., 1990, 1995). The coupling (conjugation) of the hapten with a carrier protein must result in a product that is of adequate solubility and stability under the reaction conditions and that contains the appropriate functional groups (Brinckley, 1992; Erlanger, 1980). The unreacted products are then separated from the conjugate by dialysis, gel filtration, or other methods. Animal immunization to obtain polyclonal or monoclonal antibodies is conducted under conditions that enhance the immune response (Harlow and Lane, 1988; Tijssen, 1985). The affinity of the antibodies for the immunizing hapten is then evaluated to determine antibody titers against it. Inhibition experiments are then conducted to determine the potential for each target analyte toward inhibition of binding between the hapten and the antibody. Validation of the lA method is conducted after matrix effects and the influence of pH, salts, solvents, and other components are identified. As with other animal methods, validation involves determination of fortified recoveries in the appropriate matrices. Schneider et at. (1995) detail troubleshooting procedures that may be considered during method development and validation. Numerous examples of successful lA methods are reported for organophosphates, pyrethroids, triazines, urea herbicides, and other compound classes (Gee et aI., 1995; Harris et aI., 1998). For the analysis of specific agrochemicals, commercially available immunoassay kits are generally more cost effective than traditional instrumental analysis. This factor makes the continuing development of lA methods attractive, in particular, for use in monitoring studies and for pesticide analysis in developing countries. However, substantial purification of analyte from matrix is required before lA methods can be applied. 30.3.6 CAPILLARY ELECTROPHORESIS Electrophoresis refers to the migration of electrically charged species when dissolved in an electrolyte through which an electric current is passed. Capillary electrophoresis (CE) combines a variety of modem analytical techniques with a wide range of applications, including the analysis of biopolymers [such as deoxyribonucleic acid (DNA), proteins, peptides], natural products, pharmaceuticals and drugs, and fine chemicals, including agrochemicals. CE makes use of various separation modes with distinct ranges of application, such as the following:
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1. Free-solution capillary electrophoresis (FSCE), which is mainly used for the separation of ions based on differences in their charge-to-mass ratios, whereby the pH controls the dissociation and protonation of functional groups on the analytes. 2. Micellar electrokinetic capillary chromatography (MECC or MEKC), which uses relatively high levels of ionic surfactants forming micelles. The separation of neutral analytes is based on the hydrophobic interaction with the micelles that migrate in the capillary. The use of chiral cyclodextrins provides a relatively cost effective and powerful method for enantioselective separations. 3. Capillary electrochromatography (CEC) is a fusion of liquid chromatography and capillary electrophoresis, where the capillary is packed with stationary phase similar to those used in liquid chromatography, and the flow of the mobile phase is caused by the electroosmotic flow (EOF) between the electrodes. The selectivity of the separation depends on partition between the stationary and mobile phases. Capillary electrophoresis has to be considered for the analysis of polar and charged analytes and thus follows the trend in pesticide chemistry to use more hydrophilic water-soluble active substances and for the requirement to include polar metabolites into the residue definition. For routine residue analysis, however, CE is generally only considered when conventional chromatographic approaches such as capillary GC, HPLC, and LC-MS fail to provide straightforward solutions, for the following reasons. First, expertise and instrumentation in GC, HPLC, and LC-MS are more readily available in residue research, contract, and enforcement laboratories, and the pressure to use CE for residue analysis only exists when the other techniques do not provide rational and cost-effective methods. Second, there are several aspects of CE that do not facilitate its use in residue analysis:
1. Whereas CE provides excellent sensitivity for many applications, the limitation in sample size, which is in the nanoliter range, results in insufficient overall sensitivity. Injection size can be increased by the use of a process called "stacking" where the analytes are concentrated in a sample zone prior to the chromatographic separation, thus reducing the starting peak width. Miniaturization of extraction and cleanup techniques may provide a means to obtain decreased final extract volumes. However, in residue analysis, there is a limit to decreasing the original sample size. 2. The most frequently used detection methods in CE are ultraviolet (UV) absorbance or UV diode array. Very short path lengths, however, again limit the sensitivity obtained by CE. On the other hand, the use of a mass spectrometric detector (quadrupole, triple-quadrupole, ion trap, or time-of-flight mass spectrometer, MSIMS techniques), predominantly with ESI sources, provides good sensitivity and high selectivity. The
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presence of electrolytes and surfactants in the caSe of MECC, however, limits applications and overall performance. In summary, with the current trend away from highly lipophilic pesticides of great environmental persistence toward more polar active substances and the necessity to detect polar metabolites, there is an increasing number of potential applications in pesticide residue analysis for CE.
30.4 SUMMARY Regulatory requirements that address pesticide concentrations in foods and environmental matrices have significantly accelerated the development of faster, less costly, and more sensitive and specific analytical techniques. Extraction techniques utilizing modem approaches such as supercritical fluid and accelerated solvent extractions allow for use of smaller sample sizes and solvent volumes. Cleanup of extracts can now be accomplished with commercial chromatography products and the process can be automated to address large sample numbers. The most dramatic improvements have taken place in the analytical instrumentation and automation options available. In particular, the development of multiple ionization techniques and secondary ion production in mass spectrometry have advanced detection limits and improved selectivity. Developments in immunoassay-based analysis promise to provide low detection limits and high analyte selectivity coupled with relatively low cost.
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quantitation using an atomic emission detector. J. Chromatogr. Sci. 31,445449. Snyder, J. L., Grob, R. L., McNaIIy, M. E., and Oostdyk, T. S. (1992). Comparison of supercritical fluid extraction with classical sonication and Soxhlet extractions of selected pesticides. Anal. Chem. 64, 1940-1946. Snyder, J. L., Grob, R. L., McNaIIy, M. E., and Oostdyk, T. S. (1993). The effect of instrumental parameters and soil matrix on the recovery of organochlorine and organophosphate pesticides from soil using supercritical fluid extraction. J. Chromatogr. Sci. 31, 183-191. Specht w., Pelz, S., and Gilsbach, W. (1995). Gas-chromatographic determination of pesticide residues after cleanup by gel-permeation chromatography and mini-silica gel chromatography. Fresenius' J. Anal. Chem. 353, 183190. Stanek, M., and KeIIer, G. (1998). Determination of pesticide residues using accelerated solvent extraction. In "Ninth International Congress of Pesticide Chemistry (IUPAC)," Vol. 2, 7A-034. Startin, J. R., Hird, S. J., Jones, A., and HilI, A. R. C. (1998). Analysis of residues of paraquat and diquat in plant and animal tissues by LC-MS. In "Ninh International Congress of Pesticide Chemistry Book of Abstracts," Vol. 2, 7A-007.
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CHAPTER
31 Risk Assessment and Risk Management: The Regulatory Process* Penelope A. Fenner-Crisp U.S. Environmental Protection Agency
31.1 INTRODUCTION In the United States, primary authority for pesticide regulation resides with the US. Environmental Protection Agency (EPA) under the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) and the Federal Food, Drug, and Cosmetic Act (FFDCA). Under FIFRA, EPA registers pesticides for use. This law also authorizes the Agency to prescribe the conditions of use of pesticide products. Under FFDCA, EPA establishes maximum allowable levels of pesticide residues ("tolerances") in foods and animal feeds. These tolerances are enforced by the Food and Drug Administration (FDA) of the Department of Health and Human Services (HHS) for most foods and by the U.S. Department of Agriculture (USDA) for meat, poultry, and some egg products.
31.2 HISTORICAL BACKGROUND OF PESTICIDE REGULATION IN THE UNITED STATES Regulation of pesticides at the federal level has been in place for nearly a century. Each time the law has been amended, the number and nature of the directives have been expanded and embellished. Some, but not all, of these changes will be described. Only those areas of the law(s) that remain in effect under current legislation are discussed. The first legislation passed was the Federal Insecticide Act of 1910 (FIA, 1910). The provisions of this act, essentially only a labeling statute, were limited to the prohibition of the manufacture of any insecticide or fungicide that was "adulterated or misbranded." No requirement for registration and no establishment of standards of safety were included at that time. Congress *This document reflects the opinions only of the author and does not necessarily represent official policy of the U.S. Environmental Protection Agency. Handbook of Pesticide Toxicology Volume 1. Principles
passed the first version of the Federal Insecticide, Fungicide, and Rodenticide Act in 1947, adding the requirement of registration by the Secretary of Agriculture before sale or distribution in interstate or foreign commerce, but without providing the US. Department of Agriculture the power to deny or cancel a registration if the registration did not comply with the provisions of the law (FIFRA, 1947) A 1964 amendment to FIFRA did provide USDA with the authority to deny or rescind a registration and to issue an immediate suspension of registration if necessary to prevent an imminent hazard to human health (FIFRA,1964). Nearly half a century after Congress passed the first federal pesticide regulatory legislation, it amended the Federal Food, Drug, and Cosmetic Act to require the Food and Drug Administration to establish maximum acceptable levels ("tolerances") for pesticide residues in foods and animal feeds (FFDCA, 1954). This requirement applied only to raw agricultural commodities. Four years later, Congress once again amended FFDCA to include a requirement for a tolerance in a processed food, but only if the pesticide residue in that processed food was expected to exceed the tolerance level in the related raw agricultural commodity (FFDCA, 1958). In 1970, under President Nixon's government Reorganization Plan No. 3, the primary federal authority for the regulation of pesticides was transferred from USDA and FDA to the newly created EPA (Nixon, 1970). Between 1970 and 1990, Congress amended FIFRA six times, each time adding to, or enhancing, the Agency's existing responsibilities. One of the provisions added in 1972 was that a pesticide could be registered only if it did not cause "unreasonable adverse effects" on human health or the environment [Federal Environmental Pesticide Control Act of 1972 (FEPCA, 1972)]. The 1972 revisions also established the requirement for reregistration of all existing pesticides within a five-year time frame. When this was not accomplished by 1978, Congress relaxed the timelines [Federal Pesticide Act of 1978 (FPA, 1978)]. Ten years later,
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following continued delays in completing the reregistration process, Congress once again established specific timetables in a five-phase program for all active ingredients registered before November 1, 1984, but this time with targeted funding to support the reregistration activity (FIFRA, 1988). Completion of the reregistration process is estimated for 2006. Other important provisions in the FIFRA amendments of 1972 included authorization (1) to require federal registration of pesticides sold within states, (2) to classify pesticides in general and/or restricted use categories (based upon their inherent acute toxicity) and otherwise regulate their usage (e.g., specify the maximum allowable application rate and frequency), and (3) to register establishments that make pesticide products and require them to maintain records, while also being able to inspect these producers, as well as those establishments which hold pesticides for sale, for compliance with the applicable provisions of FIFRA. Amendments to FIFRA in 1975 included the requirement for EPA to notify the Secretary of Agriculture in advance of issuing proposals for regulations or to cancel or otherwise change the registration status of a pesticide, and to consider the impact on agriculture when cancellation actions are being considered (FIFRA, 1975). In addition, the seven-member FIFRA Scientific Advisory Panel (SAP) was established to "comment as to the impact on health and the environment" of proposed cancellation actions and regulations. In the 25 years since this provision was added to FIFRA, the SAP has been consulted on many scientific issues reflecting a much broader range of EPA's pesticide regulatory activities, for example, the proposed classification of human cancer potential of a pesticide, the design of a testing protocol, and the risk assessment methodologies developed to address aggregate and cumulative exposure and risk assessment. Congress once again substantially revised FIFRA in 1978 (FIFRA, 1978). Provisions included granting data submitters 10 years of exclusive use of their data on new active ingredients while transferring the responsibility for managing the issue of data compensation from the Agency to outside arbitrators. The 1978 revisions also removed most trade secret protection for health and safety data, an early example of public rightto-know. Other changes included the granting of conditional registration authority which allows EPA to approve proposed uses before the full set of supporting data are submitted and reviewed; the establishment of a procedure for interim administrative review (a process known as Special Review) if "a validated test or other significant evidence raised prudent concerns of unreasonable adverse risk to man or to the environment;" a prohibition against disclosure of data to foreign or multinational pesticide producers; and a requirement for the recognition of the distinction between agricultural and nonagricultural pesticides when processing registration petitions and in setting registration standards and guidelines. Among the modifications to FIFRA in 1980 was the provision that the Scientific Advisory Panel could create its own subpanels, resolving the limitation in scope of expertise that the smaller, seven-member permanent panel may have on any
specific issue (FIFRA, 1980). This allowed expansion, of the panel's capabilities to provide more substantive expert scientific peer review of the increasingly diverse and complex issues brought to it by the Agency. The 1980 provisions also required EPA to request SAP comment "as to the impact on health and the environment" on proposed suspension actions (in addition to the requirements for consultation on proposed cancellations) and to set up a process for peer review "with respect to the design, protocols, and conduct of major scientific studies conducted under this Act by the Environmental Protection Agency or by any other Federal agency, any State or political subdivision thereof, or any institution or individual under grant, contract, or cooperative agreement from or with the Environmental Protection Agency" and of "the results of any such scientific studies relied upon by the Administrator with respect to any actions the Administrator may take relating to the change in classification, suspension, or cancellation of a pesticide." In 1988, in addition to prescribing the multi phase reregistration program, Congress made other changes to FIFRA, including creation of a "fast-track" registration process for enduse products for "me-too" registrants ("me-too" registrants are those who seek to register products similar to an already registered pesticide product); the necessity of gaining Congressional approval to indemnify registrants holding suspended or cancelled products; and a series of provisions related to the storage, disposal, and transport of suspended or cancelled pesticides and pesticide containers (FIFRA, 1988). Record keeping requirements were expanded to include all registrants and applicants for registration, in addition to those previously required of producers. The Food, Agriculture, Conservation, and Trade Act of 1990 (FACTA, 1990) added requirements for certified pesticide applicators to maintain records of their use of restricted-use chemicals, prohibited registrants of minor-use pesticides from submitting field trial data from geographic areas where the chemical would not be used, and provided discretion to the Administrator to reduce or waive registration fees if the cost would "significantly reduce the availability of the pesticide." In addition, there were several new requirements related to voluntary cancellation of minor-use pesticides, including a provision for public notice and comment upon a registrant's application for voluntary cancellation.
31.3 CURRENT STATE OF PESTICIDE REGULATION IN THE UNITED STATES-THE FOOD QUALITY PROTECTION ACT OF AUGUST 3, 1996 Amendments to FIFRA and FFDCA in 1996 brought both incremental and broad, sweeping changes to the legal foundation for pesticide regulation in the United States [The Food Quality Protection Act of 1996 (FQPA, 1996)]. The Food Quality Protection Act represents the outcome of long and complex deliberations to resolve inconsistencies in the previous legislation, both within and between the two statutes.
31.3 Current State of Pesticide Regulation in the United States
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FIFRA §4 specifies that tolerances and exemptions from tolerances must be reassessed as part of reregistration to determine whether they meet the requirements of the FFDCA. These reassessments must be made as soon as EPA has sufficient information to assess dietary risk, but no later than when it makes product reregistration decisions. These determinations for existing tolerances and exemptions and the need for any additional tolerances or exemptions must be published in the Federal Register and appropriate regulatory action under FIFRA and/or FFDCA begun promptly.
economic incentive to support the costs of registration or reregistration. In addition, the minor use pesticide must play a significant role in managing pest resistance or in an integrated pest management (IPM) program. There also must be a lack of efficacious alternatives or the alternatives must pose a greater risk to human health or the environment than does the pesticide under evaluation. Many minor-use crops are fruits and vegetables, which are significant components of the human diet. The law provides additional incentives for the development and maintenance of minor use registrations in a number of ways: extension of time to generate residue data and for exclusive use of these data; greater flexibility to waive data requirements; the option to waive some or all of the fees usually charged to support and maintain registration; and expedited review of minor-use applications by the Agency. None of these provisions would apply, however, if the minor use is determined to pose unreasonable risks or if the lack of data would significantly delay EPA decisions. FQPA establishes a USDA revolving grant program and a program for support of public health pesticides to be implemented jointly by the Public Health Service of HHS and EPA. By virtue of instituting this program, the federal government bears the cost of developing the required data to support the registration and reregistration of the public health use, as it does for minor-use pesticides used on agricultural crops under the USDA Inter-Regional Project Number 4 (IR-4) program. A public health pesticide is defined in FIFRA §2(nn) as "any minor use pesticide product registered and used predominantly in public health programs for vector control or for other recognized health protection uses, including the prevention of viruses, bacteria, or other microorganisms (other than viruses, bacteria, or other microorganisms on or in living man or other animal) that pose a threat to public health." A "vector" is defined in FIFRA §2(00) as "any organism capable of transmitting human discomfort or injury, including mosquitoes, flies, fleas, cockroaches, or other insects and ticks, mites, or rats." Perhaps the most notable example of a public health pesticide is DDT, s~ill used in some parts of the world, but not the United States, for the control of mosquitoes bearing the malaria vector.
31.3.1.4 Registration Renewal
31.3.1.6 Antimicrobial Pesticides
The requirement for periodic registration review was introduced. Formal procedures are to be established, with the aim of updating a pesticide's registration eligibility at least once every 15 years. The goal of this requirement is to ensure that all pesticides continue to meet up-to-date standards for safety testing and the protection of human health and the environment.
FIFRA §2(mm) defines an antimicrobial pesticide as one which is intended to "disinfect, sanitize, reduce, or mitigate growth or development of microbiological organisms" or "protect inanimate objects, industrial processes or systems, surfaces, water, or other chemical substances from contamination, fouling, or deterioration caused by bacteria, viruses, fungi, protozoa, algae, or slime" and, in this use, is exempt from a tolerance under FFDCA §408. Wood preservatives, antifouling paints, agricultural fungicides, aquatic herbicides, and liquid chemical sterilants intended for use on critical or semicritical medical devices as defined under FFDCA are not included within the definition. FQPA contains special provisions for antimicrobial pesticides, essentially removing them from the shadow of pesticides intended for agricultural and other nonfood uses and prompting
31.3.1 FIFRA-KEY CHANGES AND ADDITIONS 31.3.1.1 Emergency Snspension EPA may now suspend a pesticide registration in an emergency situation without simultaneously issuing a notice of intent to cancel, a change from the previous requirement for simultaneous action. A notice of intent to cancel must be issued within 90 days or the suspension will automatically expire. Determination of imminent hazard (to human health or the environment) constitutes grounds for suspension. This process is invoked to prevent unacceptable risks from occurring during the time required to cancel or otherwise modify the registration of a pesticide. Any action to suspend, cancel or modify the registration of a pesticide under FIFRA must be accompanied by a similar and simultaneous action on any associated tolerances under FFDCA. 31.3.1.2 FIFRA Scientific Advisory Panel A Science Review Board to consist of 60 scientists was established to be available to the permanent panel to assist in the scientific peer reviews conducted by the panel. Formation of the board complements the earlier modification to FIFRA which allowed the panel to create its own subpanels as needed. 31.3.1.3 Tolerance Reevaluation as Part of Reregistration
31.3.1.5 Protections for Minor-Use Pesticides, Including Public Health Pesticides A minor use is defined as one in which the pesticide is used on an animal, on a commercial agricultural crop for which the total U.S. acreage is less than 300,000 acres, or for the protection of public health, but does not, on its own, provide sufficient
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more focused attention on facilitating more timely registration decisions. The law requires the identification, evaluation, and implementation of reforms to the registration process for this class of pesticides to reduce review periods, providing explicit goals in number of days depending upon the action requested (i.e., a new use of an already-registered active ingredient; a new product; "me-too's"; and amendments to existing uses). A separate administrative unit has been established within the Office of Pesticide Programs that deals only with the registration, reregistration, and Special Review processes for antimicrobial pesticides. A separate section in 40 CFR 158 describes the data requirements necessary to support registration or registration for these products. 31.3.1.7 Reduced Risk or "Safer" Pesticides In the early 1990s, the Office of Pesticide Programs set up a system by which reduced risk or "safer" pesticides would be given priority attention in the registration process. The most current guidelines governing expedited review of conventional and biological pesticides were issued in 1997 [U.S. Environmental Protection Agency (U.S. EPA, 1997b)]. FQPA provided the statutory mandate for continuing this expedited consideration of applications for pesticides which meet one or more of the criteria for a reduced risk pesticide. A pesticide qualifies for expedited review as a reduced risk pesticide if its use "may reasonably be expected to accomplish 1 or more of the following:" (1) reduce the risks to human health; (2) reduce the risks to nontarget organisms; (3) reduce the potential for contamination of ground water, surface water, or other valued environmental resources; and (4) broaden the adoption of IPM strategies, or make them more available or effective [FIFRA §3(c)(10)(B)]. 31.3.1.8 Data Collection The keystone of FQPA is the inclusion of special provisions for infants and children. Title III ofFQPA addresses data collection activities to assure the health of infants and children. It states that USDA, in cooperation with FDA and/or EPA, "shall coordinate the development and implementation of survey procedures to ensure that adequate data on food consumption patterns of infants and children are collected"; shall ensure there will be improved data collection on occurrence of pesticide residues in foods, particularly those most likely consumed by infants and children; and shall evaluate the current status of pesticide usage information and move to improve usage information gathering activities. Information in all three of these areas is critical to the conduct of credible and accurate estimates of risk from exposure to pesticide residues in the diet. 31.3.2 FFDCA-KEY CHANGES AND ADDITIONS The most significant changes in pesticide regulation resulting from the passage of FQPA impact the tolerance-setting process described in FFDCA. Definitional and process changes
were mandated and the factors that are to be considered when conducting risk assessments and making risk management decisions were expanded. 31.3.2.1 The Delaney Clause Until FQPA was passed, pesticide residues in processed foods were considered to be "food additives" regulated under FFDCA §409. If a pesticide residue was expected to exceed the level which was allowed under a §408 tolerance for the raw agricultural commodity, it became necessary to establish a separate food additive regulation for the processed food under §409. However, the Delaney clause in §409 prohibits the establishment of food additive regulations for any substance "if it is found to induce cancer when ingested by man or animal, or if it is found, after tests which are appropriate for the evaluation of the safety of food additives, to induce cancer in man or animal. ... " Under the new law, pesticide residues are excluded from the definition of "food additive." Thus, the Delaney clause is no longer applicable to pesticide residues in processed foods. All pesticide residues, whether in raw or processed foods, are regulated only under FFDCA §408, which does not contain the prohibition against setting tolerances for carcinogens. 31.3.2.2 Definition of "Safe" Under FFDCA §408(b )(2)(A), the standard for establishing a tolerance is based on whether the tolerance is "safe." To be "safe" means that there is "a reasonable certainty that no harm will result from aggregate exposure to the pesticide chemical residue, including all anticipated dietary exposures and all other exposures for which there is reliable information." This definition is consistent with the standard applied historically to nonpesticide food additives and color additives by the Food and Drug Administration. For threshold effects (i.e., those effects for which a level can be identified that would not be expected to cause or contribute to any adverse human health consequences), the safety standard is satisfied if the aggregate exposure is lower than the no-effect level by "an ample margin of safety." Traditional regulatory policy states that, as a default, exposure estimated to be at or below a level WO-fold lower than the critical no-effect level identified in the animal toxicology database would meet the safety standard. For nonthreshold effects (i.e., those for which no no-effect level can be identified), a pesticide will satisfy the safety standard if the increased lifetime risk, expressed as a probability, is "negligible." Traditionally, "negligible" has been defined as being no greater than a one-in-a-million excess lifetime risk for nonoccupational exposures. 31.3.2.3 Aggregate Exposure Aggregate exposure is defined as that which occurs from all food uses for the pesticide, as well as from exposure that occurs from all nonoccupational sources. This would include exposures from drinking water, nonfood pesticidal uses (e.g., lawn and garden use or indoor residential, school, or public building
31.3 Current State of Pesticide Regulation in the United States
applications) and those exposures that may result from nonpesticidal uses (e.g., as a human pharmaceutical or a hazardous waste site contaminant). Principles for conducting an aggregate exposure assessment are being developed by EPA (U.S. EPA, 1999d). Among the factors that must be taken into account when establishing, modifying, leaving in effect, or revoking a tolerance or an exemption from a tolerance is the risk that may ensue from the aggregate exposure to the pesticide under evaluation. A tolerance represents a single pesticide-use combination. That is, one tolerance would be needed if Chemical X were to be used on potatoes. A separate tolerance would be required if Chemical X also were to be used on lettuce. Therefore, when making a decision with regard to anyone use, EPA must consider the exposure and risk that would occur not only as a consequence of that particular use, but also all other existing food and nonfood uses. In other words, can this new use be added to the existing "risk cup" for Chemical X? 31.3.2.4 Common Mechanism of Toxicity and Cumulative Risk Assessment Another factor that must be taken into account when establishing, modifying, leaving in effect, or revoking a tolerance or an exemption from a tolerance is the cumulative effects of the pesticide under evaluation and other substances with which it may share a common-mechanism of toxicity. "Other substances" are not just other pesticides, but may be chemicals such as drugs, commodity chemicals, or environmental contaminants. When a common mechanism finding is made, then a cumulative risk assessment is to be conducted. The first step is to determine the need for a cumulative risk assessment. This is done by conducting a hazard assessment for that pesticide. In the course of performing this assessment, information would come to light to suggest that this pesticide may share a common mechanism with at least one other pesticide. This process continues until all likely pesticide candidates are identified. After it is concluded that two or more pesticides are candidates for a common mechanism group, an effort is made to determine if reliable information exists to suggest that any non pesticides also may share the same mechanism of toxicity. EPA has developed criteria by which to judge whether substances share a common mechanism of toxicity (V.S. EPA, 1999a). At the conclusion of the process to determine those substances which actually do share a common mechanism of toxicity, those remaining substances are subjected to cumulative risk assessment. It is possible that the final cumulative risk assessment may not include all of the substances that constituted the original common mechanism group. Modifications to group membership would be informed by the results of the individual aggregate exposure assessments that also would be conducted on each original candidate for the common mechanism group. The nature, magnitude, and timing of exposure to each substance in the aggregate and how the exposures (or their biological consequences) to individual substances overlap become critical factors in determining the final group to be included in the cumulative risk assessment. Guidance is being developed by EPA for the conduct of cumulative
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risk assessments. Early articulations of principles are presented in several Agency documents (V.S. EPA, 1999a, b, 2000a, b). The first group of substances to be subjected to a cumulative risk assessment are the cholinesterase-inhibiting organophosphorus insecticides. 31.3.2.5 Special Considerations for Infants and Children When making tolerance decisions, EPA also must implement several new requirements related to assuring the safety of infants and children. The Agency must assess (aggregate) risk based upon available information about: (1) dietary consumption patterns that are likely to yield disproportionately higher exposures or risks; (2) special susceptibilities to pesticides, including neurological differences between infants and children and adults, and the effects of in utero exposure to pesticide chemicals; and (3) the cumulative effects of the pesticide residues and other substances that have a common mechanism of toxicity. The "reasonable certainty of no harm" safety standard must be ensured and a specific safety determination for infants and children must be made. The provision that has prompted the most controversy and has had the greatest impact upon the risk assessment and regulatory decision-making process under FQPA is the obligatory application of an additional safety factor. FFDCA §408(b )(2)(C) states that "in the case of threshold effects ... an additional tenfold margin of safety for the pesticide chemical residue and other sources of exposure shall be applied for infants and children to take into account potential pre- and postnatal toxicity and completeness of data with respect to exposure and toxicity to infants and children." A different margin of safety may be used only if, on the basis of reliable data, such a margin will be safe. It should noted that any different margin of safety could be greater or lesser than the default 10 x . Since the passage of FQPA, EPA has developed a series of policy guidance documents, representing the evolution of its approach to implementing the "FQPA Safety Factor" provision of the law. The most current thinking on this topic can be found in the draft document entitled The Office of Pesticide Programs' Policy on Determination of the Appropriate FQPA Safety Factor(s)for Use in the Tolerance-Setting Process (V.S. EPA, 1999b). This document describes when FQPA safety factor decisions are needed; what the FQPA lOx safety factor is "in addition to"; how to judge the completeness of the toxicology and exposure databases; when a database uncertainty factor greater than I x is applied; how to determine, and account for, the degree of concern for pre- and postnatal toxicity; and the process for determination of the appropriate FQPA safety factor(s). Earlier articulations of principles are presented in several Agency documents (V.S. EPA, 1996, 1998a, b). 31.3.2.6 Consumer Right-to-Know FFDCA §408(0) states that EPA shall publish, and provide to large grocery stores, a publication which describes the risks and benefits of pesticide residues on food purchased in those
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stores by consumers, a listing of those pesticides for which the limited benefits-based tolerances have been issued, and recommendations for consumers on how they can reduce dietary exposures to pesticides in a manner consistent with maintaining a healthy diet. EPA developed and distributed a brochure, Pesticides and Food, to 30,000 grocery stores during the winter of 1999. Copies also went to public health officials, libraries, and the medical community. Over 4 million copies are in circulation. The brochure also can be found on EPA's Office of Pesticide Programs' Web site (www.epa.gov/pesticides/food). 31.3.2.7 Estrogenic Substances Screening Program FFDCA §408(p) states the EPA shall "develop a screening program, using appropriate validated test systems and other scientifically relevant information, to determine whether certain substances may have an effect in humans that is similar to an effect produced by a naturally occurring estrogen, or such other endocrine effect as the Administrator may designate." The Agency was given two years to develop the screening program, another year to implement it, and four years to report on its findings. The Agency established an advisory committee to assist it with the development of the program. In August 1998, the committee released its recommendations, most of which were adopted by the Agency. Implementation of the program began soon thereafter, with the focus being on the standardization and validation of the proposed components of the screening and testing batteries that make up the screening program. Details on the program and the Report to Congress (August 2000) can be found on the Web site of EPA's Office of Science Coordination and Policy (www.epa.gov/scipoly/oscpendo/index.htm).
remaining post-1984 chemicals; biopesticides; and the rest of the food-use inert ingredients. EPA published a notice in the Federal Register on August 4, 1997, outlining its plans for carrying out the tolerance reassessment process and identifying the individual substances in each of the three priority groups (U.S. EPA,1997a).
31.4 CURRENT REGULATORY PROCESS The registration of pesticides in the United States is bound by a structure defined by congressional legislation, as interpreted in formal regulations and other less-formal articulations of policy and practice. Proposals for changes and final changes to the regulations are published in the Federal Register. All final, formal regulations also can be found in the Code of Federal Regulations (40 CFR Parts 150-189). The CFR is updated annually to reflect any changes in the regulations that may have been finalized during the year. Daily issues of the Federal Register and the current CFR can be found on the Web site of the U.S. Government Printing Office (http://www.gpo.gov/). Frequently, statements of policy related to pesticide regulation are published by EPA's Office of Pesticide Programs (OPP) as Pesticide Registration (PR) Notices. These PR Notices as well as other documents articulating OPP's regulatory and risk assessment policies and practices can be found on OPP's Web site (http://www.epa.gov/pesticides/). Because regulatory approaches and practices are continually evolving as the state-ofthe-science and its interpretation mature, prospective pesticide registrants are strongly encouraged to meet with pesticide officials before proceeding with data generation and submission of petitions for registration or before executing changes in the registration status of their products.
31.3.2.8 Tolerance Reassessment FQPA required EPA to reevaluate all tolerances and exemptions in effect on the day before enactment of the act. A schedule was imposed that was to assure that 33% of such tolerances and exemptions were reviewed within 3 years of enactment; a second 33% within 6 years; and the remaining number within 10 years. Priority was to be given to those tolerances or exemptions that appeared to pose the greatest risk to public health (i.e., review the "worst first"). EPA divided all chemicals into three groups, with Group 1 containing the pesticides that appeared to pose the greatest risks. Group 1 is made up of several subgroups: organophosphorus compounds (OPs); carbamates; pesticides which had previously been characterized as probable human carcinogens (Groups Bland B2) according to EPA's classification scheme published in its cancer risk assessment guidelines in 1986; high-hazard food-use inert ingredients; and any chemicals that exceed their reference dose (RID) by unacceptable levels. Priority Group 2 contains those pesticides characterized as possible human carcinogens (Group C) and all reregistration chemicals which remained unfinished in 1996 (i.e., of those registered before 1984). Priority Group 3 contains all remaining pre-FQPA chemicals for which reregistration eligibility decisions already had been made by August 1996; all
31.4.1 REGISTRATION Registering a pesticide product for use in the United States under the regulations of the Federal Insecticide Fungicide and Rodenticide Act is equivalent to acquiring a federal license to sell or distribute a product in commerce. To do so without EPA approval is a federal crime. In theory, all pesticide products destined for use in the United States must be registered. A pesticide product is generally made up of more than one constituent. It may include one or more "active" ingredient(s) and one or more "inert" ingredient(s). A pesticide active ingredient is defined as "(1) any substance or mixture of substances intended for preventing, destroying, repelling, or mitigating any pest; (2) any substance or mixture of substances intended for use as a plant regulator, defoliant, or dessicant, and (3) any nitrogen stabilizer" [FIFRA §2(u)]. An "inert ingredient" is any substance or group of similar substances, other than the active ingredient, which is intentionally included in a pesticide product. Both the active ingredient(s) and the formulation(s) which constitute the product(s) are subject to registration requirements. If the intended use of the product includes application to agricultural
31.4 Current Regulatory Process
commodities destined for human or animal consumption, a tolerance or exemption from a tolerance also must be granted under FFDCA §408. New animal drugs, animal feeds containing a new animal drug, and liquid chemical sterilants for use on critical or semicritical medical devices are excluded from the definition of "pesticide." Although FIFRA §3 requires the registration of all materials which meet the definition of "pesticide" (i.e., either an active or inert ingredient) this section along with several other sections of the law provides for exemptions. For instance, exemptions may be granted if the material is being transferred from one site of an establishment to another when both are operated by the same producer, if it is regulated by another federal agency (e.g., certain biological control agents and human drugs), if it is "of a character which is unnecessary to be subject to this Act" [FIFRA §25(b)], or if it is in the preregistration status of having been granted an experimental-use permit under FIFRA §5 or an emergency use under FIFRA § 18. Approval of a registration is dependent upon the successful fulfillment of a series of data requirements, among other factors. The number and types of studies to be conducted vary with the intrinsic chemistry, anticipated inherent toxicity, and proposed use pattern of the pesticide. Pesticides of conventional chemistry proposed for use on agricultural commodities generally require the greatest amount of information, whereas nonfooduse conventional chemicals, antimicrobials, and biopesticides such as microbials and biochemicals generally require less. Part 158 of 40 CFR presents the regulatory roadmap specifying the types and amounts of data and other information needed by EPA to decide whether to approve an application for a new or amended registration or reregistration under FIFRA §3, for an experimental-use permit under FIFRA §5, or for a emergency exemption under FIFRA § 18. The data requirements specified in this part cover the areas of product chemistry, toxicology for human health and terrestrial mammals, wildlife and aquatic toxicology, nontarget insects (e.g., honey bees), environmental fate, aerial drift evaluation, reentry protection (primarily in the occupational setting), plant protection, product performance, residue chemistry (for food uses), and biochemical and microbial pesticides. Some of these kinds of data are always required for the evaluation of some or all types of products. Other kinds of data are required only under certain conditions; that is, if the product's proposed pattern of use, the results of earlier studies, or other circumstances warrant the development of such data. For example, the acute delayed neurotoxicity study in the hen is required only if the pesticide is an organophosphate or a metabolite thereof and causes inhibition of acetylcholinesterase or is structurally related to a substance that is known to cause delayed neurotoxicity. Another example would be the requirement to develop residue chemistry data only if the pesticide is proposed for use on food crops. Some, but not all, subparts of Part 158 have been modified since their original promulgation in 1984. Other data requirements have been added without the benefit of formal promulgation of regulations. This has occurred because the state-of-the-science has evolved and matured in many areas in
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the intervening years. The Office of Pesticide Programs has modified many of its data requirements, communicating these changes via Pesticide Regulation Notices and other written materials. Thus, although the current Part 158 establishes data requirements that are applicable to various general use patterns, some unique aspect of a proposed use and/or the possibility of modification of the original data requirements for any use, general or unique, argues strongly for consultation between the prospective registrant and the Agency before beginning any data generation or information development. 31.4.2 REREGISTRATION AND REGISTRATION RENEWAL
Four times, over approximately 25 years (1972, 1978, 1988, 1996), Congress acted to require EPA to update the registration status of existing pesticide products. The most proscriptive directive was introduced in the amendments to FIFRA in 1988. The reregistration scheme articulated at that time remains in place, now made more complex by the requirement to reassess all tolerances in place at the time FIFRA and FFDCA were amended in the Food Quality Protection Act in 1996. The reregistration directives in the 1972 and 1978 amendments to FIFRA covered those pesticides registered up to that point. The accelerated reregistration program under FIFRA 1988 (FIFRA §4) encompassed all pesticide active ingredients initially registered before November 1, 1984. At that time, there were approximately 1150 active ingredients and over 20,000 product formulations registered in the United States. Because many of these 1150 active ingredients were related to one other (e.g., different salts of the same substance, such as sodium and calcium hypochlorite), they were organized into about 600 "cases" or groups of related pesticide active ingredients. These 600 cases were divided into four lists: List A-List A, which contains most of the food use pesticides, consists of the 194 chemical cases (or 350 individual active ingredients) for which EPA had issued registration standards prior to FIFRA 1988. Each registration standard document summarized the data available for a pesticide, called in any additional studies needed for reregistration, and required necessary product labeling changes. Lists B, C, and D-The remaining pesticides requiring reregistration, and for which no registration standard had been developed in previous reregistration attempts, were divided into three lists based on their potential for human exposure and other factors, with List B containing pesticides of greater concern and List D pesticides of less concern. Some of the classification criteria included the potential or known occurrence of residues in food or drinking water, significance of outstanding data requirements, potential for worker exposure, Special Review or restricted-use status, and unintended adverse effects on animals and plants.
FIFRA 1988 established a reregistration process consisting of five phases, with time frames and responsibilities for both
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EPA and the pesticide producers or registrants. The pesticides on Lists B, C, and D went through all five phases. Because EPA had already substantially reviewed them under the Registration Standards program, the List A pesticides moved directly to Phase 5.
Phase 1, list active ingredients-As required, EPA published Lists A, B, C, and D within 10 months ofFIFRA 1988 (by October 24, 1989) and asked registrants of these pesticides whether they intended to seek reregistration. Phase 2, declare intent and identify studies-Phase 2 required registrants to notify EPA whether or not they intended to reregister their products; to identify and commit to providing necessary new studies; and to pay the first installment of the reregistration fee. During this phase, EPA issued guidance to registrants for preparing their Phase 2 and Phase 3 responses. Phase 2 activities were completed in 1990. Nearly 250 cases, which included nearly half of the existing products, did not proceed past this phase, as the registrants chose not to support their continued registrations. Phase 3, summarize studies-During Phase 3, following EPA guidance, registrants were required to submit summaries and reformat acceptable studies, "flag" studies indicating adverse effects, recommit to satisfying all applicable data requirements, and pay the final installment of the reregistration fee. Phase 3 ended in October 1990. Phase 4, EPA review and data call-in-During Phase 4, EPA reviewed all Phase 2 and Phase 3 submissions and required registrants to meet any unfulfilled data requirements within four years. Phase 4 was completed in 1993. Phase 5, reregistration decisions-In this final phase, which remains ongoing, EPA reviews all the studies that have been submitted and decides whether or not the active ingredient(s) and the pesticide products containing the active ingredient(s) are eligible for reregistration-whether the data base is substantially complete, and whether or not the pesticide causes unreasonable adverse effects to humans or the environment when used according to product labeling. EPA also considers whether the pesticide meets the new safety standard of the FQPA and conducts tolerance reassessment for those pesticides which have food uses. The results of the Agency's review are presented in a Reregistration Eligibility Decision (RED) document. Products containing the pesticide active ingredient are reregistered after certain product-specific data and revised labeling are submitted and approved. All the active ingredients in a pesticide product must be eligible before the product is considered to be reregistered. Because of the FQPA requirement that pesticides sharing a common mechanism of toxicity with other substances (with or without pesticidal uses of their own) shall be evaluated for inclusion in a cumulative risk assessment with these other substances, the Agency has been issuing Interim REDs (IREDs) for some pesticides, reflecting its judgment concerning reregistration eligibility based solely on the individual pesticide's aggregate risk assessment, but reserving judgment on full reregistration eligibility until the cumulative risk assessment process is completed.
The timetable for reregistration that had been established in response to the congressional directives in FIFRA 1988 required modification following the passage of FQPA. Among the new dimensions added in FQPA was the requirement to reassess, within a lO-year time frame, all previously granted tolerances and exemptions from the requirement for a tolerance against the new safety standard that FQPA had established. This directive applied to tolerances (and exemptions) for all food-use pesticides, without regard to the date of their original registration. The consequences of this directive were, among others, that those food-use pesticides for which REDs already had been completed under the FIFRA 1988 reregistration process needed to be revisited for tolerance reassessment. Some of these also would need to be considered for inclusion in a cumulative risk assessment. Because EPA is using the reregistration program to accomplish tolerance reassessment, the timetable for completion of reregistration has been extended to encompass the 10-year time frame for tolerance reassessment, 1996-2006. At the end of 2000, reregistration was about 75% complete. Persons interested in obtaining details on the status of individual cases are referred to EPA's Office of Pesticide Programs Web site, where status reports and other materials on reregistration can be found (http://www.epa.gov/pesticides/reregistration. htm). FQPA also required EPA to establish a new registration review program ("registration renewal"). This new program obligates EPA to review every registered pesticide on a IS-year cycle. This new program would include all pesticides registered since November 1, 1984, as well as those that had been through earlier reregistration processes. Implementation of such a program would assure that pesticides are being reviewed periodically and updated to meet current scientific and regulatory standards. 31.4.3 SPECIAL REVIEW EPA not only has the authority to register pesticides, but also to cancel, suspend, or modify the registration of any pesticide or use of such pesticide that the Agency has determined to have the potential to "cause unreasonable adverse effects on the environment" [FIFRA §6(b )]. FIFRA §(2)(bb) states that, in making a final judgment whether to cancel or modify the conditions of registration of a pesticide or any of its uses, the Agency must weigh the potential for adverse effects against the costs and benefits ("economic, social, and environmental") derived from the use(s) of the pesticide. Dietary risks must be judged against the "reasonable certainty of no harm" safety standard under Section 408 of FFDCA. Risks related to use of public health pesticides are weighed against the health risks such as those from the diseases transmitted by the vector to be controlled by the pesticide. The formal procedure for conducting the necessary regulatory assessment under FIFRA §6 is commonly known as Special Review. Regulations governing the Special Review process are articulated in 40 CFR Part 154. The formal Special Review
References
process which includes a FIFRA §6 cancellation or suspension hearing is resource-intensive and time-consuming. In practice, the Agency has more often used less formal procedures to achieve the same goal of reducing the potential risks to acceptable limits. In more recent times, the reregistration process has been the principal mechanism for negotiating changes in the registration status of pesticides. It is more efficient and, whereas the formal Special Review process generally is a dialogue only between the Agency and the registrant(s) (and, sometimes, the hearing judge), with user groups free to submit benefits and other economic information to USDA, the less formal procedure encourages and supports a more active role for user groups and other interested stakeholders, such as public interest groups and the public health community. Persons interested in obtaining details on the status of chemicals in Special Review are referred to EPA's Office of Pesticide Programs Web site, where its report is available (http://www.epa.gov/docs/SpeciaIReview/srOOstatus.pdf).This report is updated annually. Risk reduction measures taken as a result of assessments conducted during the reregistration process are detailed in individual REDs and IREDs and summarized periodically in status reports. These can be found on EPA's Office of Pesticide Programs' Web site, where status reports and other materials on reregistration can be found http://www.epa.gov/pesticides/reregistration.htm). Portions of this Web page are updated annually at a minimum.
31.5 WEB SITES Government Printing Office Code of Federal Regulations and Federal Register (http://www.gpo.gov/) U.S. Environmental Protection Agency Estrogenic Substances Screening Program (http://www.epa.gov/scipoly/oscpendo/index.htm) Pesticides and Food brochure (http://www.epa. gov/pesticides/food) Reregistration (http://www.epa.gov/pesticides/reregistration.htm) Special Review (http://www.epa. gov/docs/SpecialReview/srOOstatus. pdf) Statements of policy (http://www.epa.gov/pesticidesl)
REFERENCES Code of Federal Regulations (1990). 40 CFR Parts 150-189. Protection of the environment. Subchapter E-Pesticide Programs. FACTA (1990). Food, Agriculture, Conservation, and Trade Act of 1990, Public Law No. 101-624, secs. 1491-1496, 104 Stat. 3359. FEPCA (1972). Federal Environmental Pesticide Control Act of 1972, Public Law 92-516,86 Stat. 973. FFDCA (1954). Federal Food, Drug, and Cosmetic Act of 1954 Miller Amendment to FFDCA §408, Public Law No. 518, 68 Stat. 511.
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FFDCA (1958). Federal Food, Drug, and Cosmetic Act of 1958 Food Additives Amendment to FFDCA §409, Public Law 85-929, 72 Stat. 1785. FIA (1910). Federal Insecticide Act of 1910, Chap. 191,36 Stat. 331. FIFRA (1947). Federal Insecticide, Fungicide, and Rodenticide Act of 1947, Public Law No. 80-104,61 Stat. 163. FIFRA (1964). Federal Insecticide, Fungicide, and Rodenticide Act of 1964, Public Law 88-305,78 Stat. 190. FIFRA (1975). Federal Insecticide, Fungicide, and Rodenticide Act of 1975, Public Law No. 94-140, 89 Stat. 751. FIFRA (1980). Federal Insecticide, Fungicide, and Rodenticide Act of 1980, Public Law No. 96-539, 94 Stat. 3194. FIFRA (1988). Federal Insecticide, Fungicide, and Rodenticide Act Amendments of 1988, Public Law No. 100-532, 102 Stat. 2654. FPA (1978). Federal Pesticide Act of 1978, Public Law No. 95-396, 92 Stat. 819. FPA (1996). Food Quality Protection Act of 1996, amending the Federal Insecticide, Fungicide, and Rodenticide Act and the Federal Food, Drug, and Cosmetic Act, Public Law No. 104-170, 11 Stat. 1513. Nixon, R. M., President (1970). Reorganization Plan No. 3 of 1970. 40 CFR pt. I and Fed. Reg. 35, 15623. U.S. Environmental Protection Agency (1996). "Is an Additional Uncertainty Factor Necessary and Appropriate to Assess Pre- and Postnatal Developmental and Reproductive Effects in Infants and Children Exposed to Pesticide through Chronic Dietary Exposure?" Presented to the FIFRA Scientific Advisory Panel, October 1996. U.S. Environmental Protection Agency (1997a). Raw and processed food schedule for pesticide tolerance reassessment. Fed. Reg. 62(149), 4201942030, August 4, 1997. U.S. Environmental Protection Agency (1997b). "Guidelines for Expedited Review of Conventional Pesticides under the Reduced-Risk Initiative and for Biological Pesticides." Pesticide Registration Notice 97-3, September 4, 1997. U.S. Environmental Protection Agency (1998a). "Use of IOx Safety Factor to Address Special Sensitivity of Infants and Children to Pesticides." Presented to the FIFRA Scientific Advisory Panel, March 1998. U.S. Environmental Protection Agency (1998b). "Standard Operating Procedures of the Health Effects Division's FQPA Safety Factor Committee." Presented to the FIFRA Scientific Advisory Panel, December 1998. U.S. Environmental Protection Agency (I 999a). "Guidance for Identifying Pesticide Chemicals and Other Substances Which Have a Common Mechanism of Toxicity." February 1999. U.S. Environmental Protection Agency (l999b). "The Office of Pesticide Programs' Policy on Detennination of the Appropriate FQPA Safety Factor(s) for Use in the Tolerance-Setting process." May 1999. U.S. Environmental Protection Agency (1999c). "Proposed Guidance on Cumulative Risk Assessment of Pesticide Chemicals That Have a Common Mechanism of Toxicity: Issues Pertaining to Hazard and Dose Response Assessment." Presented to the FIFRA Scientific Advisory Panel, September 1999. U.S. Environmental Protection Agency (l999d). "Chapter 4: Exposure Assessment and Characterization and Chapter 6: Estimation and Characterization of Cumulative Risk. Proposed Guidance on Cumulative Risk Assessment of Pesticide Chemicals That Have a Common Mechanism of Toxicity." Presented to the FIFRA Scientific Advisory Panel, December 1999. U.S. Environmental Protection Agency (2000a). "End Point Selection and Detennination of Relative Potency in Cumulative Hazard Assessment: A Pilot Study of Organophosphorus Pesticide Chemicals." Presented to the FIFRA Scientific Advisory Panel, September 2000. U.S. Environmental Protection Agency (2000b). "Cumulative Risk: A Case Study of the Estimation of Risk from 24 Organophosphate Pesticides." Presented to the FIFRA Scientific Advisory Panel, December 2000.
CHAPTER
32 Risk Assessment for Acute Exposure to Pesticides Roger C. Cochran* Department of Pesticide Regulation, California Environmental Protection Agency
32.1 INTRODUCTION
32.2 TOXICOLOGICAL DATA
Regulatory agencies have tended to focus their assessments of pesticide risk on the potential for toxicological effects to arise from repetitive, long-term usage of the chemicals (Barnes and Dourson, 1988; WHO, 1978). This emphasis on developing reference doses (RIDs) for the potential effects of chronic exposure to pesticides may have gained impetus from public concerns about cancer or the impact of pesticides on the environment (Carson, 1962; NRC, 1987). Consequently, much of the toxicological database required under the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) examines the effects of repetitive, subchronic or chronic dosing by the oral route. From a public health perspective, however, the recorded human illnesses attributed to acute exposures to pesticides may be of greater significance than those connected with potential chronic exposures (Mehler et aI., 1992). Risk assessment, as defined by the National Research Council (1983), consists of five components: (1) Hazard identification encompasses examination of the toxic effects of the chemical; (2) dose-response assessment evaluates the dose level of the chemical necessary to cause manifestation of toxic effects; (3) exposure assessment estimates the amount of the chemical that people are likely to absorb; (4) risk characterization predicts the likelihood that people, exposed to the chemical to the degree estimated, will become ill; and (5) risk appraisal examines the strengths and weaknesses of the estimates of the various toxicological and exposure parameters and expresses the degree of confidence in the projected risks. Acute exposure, here, refers to human encounters with pesticides in the course of one day or less. A pesticide, as defined by the U.S. Environmental Protection Agency (EPA), is any chemical, or mixture of chemicals, intended to be used in preventing, destroying, repelling, or mitigating any pest (Federal Register, 1998).
The first two of the five components needed for risk assessment require an extensive knowledge of the toxicological effects of a chemical. Under FIFRA, the toxicological database for a pesticide is defined by the guideline requirements (EPA, 1984). This database includes acute lethality studies (oral, dermal, and inhalation), subchronic toxicity studies (90-day oral, inhalation, and dermal toxicity; 21128-day dermal toxicity; developmental toxicity; reproductive toxicity), chronic toxicity studies (I-year nonrodent toxicity; oncogenicity; and combined chronic toxicity/oncogenicity), and neurotoxicity studies (neurotoxicity screening battery; 90-day neurotoxicity; developmental neurotoxicity) (Federal Register, 1998). Only a few of these study types contain data that can be used to explore the toxicological effects from a single day's (acute) exposure to a pesticide. Acute lethality studies, for example, use a range of single doses to elicit toxic effects. However, these studies are designed to set toxicity categories for labeling information (EPA, 1998a). Virtually all of the older acute lethality studies, regardless of the route of exposure (oral, dermal, or inhalation), generally do not have data on nonlethal, systemic effects that occurred at less than lethal dosages. The single-dose, neurotoxicity screening battery is currently being required only for those pesticides designed to be neurotoxins (e.g., organophosphates, carbamates, and pyrethroids) (EPA, 1998b). Consequently, data from this test, which includes components of histopathology, tissue and blood chemistry, as well as clinical signs and performance testing, are not available for most pesticides. Thus, data on acute effects for most pesticides have to be teased out of repetitive dosing studies. Subchronic, reproductive, and chronic toxicity studies may have data concerning clinical signs that appear within 1-2 days at the beginning of the studies. All other data on potential systemic toxicity in these study types are obtained at the end of the study period and cannot be attributed to acute toxicity. Developmental toxicity studies provide an exception. Because develop-
*The opinions expressed in this chapter represent the views of the author and do not necessarily reflect the views and policies of the Department of Pesticide Regulation. The mention of trade names or commercial products does not constitute endorsement or recommendation for use. Handbook of Pesticide Toxicology Volume 1. Principles
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Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
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mental toxicity may be manifested as the result of a single dose (EPA, 1991; Ogata et aI., 1984; Schardein, 1985), it is assumed, in the absence of data to the contrary, that the observed developmental effects are elicited from a single dose. This assumption mayor may not be valid. Nonetheless, developmental toxicity studies (see Chapter 16) tend to be a major source of critical no observed effect levels (NOELs) for conducting risk assessments on potential acute exposures to nonneurotoxic pesticides. Although a developmental endpoint for exposure to toxins is only relevant in women of child-bearing age, the assumption that all other population subgroups are as sensitive results in margins of safety (MOSs) that protect the health of these other subgroups for other endpoints that may occur at higher dosages. The MOS is defined as the ratio of the critical NOEL to the estimated exposure. Published research studies may also provide sufficient data for dose-response assessment. These studies, however, tend to be designed to clarify the mechanism of action of a specific type of pesticide toxicity. Nonetheless, the peer-reviewed reports sometimes describe a range of concentrations used to elicit an effect from a single dose. Such a study may provide the basis for a regulatory NOEL, particularly in the case of experiments with human subjects. The main drawbacks to these published studies are (l) the lack of individual animal data because they are typically not reported or archived and, thus, (2) the need to rely on the author's interpretation of the results. Under FIFRA, pharmacokinetic data are sought to obtain information on how a pesticide is absorbed, distributed, biotransformed, and excreted, as well as to aid in understanding the mechanism of toxicity (EPA, 1998c). Information may also be obtained about potential tissue-specific accumulation and induction of biotransformation. Most of the pharmacokinetic data are derived from studies using the oral route of exposure. Dermal pharmacokinetic studies tend to consider solely dermal penetration and/or absorption. Pharmacokinetic studies on the inhalation of pesticides are comparatively rare, seemingly limited to fumigants. Pharmacokinetic data can have a profound effect on the dose-response assessment for a pesticide. The estimated absorbed dose of a pesticide necessary to cause toxic effects may be modified downward if there is evidence of less than 100% absorption through the route used in the dose-response assesment. Information regarding bioavailability via the oral route is useful as many pesticides and their metabolites are excreted in variable amounts in the feces. Estimations of absorbed dosages from inhalation toxicity studies rely on default assumptions concerning breathing rates, tidal volumes, and chemical retention and absorption to estimate absorbed dosages (Raabe, 1986, 1988; Zielhuis and van der Kreek, 1979). Such estimates, when derived from whole-body inhalation studies, can be confounded by the fact that rats exposed to dusts or chemical vapors via whole body absorb 5-8 times more material than rats exposed via nose only (Blair et aI., 1974; Hext, 1991; Iwasaki et aI., 1987; Jaskot and Costa, 1994; Landry et aI., 1986; Langard and Nordhagen, 1980; Ty1 et al., 1995; Wolff et al., 1982). The additional absorption noted in whole-body inhalation exposure
studies appears to be due to an unquantifiable oral component, possibly from grooming behavior (Cochran et aI., 1997). Even nose-only inhalation toxicity studies may have a significant oral component due to grooming activity (Hext, 1991).
32.3 EXPOSURE DATA The second, and equally important, half of the risk assessment equation is the estimate of human exposure. The chief source of exposure to pesticides through the oral route is from the diet (see Chapter 19; Cochran et aI., 1994). There can also be an oral contribution from hand-to-mouth activity in adults and children or pica in children (Binder et aI., 1986; Calabrese and Stanek, 1992; Calabrese et aI., 1989, 1991; Carlisle, 1992; Clausing et aI., 1987; EPA, 1996). Pica in children, however, appears to be highly unusual, as only a single instance of intentional imbibing of dirt was reported out of more than 200 children whose soil ingestion from hand-to-mouth activity was documented in the preceding publications. Currently, dietary exposures are estimated by most governmental agencies through a process that combines data on dietary consumption with data on pesticide residues measured on food (Cochran et aI., 1995a; FAOIWHO, 1988, 1997). Dietary consumption data are generally derived from government surveys (Cochran et aI., 1995a; FAOIWHO, 1997; Trichopoulou, 1994; USDA, 1989-1991). Data for potential pesticide residues associated with EPA or European Union (EU) label-approved direct food uses, as well as information about possible secondary residues in animal tissues, are also necessary for estimating human dietary exposures. These data are derived from governmental monitoring programs (CD PR, 1997; FAOIWHO, 1999; USDA, 1996). However, dietary exposure to pesticides is only a fraction of the total human exposure experience. Much of human occupational (persons engaged in the process of pesticide application) or nonoccupational (other than dietary) exposure to pesticides results from the handling of pesticides or other activity patterns that place people in contact with the pesticides. In general, most of the occupational and nonoccupational exposure to pesticides is through the dermal and/or inhalation routes (Ross et aI., 1992; Wolfe, 1976). Exposure estimates for these scenarios are based on environmental monitoring, passive dosimetry, or biological monitoring of individuals involved in the active handling of pesticides or engaged in activities in areas treated with those pesticides (Bonasall, 1985; Lavy and Mattice, 1986). An extensive, detailed discussion of the techniques used for estimating occupational and nonoccupational exposures may be found in Chapter 21. Environmental monitoring involves measurements of pesticide concentrations in the ambient air and on surfaces. The translation of measured air concentrations into an estimated absorbed dose for humans requires assumptions on respiratory frequency, volume, and absorption of the pesticide (EPA, 1996; Raabe, 1986, 1988; Zielhuis and van der Kreek, 1979). Estimation of human dermal exposure from surface concentrations of
32.4 Examples
pesticides in the environment relies on the precision of various generic transfer factors (EPA, 1996; Pandian et aI., 1999). Passive dosimetry gauges air concentrations in the breathing zone and measures dermal concentrations of pesticides through the use of hand washes, dermal patches, and/or articles of clothing (Wolfe, 1976). The same assumptions for inhalation are used with air concentrations of pesticides measured in the breathing zone as were used for those detected in the ambient air. Concentrations of pesticides extracted from monitoring patches attached to the skin are assumed to be representative of chemical concentrations over a specified body surface area (Wolfe, 1976). A single value, based on submitted, chemical-specific studies (a default of 100% has been used if specific data were not available), serves as the basis for estimating the absorbed dose (EPA, 1992a). It is known that the percentage of pesticide absorbed through the skin varies inversely with the concentration of the chemical (Wester and Maibach, 1976). However, at the present time, there are no scientific models available that examine the effect of multiple concentrations of pesticides on the skin, separated spatially and/or chronologically, on the absorbed daily dosage (Wester and Maibach, 1993). Biological monitoring provides an estimate of the aggregate exposure to a pesticide from all routes. Unfortunately, very few biomonitoring studies have been conducted for more than a handful of pesticides. Chemical-specific, human stoichiometric data are essential to the process of estimating absorbed dosages from excreted pesticide metabolites. Consequently, the principal limiting factor seems to be the lack of human pharmacokinetic data on most pesticides. Chemical-specific information is preferred for exposure data from either environmental monitoring or passive dosimetry. Surrogate exposure data (from pesticides with similar chemical and physical properties, as well as similar preparation and application practices) and generic databases, such as the Pesticide Handlers Exposure Database (PHED, 1995), are used as substitutes. The use of surrogate exposure data increases the level of uncertainty in exposure estimates. Differences in volatility between the chemical under consideration and surrogate chemicals may affect air concentrations in an unquantifiable manner. Likewise, differences in chemical properties could affect transfer factors, clothing penetration, and dermal adsorption. Differences in application rates cause assumptions to be made on the relationship between the amount of chemical handled and the amount of exposure through all routes. The principal difficulty associated with the use of PHED to estimate exposure data is that the data subsets, which are combined by the program to form work categories, are not homogeneous (van Hemmen, 1992). For example, one source of variability is that each of those studies has a different minimum detection level for the analytical method. It should be noted that the detection of dermal exposure to the body regions is not standardized. Some studies observe exposure to only selected body regions, such as the hands, arms, and face, with other body regions considered 100% protected from exposure by work clothing. Other studies have more extensive dermal measurements.
693
Consequently, the subsets derived from the database for dermal exposure have different numbers of observations for each of the body regions. Finally, the PHED database is predicated on the relationship between the amount of pesticide handled and the degree of occupational exposure. Yet, for example, within the data set used to estimate exposures for groundboom applications without the presence of a cab, there is no correlation between the amount of pesticides being used and the amount of dermal or inhalation exposures that workers receive. The net effect of this lack of correlation between exposure and the amount of chemical used is an inability to predict, with accuracy, what exposures any worker will receive in a given work category. When PHED is used for estimating potential acute (single day) occupational exposures, the only data point that can be provided is the average exposure value. Because the variability in each of the data subsets in a given category is unrelated to that in any of the other data subsets, it is not possible to estimate the overall variability in exposure. Yet, in a given study, there is variability in worker exposure. Even though individuals confine their activities to label-approved personal protective equipment and labor practices, they do not receive the same exposure. Depending on the shape of the distribution curve, the average exposure value may represent the maximum potential exposure of as few as 50% of the workers. The amount of exposure for the other workers, who also follow label requirements, could be greater-though the magnitude of the exposure cannot be calculated.
32.4 EXAMPLES Thus far, we have examined the nature of the toxicological and exposure databases used in generating a risk assessment. How the data fit together can best be explored through critical examination of some examples of completed risk assessments for acute exposure to pesticides. The examples provided are risk assessments conducted for the California Department of Pesticide Regulation (CDPR). 32.4.1 ETHOPROP The first case study is the risk characterization document (RCD) for ethoprop (Cochran et aI., 1995b). Ethoprop (O-ethyl-S,Sdipropyl phosphorodithioate) is an organophosphate pesticide used as an insecticide, nematicide, and fungicide (suppression of white mold on peanuts) on food and nonfood crops. The oral LDso for ethoprop was 61 mglkg in male rats and 33 mglkg in female rats. Examination of the toxicological database indicated that rabbits were the most sensitive laboratory species to ethoprop exposure, with a dermal LDso of 24 mglkg. Clinical signs of acute toxicity were characteristic of cholinesterase inhibition and included diarrhea, excessive urination, lacrimation, tremors, and convulsions. As the principal route of exposure for most pesticide applicators using ethoprop was through the skin, it would have been
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preferable to use the dose-response data of adverse effects observed in short-term dermal toxicity studies as the basis for calculating margins of safety for workers with short-term exposure to ethoprop. However, none of the submitted or published data established a single-dose dermal NOEL for clinical signs in rats or rabbits using technical-grade ethoprop. In rats, a single dermal dose of a formulation, Mocap® 6EC, resulted in cholinergic signs (salivation, irregular respiration, prostration, and morbidity), with a NOEL of 160 mg formulationikg. Dermal exposure of rabbits to Mocap 6EC resulted in clinical signs (ataxia, depression, dilation of pupils, excessive salivation, and loss of the righting reflex) with a NOEL of 8.7 mg active ingredientlkg. However, examination of the database suggested that ethoprop, diluted in formulations, had more toxicity through the dermal route than technical-grade ethoprop. This suggested that the inert ingredients in the formulation had facilitated the passage of ethoprop through the skin. As the inert ingredients in formulations are frequently changed, none of the short-term dermal toxicity studies were considered appropriate as the basis for assessing the risk of short-term occupational exposures to ethoprop. Instead of a dermal NOEL, the absorbed dose from an oral NOEL was used to estimate margins of safety from short-term exposure to ethoprop. Short-term oral NOELs were derived from developmental studies in rats and rabbits and from a single-dose neurotoxicity study in rats. Ethoprop did not produce developmental malformations in rats. Fetal toxicity in rat studies was manifested as decreased fetal weight. In rats, the lowest observed effect level (LOEL) for maternal toxicity [cholinergic signs-soft stools (8/25 animals) and anogenital staining (3/25 animals)] was 18 mg/kg-day with a NOEL of 9 mg/kg-day. The effects were manifested after 2 days of dosing. A lower NOEL for cholinergic signs (1.6 mg/kg-day) was reported in an earlier rat developmental study. However, the happenstance of dose selection appeared to determine this NOEL. Considering the two studies together, the NOEL (9 mg/kg-day) from the later study was not precluded as a possible NOEL for the earlier study as well. The single-dose LOEL for cholinergic signs, reduced motor activity, and reduced scores on the functional observational battery was 25 mg/kg with a NOEL of 5 mg/kg in both male and female rats. Again, 9 mg/kg was not precluded as the actual NOEL. Developmental toxicity was not observed in rabbits at any dose. Maternal toxicity in rabbits, characterized by signs of cholinesterase inhibition [soft stools (2/8 animals), anogenital staining (2/8 animals), and death (1/8 animals)], was observed by day 2 at 5.0 mg/kg-day, with a NOEL of 2.0 mg/kg-day. A lower NOEL in rabbits, 0.125 mg/kg for decrement in maternal weight gain (14%), was noted in an earlier study. However, the endpoint (decrement in weight gain) required 12 days of dosing to be manifested. Consequently, this NOEL could not be used to assess health risks associated with potential single-dose exposures to ethoprop. The oral NOEL (2 mg/kg-day) for maternal toxicity in rabbits (cholinergic signs and death) was used to assess the health risks from potential short-term exposures to ethoprop.
Exposure estimates for the various occupational categories were based on monitoring data from ethoprop exposures and calculations from monitoring data for surrogate active ingredients (diazinon, turbofos) with similar application rates and chemical properties. These estimates were based on 8-h workdays during the application season and assumed 100% dermal absorption, as no dermal absorption data were available. This health protective assumption probably overstated the exposure, as reported in vivo human dermal absorption for five other organophosphate pesticides ranged from 8 to 46% (Wester and Maibach, 1985, 1993). Uptake of ethoprop via the inhalation route was assumed to involve 50% retention by the lungs and 100% absorption (Raabe, 1986, 1988). The average daily dosage (ADD), actually the geometric means of exposure, ranged from 0.2 J.lg/kg-day for irrigators to 139 J.lg/kg-day for incorporators (workers incorporating the applied ethoprop into the soil) working with the EC formulation. The use of geometric mean values underestimates potential short-term exposures of populations of workers (EPA, 1992b). Consequently, the 95th percentile [geometric mean x (standard deviation)1.645] of short-term worker exposure was also examined. The 95th percentiles of short-term exposure for loader/applicator/incorporators working with the 5G and lOG formulations were 71 and 45 J.lg/kg-day, respectively. All but two of the tolerances for ethoprop are for "negligible residues," as the EPA does not expect that any residues of ethoprop will be found on raw agricultural commodities (EPA, 1988a). Nonetheless, to be health protective, CDPR has a policy of conducting dietary risk assessments if tolerances exist for a pesticide on edible commodities. The CDPR surveillance programs from 1987 to 1991 indicated that ethoprop levels in raw agricultural commodities (RACs) were nondetectable. The minimum detection limit (MDL) was 0.05 ppm. Crops monitored in this survey between 1989 and 1990 were cabbage and potatoes, where ethoprop was mostly used. Field studies indicated that ethoprop residues on registered crops were less than 0.02 ppm. Examination of the FDA program for fiscal year (FY) 1985FY 1990 revealed only two values. These were 0.680 ppm in strawberries (1987) and 0.140 ppm in apples (1989). The mean theoretical (acute) daily dietary exposure for all population subgroups ranged from 0.02 to 0.08 J.lg/kg-day, with children (1-6 years of age) having the highest theoretical exposure. Although theoretical acute dietary exposure was combined with acute occupational exposure in the RCD, the dietary contribution was negligible. Consequently, it is not considered here. The MOS for exposure to ethoprop was calculated as the ratio of an oral NOEL, established in laboratory animal studies, to the potential exposure dosage (greater than 95% through the dermal route) estimated for the human population. The MOSs for potential acute exposure, based on an oral NOEL of 2.0 mg/kg-day for cholinergic signs and death in rabbits, ranged from 14 for incorporators using the EC formulation to 10,000 for the irrigators (Table 32.1). If the 95th percentile of shortterm exposure were considered for workers using the 5G and lOG formulations, the MOSs would be 29 and 40, respectively.
32.4 Examples Table 32.1 Margins of Safety for Potential Acute (Daily) Exposures to Ethoprop Acute
Work task
Mosa
EC formulation MixerlIoader/applicator
32 14
Incorporator
10,000
Irrigator 50 formulation Loader/applicator/incorporator
400
100 formulation Loader/applicator/incorporator
425
aBased on an NOEL of 2.0 mg/kg-day for cholinergic signs and death in a rabbit study MOS = NOEL(2000 ~g/kg-day) ADD
In the absence of scientific evidence to the contrary, effects reported in laboratory studies are expected to occur in humans at similar dosages. When the NOEL is from a laboratory animal study, a MOS of 100 is generally considered adequate for protection against potential acute toxicity of a chemical. This uncertainty factor assumes that humans are 10 times more sensitive to the acute effects of a toxin than are laboratory animals and that the difference in susceptibility to the toxicity of a compound within the human population spans only an order of magnitude (Davidson et aI., 1986; Dourson and Stara, 1983, 1985; EPA, 1986). If the critical NOEL is derived from a human study, a different number, 10, is used, incorporating the single uncertainty factor for human variability. After the RCD for cthoprop was released, the manufacturer dropped production of the EC formulation and changed the method of handling the granular formulations in order to reduce the estimated exposure. Certain generic exposure reduction factors were dictated by the addition of personal protective equipment and procedures (EPA, 1997; Thongsinthusak et al., 1993). These changes in handling procedures of the granular formulations theoretically resulted in a substantial decrease in the estimated exposure to ethoprop and a concomitant increase in the estimated MOSs to more than 100 for the 95th percentile of worker exposure. 32.4.2 MEVINPHOS Mevinphos (2-carbomethoxy-l-methyl-vinyl dimethyl phosphate) is an organophosphate insecticide used to control aphids, mites, grasshoppers, cutworms, leafhoppers, caterpillars, and many other insects on a broad range of field, forage, vegetable, and fruit crops. This highly toxic pesticide (rat oral LDso "-'2 mg/kg) had a large number of human illness reports associated with its use (Cochran et al., 1996). Examination of the toxicological database for mevinphos indicated that the principal adverse effects (cholinergic signs) were associated with inhibition of acetylcholinesterase activity. Both plasma and red
695
blood cell cholinesterase activities were significantly reduced compared to controls in several acute studies. The EPA, in its Guidelines for Neurotoxicity Risk Assessment, lists alteration in the degradation of neurotransmitters as a possible adverse effect because it can lead to unwanted changes in the function of the nervous system (EPA, 1998b). However, the guidelines do not specify the level of inhibition of brain acetylcholinesterase activity that constitutes an adverse effect. Even statistically significant brain cholinesterase inhibition caused by organophosphorous insecticides may not lead to cholinergic signs in laboratory animals (Bushnell et aI., 1993, 1994; Chanda and Pope, 1996; Stanton et al., 1994). Organophosphorous insecticide poisoning in humans may lead not only to cholinergic signs but also symptoms, for example, headaches (Ellenhom et aI., 1997), which cannot be ascertained in laboratory animals. Consequently, statistically significant levels of inhibition of brain acetylcholinesterase activity in laboratory animals may be used as a surrogate for this manifestation of an impaired nervous system function, which is detectable only in humans (EPA, 1998b; JMPR, 1998). The depression of plasma or red blood cell cholinesterase activity is generally considered an indication of exposure to a neurotoxic substance, rather than an adverse effect in itself (Carlock et aI., 1999; EPA, 1988b, 1990, 1993, 1998b; JMPR, 1998). The biological significance of inhibition of blood cholinesterase has remained controversial (Carlock et aI., 1999; Chen et aI., 1999; EPA, 1988b, 1990, 1993; JMPR, 1998; Lotti, 1995). The cholinesterase in the blood does not appear to act as a "sink" to reduce paraoxon toxicity, as that same toxicity was not potentiated by prior immunological reduction of blood acetylcholinesterase activity (Padilla et aI., 1992). Consequently, these parameters were not used to characterize the potential risk to humans. As in the case of ethoprop, the principal route of exposure for most mixerlloaders associated with mevinphos use was through the skin. No single-dose dermal toxicity studies were available. However, the toxicological database for mevinphos indicated that the inhibition of brain cholinesterase activity following continuous dosing was likely the result of the last dosage, rather than a cumulative effect of the repeated dosages. Thus, the NOEL (1 mg/kg-day) for significant (p < 0.05) brain cholinesterase inhibition (>50%) in a 21-day dermal toxicity study in rabbits might be considered a single-dose NOEL. The single-dose oral NOEL for neurotoxicity (clinical signs, sensorimotor alterations, reduced neuromuscular performance, and inhibition of brain cholinesterase activity) in the rat was 0.1 mg/kg. Cholinergic signs (loose stool and undescribed other effects) were noted in a study of 20 human subjects following oral ingestion of capsules with mevinphos up to an estimated dosage of 0.036 mg/kg (Rider et aI., 1975). Unfortunately, individual data from that study were unavailable, and a NOEL could not be established. In a subsequent study, a different group of investigators (Verberk, 1977; Verberk and Salle, 1977) built upon the earlier study. A NOEL for cholinergic signs in eight humans was established as 0.025 mg/kg or greater (only one dose level and no cholinergic effects) (Verberk, 1977; Ver-
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berk and Salle, 1977). Although the study ran for 28 days, no cholinergic signs were reported at any time. Thus, the NOEL for cholinergic signs and symptoms (0.025 mg/kg) in humans applies to both acute (single dose) and subchronic (28 days) time periods. The actual I-day NOEL for cholinergic signs may be higher, as a toxicological endpoint caused by a single dose of a chemical can generally be achieved by a lower repetitive dose (Klaassen and Eaton, 1991). For the short-term time frame, this human I-day oral NOEL was approximately the same as the oral NOELs for cholinergic signs from studies on dogs, rats, and rabbits. Because of the uncertainty over the rabbit dermal NOEL, the human oral NOEL (0.025 mg/kg) for cholinergic signs was used in the ReD to calculate margins of safety for potential short-term occupational and acute dietary exposures. The studies and data forming the basis for estimating worker exposure were based on both passive dosimetry data and calculations from foliar residue data. The mean exposure values used for the risk assessment are shown in Table 32.2. The exposure estimates for pilots of fixed-wing aircraft and helicopters, loaders for helicopters, and fiaggers were obtained from studies conducted in Monterey and Imperial counties (Maddy et aI., 1981, 1982). In those studies, the mean absorbed dosage was calculated from the amount of mevinphos collected by cotton gauze and cloth patches placed on the bodies of workers. Approximately 80% of the patches had no detectable levels of mevinphos. Accordingly, each of those patches was assigned a default value equivalent to 50% of the MDL. The measured
or theoretical levels of residue on patches from specific body regions were then multiplied by the total surface area of that zone to obtain an estimate of the dermal exposure. The estimates for each of the zones were then summed for total body exposure. Inhalation exposure was estimated from measured air concentrations of mevinphos in the breathing zone. The 95th percentile of short-term worker exposure ranged from 1.1 f.Lg/kg-day for helicopter pilots to 33.8 f.Lg/kg-day for helicopter mixerlloaders. These values are representative of the maximum acute occupational exposures that workers might be expected to encounter. Potential acute dietary exposure to mevinphos for alllabeled uses, based on the 95th percentile of user-day exposure for all population subgroups, ranged from 1.0 to 3.3 f.Lglkg-day. The potential dietary exposure of the population subgroup of males, aged 20 and over, was chosen as a surrogate sUbpopulation for the purposes of estimating combined occupational and potential dietary exposures. The choice was based on two factors: (1) occupational exposures were derived using passive dosimetry of agricultural workers from this population subgroup, and (2) the dietary exposure values are approximately the same as those of any other population subgroup that might contribute to the agricultural workforce. The potential acute dietary exposure of this population subgroup was 1.3 f.Lglkg-day. This value was added to the mean estimated occupational exposures. However, the theoretical combined acute exposures were probably overestimates of the actual exposures. It was unlikely that the agricultural workers engaged in activities associated with mevinphos
Table 32.2 Estimates of Acute Occupational Exposure to Mevinphos and the Respective Acute MOSs
Work task
ADDa
Acute
Combinedc
Combined
(~g/kg-day)
MOSb
ADD
acute MOS
Helicopters Mixerlloaders
2.4
10
3.7
7
Pilots
0.5
50
1.8
14
Fixed-wing aircraft Mixerlloaders
1.6
16
2.9
9
Pilots
0.5
50
1.8
14
Flaggers
0.04
625
1.34
19
Ground application Mixerlloader/applicators
3.8
7
5.1
5
3.2
8
4.5
5
(open cab) Mixerlloader/applicators (closed cab) Harvesters Field workers (vegetables) Field workers (fruits) Field workers (grapes)
neg.-1.5 0.8-11 0.4-1.1
17-50
1.3-2.8
2-31
2.1-12.3
2-12
25-63
1.7-2.4
11-15
9-14
aThe geometric mean of the estimated absorbed daily dosage (ADD) for helicopter, fixed-wing aircraft, and ground application. For harvesters, the values represent the range of exposures for workers harvesting different commodities. The dermal absorption was 16.8%. Inhalation retention and inhalation absorption were 50% and 100%, respectively, assuming a 75.9 kg body weight for workers. bMOS based on an NOEL of 0.025 mg/kg for cholinergic signs in human studies. cCombined occupational and dietary exposures.
32.4 Examples
use would also be in the 95th percentile of dietary exposure to RACs, each with the maximum measured level of mevinphos residues. The MOSs for mean acute occupational exposures, based on the NOEL of 25 ~g/kg for human cholinergic signs, ranged from 2 (apple harvesters) to 625 (flaggers in enclosed vehicles). If the 95th percentile of short-term exposures were considered for each of the job categories, the MOSs would range from less than 1 (mixerlloaders involved in helicopter applications) to 23 (helicopter pilots). Combining potential acute dietary exposure with mean occupational exposures caused a substantial drop in the MOS for all job categories. The combined MOSs ranged from 2 (apple harvesters) to 19 (flaggers in closed cabs). Arguably, it might have been preferable to use the dermal NOEL of 1 mg/kg for inhibition of brain cholinesterase activity from the 2l-day rabbit study. The specificity of the route of exposure can affect the time course of systemic absorption as well as the chemical nature of the toxin. If this NOEL (l mg/kg) were applied to the calculated mean dermal exposures, the MOSs would then be 69, 100,64, and 76 for mixerlloaders associated with helicopter, fixed-wing aircraft, open-cab ground applications, and closed-cab ground applications, respectively. If the 95th percentile of short-term exposure were considered for these workers, the MOSs would then be 5, 26, 20, and 30, respectively. These margins of safety remain less than the value (100, for a critical NOEL from a laboratory animal study) conventionally recommended to protect people from the toxic effects of a chemical. The greatest uncertainty was associated with the exposure assessment. As more than 70% of the dermal patches analyzed in the occupational exposure studies involving mevinphos and a surrogate pesticide contained nondetectable levels of residues, the accuracy of the occupational exposure estimates were questioned, too. One alternative would have been to use the PHED to estimate worker exposures. However, PHED mean exposure values were approximately the same as those used. A second alternative would have been a biological monitoring study, in which the absorbed dose would have been estimated from urinary metabolites of the parent compound. However, no acceptable study of this type was available. The personal protective equipment and clothing already required for mevinphos handlers was close to the maximum level permitted in California's climate (CCR, 1989). Consequently, it did not appear possible to mitigate the estimated excessive exposures. Before any regulatory action was taken, the manufacturer voluntarily withdrew the registration of mevinphos (EPA, 1994). 32.4.3 PROPOXUR
Propoxur [2(l-methylethoxy)phenol methyl carbamate] is a carbamate insecticide used alone or in combination with other insecticides in interior crack-and-crevice treatments, room foggers, flea and tick sprays, flea and tick collars, ant and cockroach traps, insecticide tapes, ant and cockroach sprays, wasp, bee, and hornet sprays, and flea and tick dips for pets. It does
697
not have any food uses in the United States, but it is approved for crop use in the EU. As was the case with ethoprop, the principal route of exposure for most pesticide applicators using propoxur is through the skin. Consequently, the dose-response data of adverse effects observed in short-term dermal toxicity studies were the initial choice as the basis for calculating margins of safety for workers with short-term exposure to propoxur (Cochran et aI., 1997). A single dermal dose of 2000 mg/kg caused clinical signs (fasciculations, decreased motor activity, hyperreactivity) in rabbits. The single-dose dermal NOEL for clinical signs was 1000 mg/kg in both rabbits and rats. These single-dose dermal NOELs for clinical signs were considerably greater than the oral NOELs in the same species. In a rabbit developmental study, the maternal NOEL (l day) was 10 mg/kg-day, based on cholinergic signs and death at 30 mg/kg. In the rat, the LOEL (30 min) for cholinergic signs (convulsions, reduced motility, apathy, bristling coat) from a single oral dose was 25 mg/kg with a NOEL of 5 mg/kg. The LOEL for maternal toxicity (cholinergic signs) in a rat developmental study was 9 mg/kgday, with a I-day NOEL of 3 mg/kg-day. A single oral dose of 5 mg/kg resulted in cholinergic signs (muscle fasciculations) in dogs, but a dose of 4 mg/kg did not produce any signs. In a single-oral-dose neurotoxicity study, the LOEL for cholinergic signs (excessive chewing and reclining posture) and significant brain cholinesterase inhibition was 2 mg/kg-day. Differences in the effective dose were due to the slower and reduced percentage of dermal absorption compared to oral absorption. Despite the fact that dermal dosing is more germane to human exposure scenarios, the dermal NOELs, 1000 mg/kg for clinical signs in rats and rabbits, were not used as the basis for assessing the risks from acute exposure to propoxur. Nor were oral NOELs for clinical signs in laboratory animals used as the basis for risk characterization. The toxicological basis for characterizing the risk from acute exposure to propoxur was an oral NOEL from a human study. The human oral NOEL was used because (1) the use of human dose-response data eliminates the uncertainty associated with extrapolating to humans from laboratory animal studies and (2) the quality of the dermal toxicity studies was not comparable to the clinical observations in the human study. In the human study, volunteers (number unstated) were reported to have exhibited cholinergic signs (stomach discomfort, blurred vision, moderate facial redness, and sweating) after a single bolus oral dose of 0.36 mg/kg. Doses of 0.2 mg/kg administered every half hour for up to 2 1/2 h (a total of 1 mg/kg) produced no cholinergic signs. Thus, the 30-min NOEL for cholinergic signs in humans following a single bolus dose was 0.2 mg/kg. In the same study, red blood cell cholinesterase activity was depressed about 2% after the first dose and 10% after a total of 5 doses. This indicated a cumulative inhibitory effect on cholinesterase activity by multiple doses of propoxur. The NOELs for clinical signs (specified amounts-up to 1 mg/kgfor specific lengths of time-up to 2 1/2 h) were used to evaluate the health risks from potential acute exposures of different durations to propoxur.
698
CHAPTER 32
Risk Assessment for Acute Exposure to Pesticides
The studies and data, which formed the basis for estimating worker exposure, were based on passive dosimetry using patches attached to clothing. Ethanol hand washes were collected to assess hand exposure, and air levels were monitored with personal pumps. Patches with nondetectable levels of propoxur were given default values equal to 50% of the MDL. The monitored activities in one location took 1.8 h to complete. This was defined as one cycle. The geometric mean exposure values used for the risk assessment are shown in Table 32.3. The 95th percentiles of the absorbed cycle dosage for the respective work tasks were as follows: aerosol (1 %) applicator, 2.30 I-lg/kg-cycle; bait (2%) applicator, 0.61 I-lg/kg-cycle; spray (0.95%) applicator, 0.32 I-lg/kg-cycle; and spray (70WP) applicator, 8.0 I-lg/kg-cycle. Nonoccupational exposures to office workers and home residents may occur through dermal contact with treated surfaces and, to a lesser extent, via inhalation of pesticide vapors. The potential passive exposures of residents to propoxur after crackand-crevice treatment of a home were based on studies submitted by the registrant. The data were derived from wipe samples in various rooms of the home. Analysis of the samples indicated a log-normal distribution of surface residues throughout the house. Air concentrations in the home were more or less constant. It was assumed that infants (6-9 months old) had a body weight of 7.5 kg with 0.45 m2 of surface area-50% of which could be exposed to pesticides. Their breathing rate was
Table 32.3 Mean Acute Exposures to Propoxur and Their Respective MOSs Absorbed cycle dosage Activity
(iJ.g/kg-cycle )a
Acute MOSb
Aerosol (1%) applicator (N = 32)
0.95
842
Bait (2%) applicator (N = 32)
0.19
4210
Spray (0.95%) applicator (N = 32)
0.16
5000
Spray (70WP) applicator (N = 16)
1.47
544
Passive exposure Infant (6-9 months)
1.46
548
Adolescent (12 years)
0.22
3643
Adult
0.37
2000
Active exposure Dog groomer (N = 15)
10.3
97
aGeometric mean of one application-assumes that workers' body weights were 76 kg, dermal penetration was 0.351 %Jh, and respiratory uptake was 50%. The monitored activities in one location took 1.8 h to complete, which was rounded off to 2 h. This is defined as one cycle. For people living in homes treated with crack-and-crevice treatments of propoxur for insect control, the exposures were calculated using default assumptions. It was assumed that infants (6-9 months) had a body weight of 7.5 kg with 0.45 m 2 of surface area-50% of which could be exposed to pesticides. Their breathing rate was 0.5 m3Jh with 100% absorption. For children (12 years), it was assumed they had a body weight of 40.5 kg with 1.37 m2 of body surface area and a breathing rate of 0.9 m3Jh. For adults, the assumptions were 76 kg body weight, 2.0 m2 surface area, and 1 m3Jh breathing rate. b Acute MOSs were based on a 2-h human NOEL of 800 iJ.g/kg for cholinergic signs.
0.5 m 3Jh with 100% absorption through the inhalation route. For children (12 years old), it was assumed that they had a body weight of 40.5 kg with 1.37 m2 of body surface area and a breathing rate of 0.9 m3Jh. For adults, the assumptions were 76 kg body weight, 2.0 m2 surface area, and 1 m 3Jh breathing rate. The geometric means of passive, nonoccupational exposures ranged from 2-h absorbed dosages of 0.22 to 1.4 I-lg/kgday (Table 32.3). Infants, 6-9 months of age, had the highest potential exposure. The study used to estimate exposure to a pet owner for flea control on two dogs involved a biomonitoring study of 15 professional dog groomers who sprayed an average of 20 dogs during an 8-h workday. A metabolite of propoxur, 2-isopropoxyphenol, was measured in urine samples collected from the study's participants. The estimated 24-h absorbed dose of propoxur was based on the amount of the metabolite in the urine. The mean absorbed dose, normalized for two dogs, is presented in Table 32.3. It was assumed that the dogs could be sprayed by a nonprofessional in a period of 2 h. The 95th percentile of the absorbed dosage for an adult engaged in spraying two dogs per day for ticks was 77.1 I-lg/kg-day. The MOSs for mean acute exposure to propoxur, based on the human NOEL of 0.8 mg/kg for a 2-h period, ranged from 97 to 5000 (Table 32.3). The MOSs for the 95th percentile of the absorbed cycle dosages ranged from 100 (applicators handling 70WP) to 2500 for spray applicators using 0.95% formulation. The MOS for the 95th percentile of the absorbed cycle dosage for dog owner/groomers was 13. The acute NOEL for propoxur was based on human oral exposure leading to cholinergic signs. Even though a single bolus dose of 0.36 mg/kg produced short-lasting stomach discomfort, blurred vision, moderate facial redness, and sweating, five oral doses of 0.2 mg/kg at 30-min intervals over a period of 2 1/2 h did not cause cholinergic signs. This indicated that carbamylation of cholinesterase, caused by bolus oral doses of propoxur, was rapidly reversed in the human body (Ellenhom et aI., 1997). However, the preponderance of occupational or nonoccupational acute exposure to propoxur was through the dermal route (approximately 99% in most instances). As absorption of propoxur via the dermal route is generally slower than absorption from the gut, decarbamylation, body metabolism, and clearance probably limit the effects of acute dermal exposure to propoxur. Consequently, the margins of safety under actual exposure conditions are probably greater than indicated.
32.4.4 DIQUAT DIBROMIDE
Diquat dibromide (6,7 -dihydrodipyrido-[ 1,2-a :2', l' -c]pyrazinediium ion) is a contact herbicide that damages plant tissues quickly, causing plants to appear frostbitten because of cell membrane destruction. It also reduces plant photosynthetic activity. This nonselective contact herbicide is used for desiccation of potato vines and seed crops, control of sugarcane flowering; and industrial and aquatic weed control. As in the preceding examples, most of the exposures to diquat dibromide involve dermal absorption (Cochran et aI.,
32.4 Examples
1994). Consequently, a short-term dermal NOEL would have been desirable as the critical NOEL for assessing the risks of acute human exposure to diquat. Unfortunately, no such singledose dermal studies were available in the CDPR database or from a search of the open literature. A subchronic dermal exposure study on the effects of diquat dibromide on rats indicated systemic effects (death) began after 6 days of repetitive dosing. In rabbits, ulceration of the gastric mucosa, degeneration of the convoluted tubules in the kidneys, areas of hemorrhage in the thymus, and congestion of the lungs and lung blood vessels accompanied by death were observed at 12.5 mg/kg-day after 3 days of repetitive dermal dosing with diquat dichloride. However, deficiencies in the dosing regime, and the difference in the chemical identity of the test material, precluded the use of this study as the basis for regulating short-term exposure to diquat dibromide. The toxicological basis for assessing the risks associated with potential short-term exposure to diquat, therefore, was identified in oral dosing studies. A single oral dose of diquat dibromide at 75 mg/kg caused clinical signs (diarrhea and stained nose) in female rats. The NOEL for clinical signs was 25 mg/kg. In a rat developmental study, the developmental NOEL (delayed ossification) was 12 mg/kg-day, whereas the maternal NOEL (decrement in body weight gain) was 4 mg/kg-day for exposure to diquat by gavage. Mice appeared to be more sensitive to diquat than were rats. The mouse NOELs for maternal toxicity (clinical signs, death) and developmental toxicity (skeletal anomalies, exencephaly, and umbilical hernia) were both 1.0 mg/kg-day. The rabbit appeared to be the most sensitive laboratory animal to diquat in developmental toxicity studies. The maternal NOEL was 3.0 mg/kg-day (histopathological changes in the liver, intestine, and vasculature; mortality), but there was no developmental NOEL. The ossification of the ventral tubercle of the cervical vertebrae was delayed significantly in all treatment groups compared to controls. In addition, the incidence of fetal malformations was significantly greater in the low-dose (1 mg/kg-day) and high-dose (10 mg/kg-day) groups compared to the controls. The incidence at the mid-dose (more than a twofold increase over controls) lacked statistical significance, but may have represented a biologically significant finding, supportive of a treatment-related effect. This hypothesis was consistent with the suggested common mechanism (interference with cell migration) for the observed anomalies across the different treatment groups. Although the fetal malformations in the low-dose group (1 mg/kg-day) and the high dose group (10 mg/kg-day) were quantitatively different from the controls, the malformations were not qualitatively different from either the concomitant or the historical controls. Nonetheless, the possibility that diquat caused a significant (p < 0.05) increase in the number of malformations found at the low dose (1 mg/kgday) could not be ignored. As there was no NOEL, an estimated no effect level (ENEL) was calculated. Because the magnitude (incidence) of the effect was small, and the slope of the dose response was fairly shallow, an uncertainty factor of 3 was used to derive an ENEL of 0.33 mg/kg-day. As absorption of diquat
699
across the gut in a rat pharmacokinetic study was approximately 10%, the ENEL was divided by a factor of 10 to reflect the absorbed dosage that would be expected to have no toxic effect. This adjusted ENEL, 0.033 mg/kg-day, based on the observations of delayed ossification and fetal malformations, was used to calculate margins of safety for potential short-term exposure to diquat. It was assumed that the developmental toxicity observed in pregnant rabbits could occur as the result of a single dose. Further, it was assumed that absorption of oral dosages by the rabbit and human would be limited to the same degree as in the rat. The data that formed the basis for estimating worker exposure were based on monitoring studies for diquat and calculations from studies involving a surrogate active ingredient (paraquat) with similar application rates and chemical properties. Dermal absorption constituted the principal route of exposure. Monitoring data for both aquatic use and ground spraying indicated that less than 1% of the total exposure came through the inhalation route. Exposure data from aerial application of the surrogate herbicide, paraquat, indicated that pilots and flaggers could be exposed as much through the inhalation route as through the dermal route. The exposure estimates used for the risk assessment are shown in Table 32.4. Potential short-term exposures ranged from 0.2 !-lg/kg-day for mixers and applicators injecting diquat into aquatic environments to 106 !-lg/kg-day for ground applicators driving tractors with no cabs and a normal ground clearance. The potential short-term exposure to drift was 0.5 !-lg/kg-day at 50 and 0.01 !-lg/kg-dayat 1600 m. Potential short-term exposures (4 h) for adult males swimming in treated water 24 h after application ranged from 0.2 to 1.3 !-lg/kg-day. Only mean exposure values were available; potential upperbound exposures would likely have been greater. MOSs for short-term occupational exposures, based on the adjusted ENEL of 33 !-lg/kg for developmental toxicity and maternal clinical signs, ranged from less than 1 (ground applicators) to 165 (aquatic mixers and injectors). The MOSs for potential short-term exposure of swimmers to diquat dibromide ranged from 26 (theoretical water concentration) to 165 (measured water concentration). If the developmental ENEL of 0.033 mg/kg-day had not been used as the basis for calculating the MOSs for acute exposure, the next best oral NOEL was 1 mg/kg-day for clinical signs and death in the mouse (LOEL = 2 mg/kg-day) from a developmental study. This NOEL would also have been adjusted to 0.1 mg/kg-day because of the 10% oral absorption. Using the adjusted NOEL of 0.1 mg/kg-day for clinical signs and death in the mouse, the MOSs for mean acute occupational exposures would have ranged from less than 1 to 500, and the MOSs for mean acute nonoccupational exposures would have ranged from 77 to 1000. Consequently, the conclusions would not have changed. Margins of safety for some exposures would have remained less than 100. A mitigation process was initiated to ascertain whether changing the manner in which diquat was applied would reduce exposure. Additional exposure studies and a new acute dermal
700
CHAPTER 32
Risk Assessment for Acute Exposure to Pesticides
Table 32.4 Potential Mean Absorbed Daily Dosages and Margins of Safety from Exposures to Diquat Dibromide
Activity
ADD
Short-term
(j.lg/kg-day)
Mosa
Aquatic Mixer (injection)
0.2
165
Applicator (injection)
0.2
165
Applicator (handgun)
3.6
9
Boat driver (handgun)
0.9
37
Aerial Mixerlloader
7.8
4
Pilot
0.3
III
FJagger
8.1
4
Ground application Applicator-normal clearance, cab
106
<1
Applicator-normal clearance, no cab
7.4
5
Applicator-high clearance, no cab
5.3
6
Applicator-hand sprayer (right-of-way)
0.35
95
Gardenerllandscaper (ready-to-use formulation)
0.4
83
11.6
3
Gardenerllandscaper (knapsack sprayer) Non-occupational Aerial drift (50 m)
0.5
67
Aerial drift (1600 m)
0.01
330
Swimmer
0.2-1.3
26-165
aBased on an adjusted ENEL of 33 j.lg/kg-day for developmental toxicity in rabbits.
toxicity study were submitted by the registrant during the mitigation process. Estimates of occupational and nonoccupational exposures from the original exposure assessment are contained in Table 32.5. The dermal dose was the amount of diquat estimated to be deposited on the skin daily in each of the work tasks. This represented 99% of the exposure to diquat. However, the exposure estimates for a mixerlloader/applicator were updated because a new surrogate biological monitoring study of mixerlloader/applicators using paraquat was submitted. As paraquat is chemically similar to diquat, the former had been used as the surrogate for diquat in all exposure categories where specific data for diquat were not available. The new study indicated that the amount of diquat absorbed by certain workers was 0.3 !-1g/kg-day. This value was over 300 times less than the previous estimate of absorbed dosage (103 !-1g/kg-day) for ground applicators, which was based on passive dosimetry and a dermal absorption factor of 1.4% from an earlier worker exposure study involving paraquat. In a new dermal toxicity study, New Zealand white rabbits (5/sex/dose) were given a single, 6-h dermal dose of diquat dibromide (20.5% diquat ion) at 0, 50, 100, or 200 mg/kg, and then were assessed for 14 days. Blood levels of diquat were determined 6 h after dosing and indicated that diquat was passing through the skin in a generally dose-dependent manner. No clinical signs were noted in any of the animals, though one female at the high dose (200 mg/kg) was terminated in ex-
tremis on day 11. No treatment-related lesions were noted in the terminated animal, and no signs of systemic toxicity were noted in the survivors. There was significant (p < 0.01) reduction in food consumption (33-71 %) at the high dose for the first 5 days after treatment. Slight to moderate skin irritation was noted with all doses of diquat. The single-dose NOEL for death (1/10) and reduced food consumption in the rabbit was 100mg/kg. Acute MOSs were calculated as the ratio of the dermal NOEL (100 mg/kg) divided by the dermal exposure dose, rather than the estimate of the amount absorbed though the skin (Table 32.5). Only one of the work tasks (ground applicator on a tractor with normal clearance and no cab) indicated an MOS less than 100 (Table 32.5). However, the new exposure study generated a new estimate of absorbed dose (0.3 !-1g1kg-day) for this work category. To obtain the MOS for this work category, the adjusted oral ENEL (33 !-1g/kg-day for developmental toxicity in rabbits), which represents the no effect level for an absorbed dose though the oral route, was divided by the estimated absorbed dose of applicators (0.3 !-1g/kg-day) to obtain the MOS (110).
32.4.5 MOLlNATE Molinate (S-ethyl hexahydro-1H-azepine-l-carbothioate) is a selective, preemergence thiocarbamate herbicide registered for
32.4 Examples
701
Table 32.5 Potential Daily Dermal Doses and Margins of Exposure for Occupational and Nonoccupational Exposures to Diquat Dibromide Applied dermal dosea Work task
( ~gIkg-day)
MOEb
Aquatic Mixer (injection)
14.2
7,042
Applicator (injection)"
14.2
7,042
Applicator (handgun) Boat driver (handgun)
257 64.3
389 1,555
Aerial MixerlIoader Pilot Fiagger
557 21.4 579
180 4,673 173
Ground application Applicator-normal clearance, no cab
7,57F
BC
Applicator-normal clearance, cab
529
Applicator-high clearance, no cab
379
264
25
4,000
28.6
3,497
Applicator-hand sprayer (right-of-way) GardenerlIandscaper (ready-to-use formulation) GardenerlIandscaper (knapsack sprayer)
829
189
121
Nonoccupational Aerial drift (50 m) Aerial drift (1,600 m) Adult, swimming
35.7 0.71 14.2-92.9
2,801 140,000 1,076-7,042
aDermal exposure derived from the exposure assessment. bBased on an NOEL of 100 mg/kg for reduced food consumption and death (1/10) in rabbits at 200 mg/kg. C See text for further explanation.
use on rice. As in all of the previous examples, early exposure studies indicated most of the occupational exposure was through the skin. However, in the case of molinate, the absorbed dose in occupational exposures used for risk assessment were determined by biological monitoring of urinary metabolites. The toxicological studies that most thoroughly gauged the absorbed dose response level were administered through the oral route (Cochran et at., 1997). A small number of studies examined the toxic effects of short-term exposure to molinate. These effects occurred in developmental and reproductive toxicity studies after repetitive daily dosing. In rabbits, the 13-day NOEL for maternal toxicity (decrement in weight gain) and developmental toxicity (increased resorptions, intrauterine growth retardation) was 20 mglkg-day. In rats, the 9-day NOEL for maternal toxicity (cholinergic signs, decrement in food consumption, and weight gain) and developmental toxicity (increased resorptions, intrauterine growth retardation) was 35 mglkg-day. The 5-day NOEL for reduced fertility was 11.5 mglkg-day in male rats. A significant (p < 0.01) decrement in maternal body weight gain was seen at 4 days in rats receiving oral doses of 150, 75, or 15 mglkg-day (there was no NOEL) in a range-finding developmental toxicity study. The toxicological significance of the decrement in body weight gain, though, is unclear. Clinical signs (salivation, subdued appearance, hunched posture, piloerection, ocular discharge, stained nose and mouth, urinary in-
continence, irregular breathing) were noted in rats at 3 days after being dosed with 150 mg molinatelkg-day. The 7-day NOEL for clinical signs (salivation) in the rat in the same study was 15 mglkg-day. Rats given 350, 100, or 25 mglkg of molinate in a single-dose neurotoxicity study exhibited clinical signs (including upward curvature of the spine; hypersensitivity) and a significant dose-related decline in performance on functional observational battery tests at 4 h after molinate was administered. There was no NOEL in the study. The critical NOEL used to calculate a margin of safety for potential acute exposure to molinate was 11.5 mglkg-day for reduced fertility in rats. Although the effect was noted after five daily doses, this was the lowest short-term NOEL. In addition, this NOEL addressed one of the two toxicological endpoints of concern (reproduction and neurotoxicity). No single-dose studies that examined this effect were available. As oral absorption of molinate in a pharmacokinetic study was effectively 100%, the NOELs for oral administration represent the NOEL for an absorbed dose. The absorbed dosage was estimated from measured levels of 4-hydroxymolinate in the urine of workers handling 1200-lb bags during applications in California. The measured absorbed dosage included any potential dietary, drinking water, or other nonoccupational exposures as well. It was assumed that the absorbed dosage was directly proportional to the amount of active ingredient handled. To calculate the max-
702
CHAPTER 32
Risk Assessment for Acute Exposure to Pesticides
Table 32.6 Absorbed Daily Dosages and Margins of Safety for Short-Term Exposure to Molinate for Workers, Farmers, and Residents ADDa
Work task/protective clothing
(~g/kg-day)
AcuteMOS
Driver (no suit)
0.76
15,000
Driver (carbon suit)
0.56
21,000
Direct loader (Tyvek suit)
10.58
1,000
Direct loader (carbon suit)
6.89
2,000
Direct loader/transfer loader (Tyvek suit)
3.70
3,000
Direct loader/transfer loader (carbon suit)
4.85
2,000
FJagger
1.1
10,000
Pilot
3.5
3,000
Farmer
0.12
96,000
MaxweII, adult
0.49
23,000
MaxweII, infant
1.02
11,000
aBased on biomonitoring data (urinary metabolites).
imum permissible exposure to molinate, the measured geometric mean of the absorbed daily dosage was multiplied by the proportion of molinate handled, as well as the range of absorbed dosages (Table 32.6). Farmers entering the rice field shortly after molinate application also had potential exposure to the airborne herbicide. The potential exposure duration for farmers was estimated to be 1 h per day. Measurement at 48 cm above the water surface of the rice field showed an air concentration of 48 J.l.g/m 3 immediately after the application of molinate in 1985 (Ross and Sava, 1986). Under the study conditions, the concentration decreased to 8.3 J.l.g/m3 3 days after the treatment. The arithmetic average concentration was 21.95 J.l.g/m 3 within 3 days after molinate application. The potential arithmetic average exposure dosage for farmers entering the rice field within 3 days after molinate application is shown in Table 32.6. Nonoccupational exposures were principally through the inhalation route. In 1986, the ambient air concentrations of molinate were measured on the rooftops of the public buildings in four Sacramento Valley towns (Seiber et aI., 1989). Sampling was carried out for four 24-h intervals (Monday AM through Friday AM) for 4 weeks during the period selected to represent the highest uses of molinate. The maximum concentration was detected at the town of Maxwell at 1.7 J.l.g/m3. Absorbed dosage was calculated using the standard default value for human inhalation of 0.29 m3/kg-day for adults and 0.6 m3 /kg-day for infants (Anderson et aI., 1983) and assuming a 50% retention and 100% absorption (Raabe, 1986, 1988). The estimated acute dosages for adults and infants in Maxwell in 1986 are shown in Table 32.6. Margins of safety for the geometric mean potential exposures of agricultural workers and local residents to molinate ranged from 1000 to 96,000 for short-term exposures (Table 32.6). If the 95th percentile of short-term exposure were considered, the MOSs would range from 214 for direct loaders (wearing Tyvek) to 6765 for fiaggers.
32.5 ANALYSIS AND CONCLUSIONS The toxicological database for pesticides as a chemical class is the most complete for any category of chemicals, with the exception of pharmaceuticals. The FIFRA- and EU-required toxicological studies are designed to produce data on the types of toxic effects a chemical might possibly cause, as well as the dose that is required to elicit that effect. Nonetheless, these studies often do not contain the type of information needed for accurate estimates of the acute risks associated with the use of a pesticide. Unfortunately, only one of the currently required studies examines the amount of chemical needed to produce nonlethal systemic effects from a range of single doses. Single-dose neurotoxicity studies, though, are not yet being required for nonneurotoxic chemicals. Consequently, the critical acute NOELs for such chemicals are derived from multiple-dose studies, such as developmental toxicity. In those instances where the assessor is forced to use NOELs for systemic effects caused by multiple doses, as in the case of molinate discussed previously, there could be an overstatement of the risks from a single dose. As indicated in the preceding examples, most, but not all, of human exposure to pesticides occurs through the skin. Yet, almost all of the required toxicity studies utilize the oral exposure route for dosing. Clearly, the respective types of data do not fit together well. Dermal pharmacokinetic studies, which provide data on plasma levels and the half-lives of pesticides and their metabolites, could be used to better understand the differences in toxicokinetics for chemicals entering the body through the skin rather than via the digestive system. At the present time, however, such studies are not required.
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Crop Protection Association by Versar Inc. (6850 Versar Center, P.O. Box 1549, Springfield, VA 22151). Raabe, O. (1986). "Inhalation of Selected Chemical Vapors at Trace Levels." CARB Contract No. A3-132-33, California Air Resources Board, Sacramento. Raabe, O. (1988). "Retention and Metabolism of Toxics: Inhalation Vptake of Xenobiotic Vapors by People." CARB Contract No. A5-155-33, California Air Resources Board, Sacramento. Rider, J. A., Puletti, E. J., and Swader, J. I. (1975). The minimal oral toxicity level for mevinphos in man. Toxicol. Appl. Pharmacol. 32, 97-100. Ross, L. J., and Sava, R. J. (1986). Fate of thiobencarb and molinate in rice fields. J. Environ. Qual. 15,220--225. Ross, J. H., Fong, H. R., Thongsinthusak, T., and Krieger, R. I. (1992). Experimental method to estimate indoor pesticide exposure to children. In "Similarities and Differences between Children and Adults: Implications for Risk Assessment" (P. S. Guzelian, C. J. Henry, and S. S. Olin, eds.), pp. 226--241. ILSI Press, Washington, DC. Schardein, J. L. (1985). "Chemically Induced Birth Defects," p. 10. Dekker, New York. Seiber, J. N., McChesney, M. M., and Woodrow, J. E. (1989). Airborne residues resulting from use of methyl parathion, molinate, and thiobencarb on rice in the Sacramento Valley, California. Environ. Toxicol. Chem. 8,577-588. Stanton, M. E., Mundy, W. R., Ward, T., DuIchinos, v., and Barry, C. C. (1994). Time-dependent effects of acute chlorpyrifos administration on spatial delayed alternation and cholinergic neurochemistry in weanling rats. Neurotoxicology 15, 201-208. Thongsinthusak, T., Ross, J. H., and Meinders, D. (1993). "Guidance for the Preparation of Human Pesticide Exposure Assessment Documents." HS1612, Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento. Trichopoulou, A. (1994). Household budget surveys and international comparability of food availability: The DAFNE Report. In: "Nutrition in Europe: Nutrition Policy and Public Health in the European Community and Models for European Eating Habits on the Threshold of the 21st Century." Report of an ad hoc group. Tyl, R. w., Ballantyne, B., Fisher, L. c., Fair, D. L., Dood, D. E., Klonne, D. R., Pritts, I. M., and Losco, P. E. (1995). Evaluation of the developmental toxicity of ethylene glycol in CD-l mice by nose-only exposure. Fundam. Appl. Toxicol. 27,49-62. V.S. Department of Agriculture (VSDA) (1989-1991). "Food and Nutrient Intake by Individuals in the Vnited States, 1 Day, 1989-1992. Continuing Survey of Food Intakes by Individuals, 1989-1992." Agricultural Research Service, US. Department of Agriculture, Washington, DC. V.S. Department of Agriculture (VSDA) (1996). "Pesticide Data Program (PDP) Annual Summary Calendar Year 1994" (w. I. Franks and R. L. Epstein, eds.). Agricultural Marketing Service, V.S. Department of Agriculture, Washington, DC. V.S. Environmental Protection Agency (EPA) (1984). "Pesticide Assessment Guidelines, Subdivision F. Hazard Evaluation: Human and Domestic Animals." US. Environmental Protection Agency, Washington, DC. V.S. Environmental Protection Agency (EPA) (1986). "Human Variability in Susceptibility to Toxic Chemicals-Noncarcinogens." EPA 600/8-86-033, NTIS PB87-101242/AS, V.S. Environmental Protection Agency, Washington,DC. V.S. Environmental Protection Agency (EPA) (1988a). "Guidance for the Reregistration of Pesticide Products Containing Ethoprop as the Active Ingredient." Office of Pesticides and Toxic Substances, V.S. Environmental Protection Agency, Washington, DC. V.S. Environmental Protection Agency (EPA) (1988b). "Cholinesterase Inhibition as an Indication of Adverse Toxicologic Effect. Forum Review Draft, June 1988." V.S. Environmental Protection Agency, Washington, DC. V.S. Environmental Protection Agency (EPA) (1990). "Report of the SAB/SAP Joint Study Group on Cholinesterase: Review of Cholinesterase Inhibition and Its Effects." EPA-SAB-EC-90-014, V.S. Environmental Protection Agency, Washington, DC. US. Environmental Protection Agency (EPA) (1991). "Guidelines for Developmental Toxicity Risk Assessment." CFR 56 (No. 234): 63798--63826.
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CHAPTER
33 Risk Assessment for Chronic Exposure to Pesticides: The Triazine Herbicide Cyanazine* Derek w. Gammon and Keith F. Pfeifer Department of Pesticide Regulation, California Environmental Protection Agency
33.1 SUMMARY The human health risks from exposure to pesticides are evaluated in the United States in two ways: a value of no-observedeffect level (NOEL) for a (noncancer) long-term systemic toxic effect is divided by the estimated exposure, occupational and/or dietary, yielding a margin of safety (MOS). This numerical value is generally considered to be protective of human health if it exceeds a particular value (usually 10 if the NOEL value is based on a human study or 100 for an animal study). Cancer is regulated separately from other chronic effects. Trianize herbicides are the most heavily used pesticides in world agriculture. In 1994, a Special Review was initiated by the U.S. Environmental Protection Agency (U.S. EPA). Concerns were expressed about possible cancer risks resulting from the use of three triazines: atrazine, cyanazine, and simazine. A chronic risk assessment for cyanazine is described in detail to provide an example of the process for regulating carcinogens, in practice. The toxicology database is summarized and estimates of occupational and dietary exposure are provided. The calculations of margin-of-safety (MOS) values and cancer risk are discussed in terms of uncertainties in the data and also in terms of alternative methods of assessing the data for the determination of cancer risk arising from pesticide use.
33.2 INTRODUCTION Recent pesticide legislation in the United States has concentrated on possible acute, short-term effects of exposure and has focussed on population groups considered to be at greater potential risk [e.g., Food Quality Protection Act (FQPA, 1996)]. However, long-term risks such as cancer, which have been the *The opinions expressed in this paper are the authors' and do not necessarily reflect the views and policies of the Department of Pesticide Regulation. Handbook of Pesticide Toxicology Volume 1. Principles
subject of most of the earlier legislation [e.g., Federal Insecticide, Fungicide & Rodenticide Act (FIFRA) (U.S. EPA, 1984)], remain of great concern, to both regulators and the general public. One of the main fears, real or imagined, is that people's food, water, and air are contaminated with dangerous levels of pesticide residues (Wiles et aI., 1994). By conducting risk assessments, these fears can usually be dispelled, or else the pesticide more stringently regulated, with withdrawal from the market as the last resort. One class of pesticides which has been considered to present a cause for concern is the triazine herbicides. This class includes 3 of the 10 most heavily used pesticides in the United States, atrazine, cyanazine, and simazine. This chapter describes efforts, conducted on behalf of the California Environmental Protection Agency, Department of Pesticide Regulation (DPR), to assess the potential risks associated with the use of cyanazine in California. It is intended to serve as an example of how regulations, developed to protect people and the environment from long-term effects of pesticides, are applied in practice. As described in the chapter by Cochran (Cochran, 2001) these risks are estimated by first reviewing the studies defining the toxicity of the chemical, usually conducted on laboratory animals, and determining the NOELs. Second, possible human exposure is estimated, both in agricultural occupations (mixing, loading, applying), in the diet (food and drinking water) and combined. Uncertainties in the estimation of long-term exposure are also described. Having previously quantified the toxicological end points, for acute, subchronic, and chronic exposure, it is then possible to estimate the likely margin of safety of the chemical under typical conditions of use. Although this chapter concentrates primarily on chronic toxicity and cancer, acute toxicity is summarized for comparison. Shortterm exposure is also included because it usually provides the basis for the calculations of long-term exposure, both occupational and dietary. Subchronic toxicity is used in the event of
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Risk Assessment for Chronic Exposure
subsequently by E.!. DuPont de Nemours & Co., under the trade name Bladex (V.S. EPA, 1986b). On January 3, 1985, V.S. EPA issued a registration standard for cyanazine where data gaps were identified and registrants were required to develop the additional data within a specified time frame (V.S. EPA, 1986a). Besides certain data describing product chemistry, residues, and environmental chemistry, toxicological data were required. These included chronic toxicity and oncogenicity studies in two species, developmental toxicity studies in two species, a two-generation reproduction study, a dermal absorption study, and a complete set of genotoxicity testing. A Special Review of cyanazine was initiated by the V.S. EPA in 1985 based on its teratogenic effects in rats, fetotoxicity in rabbits, and "sufficient exposure to mixerlloaders and applicators" (V.S. EPA, 1985). The teratogenic effects were reported in Fischer 344 rats where increased incidences of anophthalmia (no eyes) and microphthalmia (small eyes) were observed (lyer et a!., 1999). At the conclusion of the special review (U.S. EPA, 1987), extensive label modifications were required, including a "Restricted Vse Classification." In 1990, California listed cyanazine as a chemical which was "known to the State to cause reproductive toxicity" under the State Drinking Water and Toxic Enforcement Act of 1986 (Proposition 65) (see California, 1986). A reference dose (RID) of 0.002 mg/kg/day, a drinking water equivalent level (DWEL) of 0.07 mgll (both based on a NOEL of 0.2 mg/kg/day for body weight loss in a 2-year rat feeding study), and a Qj of 1.0 (mg/kg/day-l), based on increased mammary gland tumors in female rats, have been established (V.S. EPA, 1993). 33.2.1 USES OF CYANAZINE A lifetime V.S. EPA Health Advisory (for noncancer toxicity) Cyanazine, (2-[[4-chloro-6-(ethylamino)-s-triazin-2-yl]amino]- of 1 ppb in drinking water is in effect. The maximum con2-methylpropionitrile), is a selective pre- or post-emergence tamination level (MCL) and allowable daily intake (ADI) are triazine herbicide registered for use to control annual grasses pending from V.S. EPA. More recently V.S. EPA initiated another Special Review and broadleaf weeds in corn, cotton, grain sorghum, winter wheat, and fallow crop land (V.S. EPA, 1986a). It is a photo- of triazines (V.S. EPA, 1994a). In this case, concerns were synthesis inhibitor, which inhibits the Hill reaction of photo- expressed about possible cancer risks, resulting not just from system 11, causing chlorosis, necrosis, and plant death (Corbett cyanazine, but also from two other related triazine herbicides, atrazine and simazine, either alone or in combination. Subseeta!., 1984). quently, the current cyanazine registrants, DuPont and Griffin, Cyanazine may be applied alone or in combination with have voluntarily agreed to gradually phase out and eventually other herbicides and fertilizers. The application rates vary and of cyanazine, under certain conditions, by the cancel the use are dependent on the soil texture and its organic matter conyear 2002. tent. Higher rates are used on heavier soils and soils with higher organic matter content. In 1992-1993, cyanazine was the fifth most heavily used pesticide in the Vnited States, with over 30 33.2.3 ENVIRONMENTAL FATE million pounds being applied annually, mostly on corn. Its use in California has accounted for only about 1% of this total. In Cyanazine is rapidly degraded in the presence of both soil and contrast to the national use pattern, cyanazine in California is sunlight to derivatives which retain the triazine ring. Hydrolmostly used on cotton (DPR, 1993). Other California uses inysis of chlorine and nitrile groups occurred under acid concluded corn, wheat, and fallow cropland. ditions (pH 5) or on soil in the dark. An additional reaction, N -dealky lation to the desethy 1 derivatives, occurred through photolysis. Anaerobic soil degradation was similar to aerobic 33.2.2 REGULATORY HISTORY but occurred more slowly. Cyanazine has medium to high soil Cyanazine was registered in 1971 by V.S. Environmental Pro- mobility, indicating a high leaching potential. There is little tentection Agency to be sold by Shell Chemical Company, and dency for residues to accumulate in crops.
a lack of adequate chronic toxicity studies and for addressing seasonal exposure. Subchronic studies are also employed to aid in dose selection for chronic studies. The top dose in such studies [also called the maximum tolerated dose (MTD)] is generally defined as one which causes a 10% fall in body weight, without any other evident effects (Eaton and Klaassen, 1996). For a carcinogen, the current regulations in the Vnited States require that a cancer potency factor be determined, wherever possible, assuming a linearized no-threshold (LNT) doseresponse. This factor is multiplied by the estimated exposure to give the excess lifetime cancer risk. The chronic (annual) and lifetime exposure estimates, which are based on the amortization of acute exposure(s), may greatly overestimate "true" long-term exposure(s), and uncertainties in this area are discussed. There are presently carcinogens for which the LNT model is not used, such as where experimental evidence suggests a nongenotoxic mechanism (of little relevance to humans) or when data indicate an effect only at the highest dose tested. In this chapter, MOSs and potential risk are presented along with a discussion of possible reasons for considering alternative method(s) for cancer risk assessment. As regulators, it is considered necessary to make "worstcase scenario" assumptions to protect human health. However, additional information on mechanisms of toxicity, particularly with respect to cancer, is always considered in arriving at appropriate regulatory decisions on how risk assessment should be conducted.
33.3 Toxicology Profile
709
Table 33.1 Cyanazine Residues in Crops (l9Sl-l985)a Preharvest
Residues b
Crop
interval (days)
(ppm) NDd
Rate Pormulation
(lb ai/A)
90DP, 4L, SOW
4-l2C
Pield or sweet corn
77-1S4
90DP,4L
4.0c
Cottonseed
92
ND
90DP, 4L, SOW
l.6-4e
Wheat grain
33-253
ND
90DPi
3.ge
Sorghum
129
ND
90DPi
2.0-4.0e
Pield corn
ll6-13S
ND
90DPi
5.6<
Cottonseed
99
ND
90DPi
4.0e
Wheat grain
343
ND
aProm studies summarized in Gammon et al. (1997). bLabel-approved preharvest interval (PHI) for cotton is 54 days. cMultiple applications (preplanting, pre- and post-emergence). dND is not detected; LOD for cyanazine is 0.01 ppm, for metabolites SD 33104 and SD 20196 is 0.03 ppm, and for SD 31223 and 31224 is 0.05 ppm. eSingle application (preplanting or preemergence).
f Tank mix with other herbicides.
33.2.4 RESIDUES Results of field tests indicate that residues of cyanazine in crops grown in soil treated with cyanazine are nondetectable (limit of detection, 0.01 ppm). Field data (1981-1985) are summarized in Table 33.1.
(male) and 369 (female) mg/kg. By inhalation, cyanazine dust had LCso > 906 mg/m3 (LDso > 152 mg/kg) after a 4-hr exposure, with an estimated NOEL of 1.6 mg/kg. Data describing the acute toxicity of metabolites are limited; two major metabolites had oral LDso values:::: 4-fold that ofthe parent.
33.3.3 SUBCHRONIC TOXICITY
33.3 TOXICOLOGY PROFILE 33.3.1 PHARMACOKINETICS Excretion in the rat following oral dosing with 14C cyanazine was approximately 34% in the urine and 18% in the feces, within the first 24 hours. Assuming that the 14C in the feces was not absorbed, 82% of the dose was absorbed from the gut. Elimination of radiolabel was fairly rapid and was nearly complete within 4 days. In the dog, 52-64% was absorbed by the oral route. Based on several rat studies, cyanazine rapidly undergoes metabolism via N -deethylation, dechlorination, and conjugation with glutathione with subsequent formation of mercapturic acids, and oxidation of the nitrile group. Dermal absorption of cyanazine in the male rat, using a solution of Bladex®4L, averaged 0.9% over 10 hr and peaked at 2.0% (group mean) and 4.6% (highest individual value) at 24 hr. Dermal absorption in the female rabbit was similar: maximally, 1-3% occurred after a 6-hr exposure period. 33.3.2 ACUTE TOXICITY Cyanazine and its formulations were not acutely toxic to the rabbit by dermal exposure, at :::: 2000 mg/kg. There was only mild dermal and eye irritation resulting from cyanazine dosing in the rabbit and no dermal sensitization in the guinea pig. Acute oral toxicity studies in the rat gave LDso values of 835
The major, consistent dose-related effect of cyanazine was loss of body weight and body weight gain. In the dietary studies there was a concomitant reduction in food intake in the rat, but not in the mouse. In dermal (rabbit) and inhalation (rat) studies, the body weight reduction did not appear to be consistently accompanied by a loss of appetite.
33.3.4 CHRONIC TOXICITY AND ONCOGENICITY Chronic toxicity manifested itself as severe weight loss in all species tested (rat, mouse, dog), usually in conjunction with reduced food intake. Many of the chronic effects, for example, chronic inanition, poor skin and fur condition, and anemia, could thus have been a consequence of inadequate nutrition. In an early rat chronic toxicity study, an increased incidence of thyroid adenomas was reported in males at the highest dose, but without showing a clear dose-response relationship. In a later rat study, reduced body weight gain was reported for males. Other effects noted include increased alveolar macrophages, reduced creatinine kinase and atrophy of the seminiferous tubules. Cyanazine resulted in an increase in malignant mammary gland tumors in females. The incidence of adenocarcinomas, considered with carcinosarcomas, was elevated significantly at the three highest doses.
710
CHAPTER 33
Risk Assessment for Chronic Exposure
33.3.4.1 Dietary Exposure-Rat
the highest dose tested (HDT), as follows: the number surviving to Day 721 was increased (p < 0.02, Fisher's exact test); for females, the increased survival was not significant (p = 0.08). The increased incidence of malignant mammary tumors in females (see below) may have compromised the increased lifespan which would have been anticipated from reduced food intake. For example, females at 50 ppm, but not at lower doses, had a shorter mean lifespan when malignant mammary tumors were present (540 ± 106 days, n = 6) than when they were not present (617 ± 107 days, n = 22), excluding animals killed by study design. Other chronic effects of dosing included an increased incidence of hyperreactivity in males at 25 ppm in 24 out of 280 observations (p < 0.05, Fisher's exact test) and 50 ppm in 34 out of 329 observations (p < 0.01), from 280 days onwards, with a positive dose-response (p < 0.01, Peto's trend test). However, the occurrence of instances of hyperreactivity in untreated rats (12/259) makes this sign of doubtful toxicological relevance. Furthermore, the dose-response relationship was discontinuous: although 1 ppm cyanazine caused a significant (p < 0.05) increase in hyperreactivity (171174), 5 ppm did not (171273). In males, increased foamy alveolar macrophages were reported at 2 years (p < 0.05, Peto's trend test), without being significantly elevated at any particular dose. Significant effects on organ weights were as follows: decreased mean absolute kidney weight (16%, p < 0.05, Dunnett's test) and increased mean relative testis weight (34%, p < 0.05, Dunnett's test) at two years, in males at 50 ppm, without histopathological changes. Creatinine kinase, a marker enzyme for energy production in muscle, was significantly reduced (p < 0.05) at
In two studies using Carworth Farm E strain rats, cyanazine was administered in the diet for 2 years (Walker et aI., 1974). In one, there was an increased incidence of thyroid C-cell tumors, particularly adenomas, in top dose males compared with concurrent controls. This was not considered to represent a treatment effect. A reevaluation of thyroid C-cell tumor incidence was undertaken, considering this study, plus the second study (discussed next), and the acceptable "combined" study (Bogdanffy, 1990). The reevaluation likewise concluded that there was not a treatment effect on the incidence of thyroid tumors. In the second study the only noticeable effect was lower mean body weight (up to 10%) at the highest dose compared to controls early in the study (Table 33.2). Potential chronic and oncogenic effects in rats were comprehensively evaluated in a more recent study (Bogdanffy, 1990). Cyanazine was given in the diet for 24 months at 0, 1, 5, 25, or 50 ppm to rats (Crl:CD@BR, 62/sex/group). Chronic toxicity included reduced body weight and body weight gain in both sexes at 25 and 50 ppm, accompanied by slight decreases in mean daily food intake of 4 and 9%, respectively. Mean body weight gain of male and female rats was reduced at 50 ppm by 20% (p < 0.05, Dunnett's test) and 16% (p < 0.05), respectively, at 1 year. At 2 years, the corresponding reductions in body weight gain were 19 and 17%. At 25 ppm, body weight gain of males and females was reduced by 7% (p < 0.05) and 13% (p < 0.05), respectively, at 1 year. By 2 years, the corresponding reductions were 7 and 4%. With cyanazine administration, longevity was increased significantly only for males, at
Table 33.2 Summary of Subchronic and Chronic Effects Caused by Cyanazine LOEL Species
Route
NOEL
(mglkg/day)
Effect
Refa
Subchronic Toxicity Decreased body weight
Mouse, 13-wk.
Diet
13-wk.
Diet
Clinical chemistry changes Decreased body wt., food consumption
Rat,
2
5
Chronic Toxicity and Oncogenicity Dog,
0.625
3.0
0.70
1.5 b
O.IY
Decreased body wt.
1.25 b
0.15 b
2
Males: hyperreactivity, decreased body wt.
1.0
0.2
3
Reductions in body wt. gain, absolute
Capsule
Liver wt. and total serum protein
I-yr.
Diet
Increased relative organ wts.
Dog,
Decreased body wt.
Mouse,
Decreased body wt.; renal cortex tubular dilation
2-yr.
Diet
2-yr.
Diet
2-yr.
Diet
Rat, Rat,
2
1.25
Oral 2-yr.
No oncogenicity at
~HDT
(150 mg/kg/day)
Females: malignant mammary gland tumors
a References: I, Gammon et al. (1997); 2, Walker et al. (1974); 3, Bogdanffy (1990). bEstimated dosage, based on Lehman (1959). CEstimated NOEL (LOELllO).
33.3 Toxicology Profile
5,25, and 50 ppm, in males, at two years, by 57, 49, and 75%, respectively. However, because of the lack of a clear doseresponse and as this enzyme was not affected at 3, 6, 12, or 18 months, the toxicological significance of inhibition is also uncertain. Three other, chronic effects were reported by U.S. EPA (1994a) to be specific to cyanazine among the triazine herbicides: granulocyte hyperplasia of bone marrow in males, extramedullary hematopoiesis of the spleen in males, and demyelination of the sciatic nerve in females. However, although there was an increased level above control at the highest dose tested, none of the effects was statistically significant (Fisher's exact test). Following the inclusion of interim sacrifice (I-year) data, the increase in extramedullary hematopoiesis in the spleen of males was significantly elevated (p < 0.05) at the highest dose tested. Atrazine also caused an increase in extramedullary hematopoiesis in the spleen of the female rat, in a 2-year study. Oncogenic effects were apparent as increased incidences of malignant mammary gland tumors in females at 5, 25, and 50 ppm (Table 33.3), with no increase in males. There was a significantly increased incidence of tumors even at the NOEL for the principal nononcogenic effect (body weight loss) of 5 ppm. It therefore seems unlikely that these tumors resulted from a secondary effect of dosing, such as impaired homeostasis or increased cell death, which is often considered to result in tumors in chronic studies with high doses of xenobiotics. The increase in malignant tumors, which were principally adenocarcinomas, showed a dose-related positive trend (p < 0.001, Peto's trend test). There was a lower rate of adenocarcinoma incidence at the highest dose (29%) compared with the next highest dose (35%). This could be associated with the large relative fall in mean body weight gain at 2 years, at the highest dose, of 17% versus only 4% at 25 ppm. It is well established that reduced food consumption and decreased body weight lead to reduced incidences of neoplastic lesions in untreated rodents (e.g., Gellatly, 1975; Tannenbaum, 1948). Similarly, in rats treated with specific carcinogens, for example, N-methyl-N-nitrosourea (Beth et aI., 1987; Chevalier et al., 1993) or 7,12-dimethylbenz[a ]-anthracene (Klurfeld et al., 1989; Kritchevsky et al., 1989), dietary restriction resulted in a reduced incidence of mammary tumors compared with free-feeding rats. The additional cancers resulting from these carcinogens were abolished at 30 and 40% dietary restriction, respectively. It is therefore possible that the reduced food intake and fall in body weight compensated, to some extent, for the increased incidence of mammary tumors that would be anticipated at the HDT compared with the next lower dose of cyanazine. The figures showing tumor incidences were considered, by the study authors, to be significant only at 25 and 50 ppm when compared to the laboratory historical controls. These indicated that the concurrent control group level (10%) of malignant mammary tumors was below the Haskell Laboratory mean of 18% (87/476) from 1984 to 1989 and at the low end of the range of 10-23%. There was, however, no significant increase in benign mammary tumors resulting from cyanazine administration. For combined (malignant plus benign) mam-
711
mary tumors, the dose-related increase (Table 33.3) showed a positive trend (p < 0.01, Peto's trend test). However, only the incidence in the 25-ppm group was significantly different from the concurrent control. There were no statistically significant, dose-related increases in other tumors or in total tumors (Table 33.3). The NOEL for non-oncogenic effects was 5 ppm (0.20 mg/ kg/day) based on reduced body weight gain in both sexes, at 25 ppm, of7 and 13% at I year.
33.3.4.2 Dietary Exposure-Mouse Cyanazine technical was fed to CD mice (50/sex/level; lOO/sex, controls) at dietary concentrations of 0, 10, 25, 250, or 1000 ppm in a 2-year feeding study (Gammon et al., 1997). Mean body weights were depressed significantly, in both sexes at all treatment levels, and were dose-related. Reductions of 9% (males) and 11% (females, p < 0.01) were reported at 1000 ppm, within a week of study initiation; at termination, body weights were reduced by 25% (males) and 32% (females, p < 0.01) at 1000 ppm, and by 11% (males) and 15% (females, p < 0.01) at 10 ppm. A corresponding reduction in mean food intake was reported, for example, over weeks 1 to 52, food intake was reduced by 7% in males at 250 ppm (p < 0.01) and by 10% at 1000 ppm (p < 0.01); for females, the reductions were 7% (p < 0.05) and 5% (p < 0.05), respectively. Reduced food intake probably contributed to the lower body weights of dosed mice, but there were also significant reductions in food conversion efficiency for both sexes at 250 and 1000 ppm. This was measured as the mean body weight gain per unit weight of food consumed. These reductions were apparent during the first week (p < 0.01), weeks 1-13 (p < 0.05), and atthe conclusion of the study (p < 0.01). Lower food intake resulted in symptoms of poor skin condition, fur loss, reduced blood glucose, anemia, and adrenal cortical lipid depletion, at 250 and 1000 ppm. Also observed at the conclusion of the study were increased cases of cutaneous ulceration, myocarditis in males at 1000 ppm (p < 0.001), myocardial fibrosis in females, both basal and nonbasal, at 250 ppm (p < 0.05) and 1000 ppm (p < 0.001), focal renal cortical tubular dilation in females at 250 ppm (p < 0.05) and 1000 ppm (p < 0.05), and epithelial vacuolation in females at 250 ppm (p < 0.05) and 1000 ppm (p < 0.001). There were no oncogenic effects from treatment. Because of the significant reduction in body weight at all doses tested, 10 ppm (1.5 mg/kg/day) was the lowest-observed effect level (LOEL). An estimated NOEL of 0.15 mg/kg/day was established (by default) by dividing the LOEL by an uncertainty factor of 10.
33.3.4.3 Oral Exposure-Dog Cyanazine was administered daily by capsule to beagle dogs (4/sex/treatment group and 6/sexlcontrol group) at 0, 0.625, 1.25, or 5 mg/kg/day for 2 years (Walker et aI., 1974). Toxic effects related to treatment occurred at the highest dose level. Dogs in this group frequently vomited, within
712
CHAPTER 33
Risk Assessment for Chronic Exposure
1 hr of dosing, and showed reduced mean body weight, absolute liver weight, and total serum protein, throughout the test. The mean body weight was reduced, at the highest dose, for the duration of the test: even at 4 weeks, males (p < 0.05) and females (p < 0.01) had reduced body weights. At 1.25 mg/kg/day females had body weight reductions of 7% (4 weeks) to 17% (104 weeks), but only at 12 weeks was the (14%) decrease significantly different from control (p < 0.01). There were no consistent hematology or clinical chemistry findings. The NOEL was 0.625 mg/kg/day, based on reduced mean body weight at the two higher doses.
33.3.4.4 Dietary Exposure-Dog In a I-year dietary study, cyanazine was administered to beagle dogs (6/sex/level) in the feed at 0, 10,25, 100, or 200 ppm (Gammon et aI., 1997). Mean body weight and body weight gain for both sexes were depressed at 100 and 200 ppm. At 13 weeks, mean body weight was reduced, at 100 and at 200 ppm, by 15% (males) and 16% and 20% (females). At termination, the decrements were 12% (males) and 25% (females) for both 100 and 200 ppm. Mean food consumption was also depressed, particularly in the 200-ppm group. For males, a significant reduction in food intake was reported only for the first week, of 28% at 200 ppm (p < 0.05) and 18% at
Table 33.3 Malignant and Benign Mammary Tumors in Female Rats Fed Cyanazine for 2 Yearsa Dose, ppmb
5
25
50
41
48
51
6
12*
17**
15*
0
o
6
12*
18**
15*
(29%)
(38%)
(29%)
Tumortype
0
No. rats at riskc
49
43
Adenocarcinomad
5
Carcinosarcoma
0 5+++
Malignant
Combined malignant
(10%)
(14%)
o
Benign
4e
Adenoma
41
Combined Malignant
+ benign
9++ (18%)
9 (21%)
14
20'
16
(34%)
(42%)
(31%)
Other tumors Fibrosarcoma
0
Fibroadenoma
o
o
19
18;
17)
24k
(43%)
(44%)
(44%)
(35%)
(47%)
0
0
o
28
28
(57%)
(65%)
Granuloma
0
tumors m
Total
o
0
21h
Fibroma
o
2
31
28
37
(76%)
(58%)
(73%)
aData are from Bogdanffy (1990). bMeasured cyanazine intake (over 2 years) was 0, 0.04, 0.20, 0.985, and 2.06 mg/kg/day in males and 0, 0.053, 0.259, 1.37, and 2.81 mg/kglday in females. crncidences are expressed as the number of animals bearing tumors per animals at risk, defined as rats subjected to necropsy after at least 335 days, excluding interim sacrifice. dRats with multiple tumors account for the following proportions: 2/5 (control), 2/6 (1 ppm), 5/12 (5 ppm), 4117 (25 ppm), and 7115 (50 ppm); the others are single tumors. eIncludes 2 rats which also had fibroadenoma. f Includes 1 rat which also had adenocarcinoma. gIncludes 1 rat which also had adenocarcinoma, 2 rats had fibroadenoma, and 2 rats had both. hIncludes 1 rat which also had adenocarcinoma and 2 had adenoma. i Includes 3 rats which also had adenocarcinoma. j Includes 5 rats which also had adenocarcinoma, 2 had adenoma, and 2 rats had both. kIncludes 5 rats which also had adenocarcinoma, 1 had fibrocarcinoma, and 1 had sarcoma. m Includes all rats bearing 1 or more tumors, listed above. ++Significant trend (p < 0.01) based on dose-weighted chi-square test (Peto et aI., 1980). +++Significant trend (p < 0.001) based on dose-weighted chi-square test (Peto et aI., 1980). 'Significantly different from control (p < 0.05) based on Fisher's exact test. "Significantly different from control (p < 0.01) based on Fisher's exact test.
33.3 Toxicology Profile
100 ppm (n.s.). The food intake of females was reduced significantly for 3 of the first 6 weeks and subsequently, only at weeks 39 and 43. During these weeks, food intake was reduced by 10-28% at 100 ppm and 16-28% at 200 ppm. Thus, the reduction in body weight may not have been caused entirely by reduced food intake. Absolute organ weights were depressed by 10-30% for heart, lung, and spleen, and were increased (20%) for adrenals, in both sexes. Absolute liver weight was reduced only in females. None of the absolute organ weights were statistically different from control but relative organ weights (heart, lung, liver, adrenals, and kidneys) were increased significantly by 19-43%, in one or both sexes, largely because of the reduced body weight. All were elevated in the 200-ppm group (p < 0.05) whereas at 100 ppm, significant increases were limited to lung (19%) and kidney (20%), in females. Sporadic increases in platelet count and inorganic phosphorus with reduced total serum protein, albumin and calcium were dose-related but not always statistically significant. Neither the organ weight changes nor the hematological and clinical chemistry changes were associated with any histopathological changes. The NOEL from this study was 25 ppm (0.7 mg/kg/day) based on decreased body weight and body weight gain along with increased relative lung and kidney weights in both sexes. Chronic and subchronic toxicity studies are summarized in Table 33.2.
33.3.5 GENOTOXICITY Cyanazine caused genotoxic effects in four types of assay using mammalian cells, in vitro: gene mutations in mouse lymphoma cells, with and without metabolic activation; clastogenic activity in chromosomes of human lymphocytes, though not in CHO cells; unscheduled DNA synthesis (UDS) in rat primary hepatocytes; transformation in a mouse cell line, although only without metabolic activation. In nonmammalian cells, cyanazine caused gene mutation in E. coli, an increased response in the Drosophila dominant lethal assay, following dosing in vivo, as well as a variety of chromosome aberrations in plant cells. However, the in vivo evidence suggests that cyanazine may not be genotoxic in mammals. For example, in rat hepatocytes and spermatocytes, cyanazine did not cause UDS after in vivo administration. Genotoxicity tests with cyanazine are summarized in Table 33.4. 33.3.6 REPRODUCTIVE TOXICITY The toxicity of cyanazine in a two-generation rat reproduction study included reduced food intake and body weight in adults. Pup body weight and food intake were also lowered, in a dosedependent manner, and pup viability (survival) was reduced. In pups the effects on body weight and viability occurred at lower
Table 33.4 Summary of Genotoxicity Tests with Cyanazine Results
Route
Test
Reference
Gene Mutation See Gammon et al. (1997)
Bacteria, S. typhimurium
In vitro
Bacteria, E. Coli
In vitro
Mouse lymphoma
In vitro
CHOcells
In vitro
See Gammon et al. (1997)
Mouse bone marrow
In vivo
See Gammon et al. (1997)
S. cerevisiae gene conversion
In vivo
See Gammon et al. (1997)
Mouse dominant lethal
In vivo
See Gammon et al. (1997)
+ +
Venkat et al. (1995) See Gammon et al. (1997)
Structural Chromosomal Aberration
Drosophila dominant lethal
In vivo
Human Iymphocytes
In vitro
Human Iymphocytes
In vitro
CHO cells, c1astogenicity
In vitro
Barley shoot tips
In vivo
Broad bean roots
In vivo
Tradescantia roots
+
Mumik and Nash (1977) See Gammon et al. (1997)
+
Roloff et al. (1992) Taets et al. (1998); Biradar and Raybum (1995a, b)
In vivo
+ + +
Ahmed and Grant (1972)
Rat hepatocytes, UDS
In vitro
+
See Gammon et al. (1997)
Rat hepatocytes, UDS
In vivo
Grilli et al. (1991)
Rat spermatocytes, UDS
In vivo
See Gammon et al. (1997)
BALB/c-3T3 cell, cytotoxicity
In vitro
Kahlon (1980) Ahmed and Grant (1972)
Other Genotoxic Effects
and transformation
w/S-9
w/o S-9
Perocco et al. (1993)
+
713
Perocco et al. (1993)
714
CHAPTER 33
Risk Assessment for Chronic Exposure
doses than did reduced body weight in adults, indicating a possible adverse effect on reproduction.
33.4 RISK ASSESSMENT 33.4.1 HAZARD IDENTIFICATION
33.3.6.1 Dietary Exposure-Rat The reproductive effects of cyanazine in rats have been reported in two studies. In the first, cyanazine was tested in Long Evans rats at 0, 3, 9, 27, and 81 ppm in the diet (10 males and 20 females per dose level) over three generations (see Gammon et aI., 1997). The report contained limited data and showed slight reduction in terminal body weights at the 81-ppm level of 5-13% (males) and 5-10% (females), in all generations. No NOEL was derived from this study due to limited data. The second investigation was a two-generation reproduction study of cyanazine (Nemec, 1987) in Sprague-Dawley COBS CD rats (28 rats/sex/dose level). These were fed cyanazine in the diet at 0, 25, 75, 150, or 250 ppm over two generations, commencing 72 days prior to the first pairing. Decreases in body weight gain and food intake during the Fo, F 1, and F2 generations were reported at the 75, 150, and 250 ppm levels. The NOEL for decreased body weight in adults (Fo) was determined to be 150 ppm, equivalent to 11.2 mg/kg/day, based on a 10% fall in body weight increase in males from week 6 to week 30 at 250 ppm (p < 0.01, Dunnett's test). At 150 ppm, there was a statistically significant fall in body weight of 5% (p < 0.01), but this was not considered biologically significant. Body weight gain in Fla pups was decreased by 18% (p < 0.01) from day 4 to day 21 at 150 ppm, but by only 10% at 75 ppm. Subsequent generations were not clearly affected by cyanazine dosing. Because body weight was reduced significantly by over 10%, for most of the dosing period, at both 150 and 250 ppm but not at 75 ppm, the latter value was selected as the NOEL for this effect in pups. Reduced pup viability occurred on days 14 and 21 in Fl a pups at 250 ppm (p < 0.01) and on days 1 and 4 in the F2a pups at 150 ppm (p < 0.01) and at 250 ppm (p < 0.05). Five out of 22 dams had total litter loss between days 11 and 19 at 250 ppm in F la. The NOEL for reduced pup viability was 75 ppm, equivalent to 5.6 mg/kg/day. The reproductive parameters (male and female fertility, gestation length and parturition) were not affected by cyanazine.
33.3.7 DEVELOPMENTAL TOXICITY Developmental toxicity of cyanazine was assessed in three oral gavage studies in the rat and in oral gavage and dermal exposure studies in the rabbit. Fetal malformations (microphthalmia and anophthalmia) were noted in two oral rat studies and in a single litter in the rabbit, after oral gavage. Maternal toxicity was noted in all studies in the form of weight loss and reduced food intake. Quantitatively, these effects occurred at similar dose levels as the fetal effects, in the rat and rabbit (see Section 33.4.1.2 for doses). These studies have previously been analyzed in depth by Iyer et al. (1999) and Gammon et al. (1997).
33.4.1.1 Summary Cyanazine consistently suppressed appetite in experimental animals, usually with a concomitant fall in body weight. A reduction in body weight was observed after the administration of cyanazine by gavage, inhalation, dermal, or dietary exposure. In all but the latter case, loss of appetite could not have resulted directly from reduced palatability. Other triazine pesticides, such as atrazine, simazine, and cyromazine, also cause this effect, regardless of the duration of exposure. V.S. EPA has used the end point of reduced body weight, along with increased hyperreactivity, to define the RID for cyanazine (0.002 mg/kg/day). Following maternal dosing, cyanazine caused anophthalmia and microphthalmia in the rat and rabbit fetus or pup, sometimes at dose levels causing little or no discernible maternal toxicity. Evidence for genotoxicity was produced in a variety of in vitro tests. In a two-generation rat reproductive toxicity study, reduced pup viability was observed at doses below that which reduced adult body weight. In a chronic study using the Sprague-Dawley rat, evidence of a compound-related increase in malignant mammary tumors in females was produced. 33.4.1.2 Acute Toxicity Acute oral toxicity studies in the rat indicated LDso values of 835 (males) and 369 (females) mg/kg. By inhalation, in the rat, cyanazine dust had LDso > 152 mg/kg after a 4-hr exposure, with an estimated NOEL of 1.6 mg/kg (see Gammon et aI., 1997). The most sensitive groups of animals for determining the acute toxicity of cyanazine were dams and/or offspring of rabbits. Cyanazine caused developmental as well as maternal toxicity. The NOEL value of 1 mg/kg/day for oral exposure of the rabbit (lyer et al., 1999) was used as the critical NOEL to assess the acute dietary and occupational exposures. This oral NOEL is of very similar magnitude to the rabbit dermal NOEL of 1.3 mg/kg, based on mean dermal absorption of 1.5%. 33.4.1.3 Chronic Toxicity The submitted summaries of studies suggest that the only, consistent dose-related effect of cyanazine was the loss of body weight. In the dietary studies there was also a reduction in food intake in the rat, but not in the mouse. In dermal (rabbit) and inhalation (rat) studies, the body weight reduction did not appear to be consistently accompanied by a loss of appetite. This effect was observed in rats, mice, and dogs. The NOEL values for this effect were: 0.2 mg/kg/day in rats, 2-year study (Bogdanffy, 1990) 0.15 mg/kg/day in rats, 2-year study (Walker et aI., 1974) 0.15 mg/kg/day (estimated) in mice, 2-year study (see Gammon et aI., 1997) 0.7 mg/kg/day in dogs, I-year study (see Gammon et aI., 1997).
33.4 Risk Assessment
Cyanazine feeding generally caused food to be poorly palatable, resulting in lower food intake, which may partly explain the reduced body weight gain in dietary studies. In addition to reduced body weight gain in mice, cyanazine feeding resulted in toxicological adverse effects of increased renal cortical tubular dilation and epithelial vacuolation and myocarditis (see Gammon et aI., 1997). The lowest measured NOEL from an acceptable study (5 ppm or 0.2 mg/kg/day), for systemic toxicity in the rat (Bogdanffy, 1990), was used as the critical NOEL for evaluating nononcogenic effects. This is the same chronic NOEL value used by V.S. EPA (1994a). The next highest dose, 25 ppm, resulted in significantly reduced body weight and body weight gain, in both sexes, and hyperreactivity in males. However, because hyperreactivity was not clearly defined in the report and showed a discontinuous dose-response, it is considered to have doubtful toxicological relevance. 33.4.1.4 Oncogenicity
Cyanazine chronic feeding in rats resulted in a statistically significant increase in combined malignant mammary gland tumors (adenocarcinomas and carcinosarcomas) in female rats administered cyanazine in the diet at 5, 25, and 50 ppm, but not at 1 ppm (Table 33.3; Bogdanffy, 1990). There was no compound-related increase in benign mammary tumors (adenomas); however, when rats having adenoma(s) were combined with those having malignant tumors, an elevated incidence of tumors was observed, but only at 25 ppm. The significant increase in mammary tumors at the highest dose tested, 50 ppm, was accompanied by significantly reduced food intake and body weight. It has been shown that, in the untreated rat, a reduction of mammary tumor incidence accompanies reduced food intake and body weight (Boorman et aI., 1990; lp, 1991; Turnbull et aI., 1985). Characteristically, a lower caloric intake is associated with increased lifespan and reduced carcinogenicity in rodents (e.g., Kritchevsky and Klurfeld, 1987; Seilkop, 1995). It is thus possible that cyanazine would be a more potent carcinogen if the dietary intake of all the rats had been restricted to that of the high-dose group. The increase in rat mammary tumors which resulted from cyanazine administration was considered toxicologically significant. However, the Carworth Farm E strain rat did not show an increase in mammary gland tumors in two earlier studies (see Walker et aI., 1974) nor were there increased tumors in a mouse study. It should be noted that other triazine pesticides cause elevated levels of mammary tumors (adenocarcinomas and adenomas) in female Sprague-Dawley rats: atrazine, simazine (Hauswirth and Wetzel, 1998), cyromazine (DPR, 1993), propazine and terbutryn (U.S. EPA, 1991a) along with a fall in body weight. There is evidence that cyanazine has genotoxic potential, as shown by results from four types of in vitro assay using mammalian cells. The same assays using S-9 metabolic systems were inactive. In vivo studies suggest that cyanazine may not have genotoxic potential in mammals. For example, in rat hepatocytes (Grilli et aI., 1991) and rat spermatocytes (see Gammon
715
et aI., 1997), cyanazine did not cause unscheduled DNA synthesis after in vivo administration. The assessment of the potential oncogenic risk of cyanazine in humans was evaluated using a quantitative, low-dose extrapolation approach. A nonthreshold mechanism was assumed because a biological mechanism has not been convincingly demonstrated. This approach is consistent with that used by V.S. EPA (1985). The linearized multistage model, Global 86 (Howe et aI., 1986) and the Weibull82 (time-to-tumor) models were used to calculate cancer potency factors in female rats. By extrapolating the dose-response curve (linearly) to low doses, potency values were estimated based on the incidences of rat combined malignant mammary tumors (Table 33.3). The malignant mammary tumors were considered because of the greater statistical significance of the increased incidence with dose and because they are of greater relevance to human health than benign tumors. Both the maximum likelihood estimate (MLE, QI) and the 95% upper confidence limit (95% VB, Qi) of the linear term of the multistage model were calculated as estimates of oncogenic potency. The equations relating the probability of cancer (P) to dose (D) and cancer potency (Q) are as follows: P(DOSE) = 1 - exp(-Qo - QID - Q2D2 ... - QxDX) P(DOSE)
=
(Global 86) 1 - exp(-Qo - QID - Q2D2 ... - QxDX) x (T - To)J
(Weibull, time-to-tumor)
Cancer potency values (QI and Qr) were identical using the two models. This indicates that the time function coefficient term (T - To)J is unity, that is, that there is not an earlier onset of tumors as a result of cyanazine administration. Equivalent human potency values were estimated using a body-weight conversion factor assuming an interspecies dose equivalence of body weight to the 3/4 power, from the rat values, using the equation (QI, human)/(QI, rat) = [(body weight, human)/(body weight, rat)]I/4. Vsing combined malignant mammary tumors, the human cancer potency values for cyanazine were 0.33 (mg/kg/day)-I for the maximum likelihood estimate QI and 0.58 (mg/kg/day)-I for the 95% upper bound confidence interval Qr. These values were used to estimate potential oncogenic risk from occupational and dietary exposures. These potency values are greater than values for other structurally related triazines. In comparing Qr values, cyanazine was 8.3-fold more potent than simazine, 4.5-fold more potent than atrazine (V.S.EPA, 1991b;V.S.EPA, 1994a), and ca. lOO-fold more potent than cyromazine (DPR, 1993). However, it should be noted that only for cyanazine are the QI values based on combined malignant mammary tumors; for the others, the QI values are based on combined malignant and benign tumors. For cyanazine, V.S. EPA originally estimated a Qi of 0.159 (mg/kg/day)-I for the rat and 0.84 (mg/kg/day)-I for the human, for adenocarcinomas and carcinosarcomas, using the linearized multi stage model (Global 86). This calculation, which was used in the V.S. EPA risk characterization, included interim sacrifice animals (V.S. EPA, 199Ib). Subsequently, V.S. EPA decided that interim animals should not, in general, be included
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in the calculation of a potency factor for lifetime exposure, regardless of whether some animals already exhibited tumors (U.S. EPA, 1993). U.S. EPA recalculated a QT value of 1.0 (mg/kg/day)-l for humans, equivalent to 0.2 (mg/kg/day)-l for the rat, based on a body-weight scaling factor of the 2/3 power (U.S. EPA, 1993). However, it is considered more appropriate to use a (body weight)3/4 instead of a (body weight)2/3 scaling factor in the calculation of animal-to-human dose equivalence (Fenner-Crisp, 1994; Travis and White, 1988). This accounts for the majority of the difference between the two calculations of the human QT values. Another difference is the calculation of daily dosage in the rat chronic study (Bogdanffy, 1990). DPR used average, measured chemical intake and U.S. EPA used rat default mean dietary intake (Lehman, 1959). 33.4.2 EXPOSURE ASSESSMENT 33.4.2.1 Occupational Exposure The absorbed daily dosage (ADD), annual average daily dosage (AADD) and lifetime average daily dosage (LADD) were estimated for workers using a study of the ground application of Bladex 4L to corn (Sanborn and Mehler, 1996). Although ca. 90% of cyanazine use in California is on cotton, the corn study is considered a suitable surrogate, once it had been adjusted for the lower application rate for cotton in California (averaging 2.0 lb per acre versus 4.5 lb per acre for corn). The 4L (liquid) formulation is only one of two cyanazine products currently registered in California, the other being the 90DF (granule). They are used in approximately equal amounts on cotton in California. Mixer-Loader-Applicator (Commercial or Farmer) The exposure of a mixer-Ioader-applicator (MLA) involved in the ground application of Bladex to corn resulted in an estimated mean absorbed daily dosage of cyanazine of 2.6 J.!g/kg/day (95% c.1. = 5.0 J.!g/kg/day), a 95th percentile (high-end) exposure of 24.6 J.!g/kg/day and an annual average daily dosage of 0.11 J.!g/kg/day, based on 15 days' use per year (Table 33.5). These dosage estimates are based on 2% dermal absorption, from a rat dermal penetration study (see Gammon et al., 1997). This was the same dermal absorption used by U.S. EPA (U.S.
EPA, 1994a, b). Over an occupational lifetime of applying cyanazine (i.e., 40 of 75 years), the lifetime average daily dosage would be 0.057 J.!g/kg/day (Table 33.5). For a farmer applying cyanazine, the ADD is expected to be the same as for the commercial applicator but, because of the reduced number of days of exposure per year (3), the AADD and LADD values would be correspondingly lower than for the commercial applicator. 33.4.2.2 Dietary Exposure Residue Data
Primary and Secondary Residues
Residue studies in raw agricultural commodities (RACs) were conducted by the former registrant, Shell Oil Co. (Table 33.1). The reasons that they were used to assess dietary exposure are as follows: a very low limit of detection (LOD) of 0.01 ppm, a complete range of crops for which registrations were being sought and for which tolerances were obtained, and a range of application rates, including levels above the current maximum label application rates. In addition, residues of four plant metabolites of cyanazine were measured in these studies, with LODs of 0.03 or 0.05 ppm. Residue studies conducted by DPR had a LOD of 0.2 ppm and revealed no cyanazine residue detections from 1988 to 1993. The U.S. Food and Drug Administration (FDA) LOD for residues of cyanazine was 0.04 ppm, using a multiresidue screen and also revealed residues below the LOD. When considered in combination with the registrant's data (Table 33.1), it is clear that residues in crops at harvest are hypothetical. There have been no determinations of residues in secondary animal products such as beef, pork, poultry, sheep, and eggs. This is because registrant studies have demonstrated that cyanazine (even if it were present in animal feed) does not concentrate in animal tissues (see Gammon et aI., 1997).
Drinking Water
Cyanazine has been frequently detected in ground water and surface water of the principal corn-growing states of the central United States, that is, Illinois, Indiana, Kansas, Missouri, Nebraska, and Ohio (see Cohen et aI., 1995; Wiles et aI., 1994). Consequently, since 1990, DPR has monitored for (parent) cyanazine in groundwater from regions of
Table 33.5 Occupational Exposure to Cyanazin& Worker
ADDb
Commercial MLAg
2.61,g 2.61,g
Farmerg
(l1g/kg/day)C 0.11
0.021
0.057 0.011
aCalculations of worker exposure, based on a Bladex 4L study on corn. b ADD = absorbed daily dosage; AADD = annual average daily dosage; LADD = lifetime average daily dosage. cGeometric mean ADD. d Applications per year = 3 days (farmer), 15 days (commercial applicator); AADD = ADD x (3 or 15)1365. e Assumes 40 years of exposure, over a 75-year lifetime; LADD = AADD x 40175. f 95th percentile = 24.6 I1g/kg/day. gBladex study conducted with 12 replicates: 3 workers and 4 loads each.
33.4 Risk Assessment
California with a high usage of this pesticide. Cyanazine has never been detected at a LOD of :s 0.1 ppb. A degradation product of cyanazine, desisopropyl atrazine (DIPA) is also a common degradate of atrazine and simazine. DIPA was the fourth most frequently detected pesticide in California groundwater in 1995 (DPR, 1996). Simazine and atrazine (parents) ranked first and fifth, respectively, for number of detections in groundwater in 1995. This ranking was similar to previous years. Acute Exposure None of the field trial or surveillance data showed any detectable residues, at LODs of 0.2, 0.04, or 0.01 ppm, for any of the RACs listed in Table 33.1. Therefore, the limit of detection of 0.01 ppm was used as a default for the estimation of acute dietary exposure (Table 33.6). For the estimation of drinking water exposure, there have been no detections of cyanazine at a LOD of :s 0.1 ppb (1111 wells from 24 counties). Therefore, 0.1 ppb was used as the default concentration in drinking water. Judging from the use patterns of these triazines, it is possible that some of the DIPA detections could have resulted from the use of cyanazine, although most of them probably resulted from the use of simazine on citrus. Chronic Exposure Field residue trials (Table 33.1) showed that cyanazine (parent) residues were not detected in any crop at harvest at the LOD of 0.01 ppm and that (four) identified transformation products were not detected at LODs of 0.03 or 0.05 ppm. Therefore, default residues of 0.005 ppm (50%
of LOD) were used for each commodity for the estimation of potential chronic (annual) dietary exposure (Table 33.6). The values presented in Table 33.6 assume that 100% of the commodities were treated with cyanazine. Data on the percentage of crop treated indicate that approximately 30% of corn and cotton and 10% of sorghum or wheat are treated with cyanazine in California (DPR, 1996). Therefore, the theoretical residue values, and resultant chronic exposure values, would be reduced accordingly. For the potential exposure to cyanazine residues in drinking water, as mentioned previously, there were no detections in ground water at a LOD of 0.1 ppb. Therefore, 0.05 ppb was used as a default residue level to estimate potential chronic exposure through drinking water. 33.4.2.3 Dietary Exposure Analysis Acute Exposure Acute dietary exposure analyses were conducted using the Exposure-4TM program of Technical Assessment Systems, Inc. (TAS). This program estimates the distribution of user-day (consumer-day) exposure for the overall U.S. population and specific population subgroups (TAS, 1992a). Exposure is calculated from knowledge of pesticide residues in each food item and the amount of each food consumed. A user-day is any day in which at least one food from the specific commodity list is consumed. The analysis uses data from the U.S. Department of Agriculture (USDA) Nationwide Food Consumption Survey (USDA, 1987-1988).
Table 33.6 Potential Dietary Exposure to Cyanazine in All Commodities with V.S. EPA Tolerances and in Drinking Water Acute exposurea
Chronic exposureb
Population subgroup
().1g/kg-day)
().1g/kg-day)
V.S. population, all seasons
0.074 (0.078)C
0.013 (0.015)d
Western region
0.069
0.013
Hispanics
0.070
0.0l3
Non-Hispanic whites
0.073
0.0l3
Non-Hispanic black
0.083
0.014
Non-Hispanic other
0.071 0.102 (0.066)C
0.004 (0.006)d
Infants (nursing)
0.0l3
Infants (nonnursing)
0.160 (0.176),
0.018
Children (1-6 years)
0.l32 (0.138)C
0.031 (0.033)d
Children (7-12 years)
0.093
0.022
Females (13-19 years, not pregnant or nursing)
0.056
0.0l3
+ years, pregnant, not nursing) Females (13 + years, nursing) Females (20 + years, not pregnant or nursing)
0.042
0.010
0.045
0.011
0.040
0.009
Males (13-19 years)
0.067
0.016
Males (20 + years)
0.046
0.010
Females (13
Seniors (55
+ years)
717
0.038
a95th percentile of dietary exposure (residues = LOD, i.e., 0.01 ppm for corn, sorghum, wheat, and cottonseed). b Annual average dietary exposure (residues = 50% of LOD, i.e., 0.005 ppm for corn, sorghum, wheat, and cottonseed). Based on 100% crop treated. cIncludes theoretical drinking water residues of cyanazine, 0.1 ppb (LOD). dIncludes theoretical drinking water residues of cyanazine, 0.05 ppb (50%LOD).
718
CHAPTER 33
Risk Assessment for Chronic Exposure
Using the 95th percentile of user-day exposures for all specific population subgroups, the potential acute dietary exposure to cyanazine from alllabeled uses ranged from 0.038 to 0.160 !l-g/kg-day (Table 33.6). Infants (nonnursing, < 1 year) had the highest and seniors (55 + years) the lowest potential acute dietary exposure to cyanazine. Potential exposure through drinking water was also estimated using the TAS Exposure-4 program. This would increase the potential exposure to nonnursing infants to 0.176 !l-g/kg-day, a 10% increase. The potential exposure of the U.S. population would be 0.074 !l-g/kg-day, without water, and 0.078 !l-g/kg-day, with water, a 5% increase. Exposure of nursing infants would be reduced, from 0.102 to 0.066 !l-g/kg-day, with the inclusion of drinking water, and children (1-6 years) would have an increased exposure, from 0.132 to 0.138 !l-g/kg-day, a 5% increase, with the inclusion of drinking water. Chronic Exposure The potential chronic dietary and drinking water exposure was calculated using the Exposure-l ™ software program (TAS, 1992b). The food consumption data for the chronic analysis was also based on the 1987-1988 USDA Nationwide Food Consumption Survey (USDA, 1987-1988). The program estimates the annual average daily exposure for specific population subgroups. In addition to calculations of theoretical dietary exposure assuming 100% treatment of each registered crop with cyanazine, calculations were made adjusting for percentage of crop treated in California. All potential dietary exposure was pooled by combining cyanazine residues in all commodities on which cyanazine use is registered (Table 33.6). The mean potential annual dietary exposure ranged from 0.004 (nursing infants) to 0.031 !l-g/kgday (children, 1-6 years), based on 100% of crop treated. Percentage of crop-treated adjustment factors were 30% for corn and cotton; 10% for sorghum and wheat. The equivalent mean potential chronic dietary exposure levels, adjusted for percentage of crop-treated, were 0.001 and 0.006 !l-g/kg-day, for the same sub-populations (not shown). In addition, potential exposure to cyanazine through drinking water was also calculated, at 0.05 ppb (50%LOD). For the U.S. population, all seasons, drinking water increased the potential chronic exposure to cyanazine from 0.013 to 0.015 !l-g/kg-day (15%). Potential exposure for nursing infants would be increased from 0.004 to 0.006 !l-g/kg-day. At the upper end of the chronic exposure range, children (1-6 years) would experience a calculated increase from 0.031 to 0.033 !l-g/kg-day (not adjusted) with the inclusion of potential drinking water residues, a 6% increase. Thus, the theoretical chronic dietary exposure to cyanazine using residue data adjusted for percentage of crop treated was reduced to between 18 and 28 % of the exposure calculated for 100% crop treated. 33.4.2.4 Combined Occupational and Dietary Exposure Assessment Acute Exposure The combined acute exposure was obtained by summing the mean (occupational) ADD of 2.6 !l-g/kg/day
(Table 33.5) and the acute dietary exposure (Table 33.6) for males, 13-19 years (0.067 !l-g/kg/day) or 20 + years (0.046 !l-g/kg/day), the subgroups most likely to experience occupational exposure. This gave a total, acute, combined occupational and dietary exposure of 2.7 !l-g/kg/day, with or without drinking water included, for males 13-19 years. For males of 20 + years, the estimated combined acute exposure was 2.6 !l-g/kg/day. Chronic Exposure The combined chronic exposure was obtained by summing the AADD values of 0.021 and 0.11 !l-g/kg/ day, for farmers and commercial applicators, respectively (Table 33.5) and the potential chronic dietary exposure for males, 13-19 years or 20 + years, of 0.016 or 0.010 !l-g/kg/day, respectively (Table 33.6). The inclusion of theoretical drinking water exposure at 0.05 ppb (50%LOD) increased the dietary exposure value by approximately 0.002 !l-g/kg/day. Total, chronic, combined occupational and dietary exposure estimates were 0.037 !l-g/kg/day for farmers and 0.126 !l-g/kg/day for commercial applicators, or 0.039 !l-g/kg/day and 0.128 !l-g/kg/day, with the inclusion of drinking water exposure. 33.4.3 RISK CHARACTERIZATION
The risk characterization process consists of calculating a margin of safety by dividing the critical acute or chronic NOEL value for a specific toxicological end point by an estimate of human exposure. The probability of excess cancer risk in a lifetime was calculated by multiplying the LADD values (occupational) and/or the chronic annual average dietary exposure, by the cancer potency factors. Additionally, the cancer risk was calculated for combined occupational and dietary exposure, through the consumption of theoretical crop residues, with and without the inclusion of theoretical drinking water residues. 33.4.3.1 Occupational Exposure
The estimates of occupational exposure, following B1adex application to cotton, are given as the ADD, AADD, and LADD (Table 33.5). These estimates were used to calculate the acute and chronic MOS, as well as the probability of excess cancer risk in a lifetime, respectively (Table 33.7). The acute MOS, based on the mean ADD, for farmers and commercial applicators was 385. For workers exposed to the 95th percentile of the ADD, the MOS was 41. The annual MOS, based on the mean AADD, was 1820 for commercial applicators and 9520 for farmers. The probability of excess cancer risk in a lifetime was 1.9 x 10- 5 (MLE) and 3.3 x 10-5 (95%UB) for commercial applicators and 3.6 x 10-6 (MLE) and 6.4 x 10- 6 (95%UB) for farmers. 33.4.3.2 Dietary Exposure Acute Exposure The margin of safety for each population subgroup for theoretical acute dietary exposure to cyanazine is given in Table 33.8. These values were derived from the theoret-
33.4 Risk Assessment
719
Table 33.7 Margins of Safety and Excess Risk from Potential Occupational Exposure to Cyanazine Acute Mosa,b
Lifetime risk'" Chronic MOSC
Worker
Mean
95th percentile
Commercial MLA
385
41
1820
1.9 x 1O~5
3.3 x 1O~5
Farmer
385
41
9520
3.6 x 1O~6
6.4 x 1O~6
MLE
95%UB
aMOS = NOEL/ADD; NOEL of 1 mg/kg/day from a rabbit oral developmental toxicity study (see Iyer et aI., 1999).
bMean ADD (2.6 flglkg/day) and 95th percentile (24.6 flg/kg/day), from Table 33.5. cMOS = NOEL/AADD; NOEL (chronic) of 0.2 mg/kg/day from a 2-year rat study (Bogdanffy, 1990). AADD values of 0.11 (commercial applicator) and 0.021 (farmer) flg/kg/day, from Table 33.5. dBased on the product of LADD values (Table 33.5) and human cancer potency factor (Q!, MLE, and Qj 95% confidence interval, UB) derived from malignant mammary tumors in the female rat (Bogdanffy, 1990).
ical dietary exposure values (Table 33,6) in which all registered commodities were assumed to contain residues at the default level of the LOD. The MOS values ranged from 6270. for nonnursing infants « 1 year), to 26,300 for seniors (55 + years). The inclusion of theoretical drinking water residues at 0.1 ppb (LOD) reduced the MOS for the V.S. population, all seasons, from 13,500 to 12,800, a 5% decrease.
Chronic Exposure The margin of safety for each population subgroup following theoretical chronic (annual) dietary exposure to cyanazine is given (Table 33.8). These values were derived from the theoretical exposure values (Table 33.6) in which all registered commodities were assumed to contain residues at the default level of 50% of the LOD. The MOS values ranged from 6,440, for children (1-6 years), to 48,100, for nursing
Table 33.8 Margins of Safety for Theoretical Dietary Exposure to Cyanazine Residues in All Commodities with U.S. EPA Tolerancesa
Population subgroup
Acute
Chronic
margin of safetyb
margin of safetyC
U.S. population, all seasons
13,50(Jd
15,oooe
Western region
14,400
15,500
Hispanics
14,200
16,000
Non-Hispanic whites
13,700
15,100
Non-Hispanic blacks
12,100
14,300
Non-Hispanic other
14,100
15,300
Infants (nursing, <1 year)
9,850
48,100
Infants (nonnursing, <1 year)
6,270
11,300
7,560 10,800
6,440
17,700
15,300
23,800
20,000
22,400
18,600
24,700
22,100
Children (1-6 years) Children (7-12 years) Females
(13~19
years)
8,950
(not pregnant, not nursing) Females (13
+ years)
(pregnant, not nursing) Females (13
+ years)
(nursing) Females (20 + years) (not pregnant, not nursing) Males (13-19 years)
14,900
12,700
Males (20 + years)
21,800
19,400
Seniors (55
+ years)
26,300
aResidues = LOD (acute), i.e., 0.01 ppm or 50%LOD (chronic), i.e., 0.005 ppm for corn, sorghum, wheat, and cottonseed. bMOS = [acute NOEL (l mg/kg/day)]/(dietary intake, 95th percentile). cMOS = [chronic NOEL (0.2 mg/kg/day)]/(annual average dietary exposure). dMOS including theoretical drinking water exposure at 0.1 ppb (LOD) = 12,800. eMOS including theoretical drinking water exposure at 0.05 ppb (50%LOD) =13,300.
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CHAPTER 33 Risk Assessment for Chronic Exposure
infants. Crop-treated adjustment factors elevated these MOS values to 31,600 and 186,000 for these two groups, respectively. The inclusion of theoretical drinking water residues at 0.05 ppb (50% LOD) reduced the MOS values for the U.S. population, from 15,000 to 13,300 (unadjusted), an 11 % fall, and from 75,100 to 50,000 (adjusted for percentage of crop treated). Lifetime Exposure The excess risk of oncogenicity calculated to result from theoretical dietary exposure to cyanazine was estimated for the U.S. population (Table 33.9). It is assumed that dietary exposure would be the same every year over a lifetime. Using the MLE for cancer potency, QI (0.33 per mg/kg/day), and the range of potential chronic dietary exposures (0.003-0.013 l-lg/kg-day, based on adjustment for percentage of crop treated), the cancer risk was 1.0-4.3 x 10-6 . For the upper bound cancer potency factor, (0.58 per mg/kg/day), the excess cancer risk from potential dietary exposure was 1.57.7 x 10-6 . Potential exposure to cyanazine through drinking water would increase the theoretical cancer risk, from dietary exposure, by 10-16% (unadjusted for percentage of crop treated) or 30-50% (adjusted for percentage of crop treated, Table 33.9). U.S. EPA (1994a) calculated a 95% UB cancer risk estimate of 2.9 x 10- 5 for potential dietary exposure to all registered RACs. However, the anticipated residues used in this calculation were above tolerances, in order to estimate combined triazine exposure, as follows: corn, 0.12 ppm; cotton seed, 0.09 ppm; sorghum, 0.10 ppm; and wheat, 0.16 ppm. The tolerances for these RACs are 0.05 ppm, except for wheat, 0.10 ppm. In addition, U.S. EPA included anticipated secondary residues in milk, poultry, eggs, and red meat at 0.28 ppb, 2.3-4.2 ppb and 3.510.3 ppb, respectively. Any residue which is detected above tolerance in a RAC or detected, at all, in a commodity for which tolerances do not exist, would be illegal and the food would not be allowed to be sold for human consumption. Such residues are therefore not used in dietary exposure calculations. U.S. EPA (1994a) calculations of 95% UB cancer risk estimates were conducted on individual crops: 1.2 x 10-5 (corn), 9.3 x 10- 8 (cotton), 1.2 x 10-7 (sorghum), 2.3 x 10- 6 (wheat) plus secondary residues in milk, eggs, chicken, and red meat, totaling 2.9 x 10-5 (UB) or 1.6 x 10-5 (UB) for just the four RACs with tolerances. If U.S. EPA had based their calculations on the residue levels at the tolerances for the RACs, the excess cancer risks would likely be similar to those calculated (Table 33.9).
Qr
33.4.3.3 Combined Occupational and Dietary Exposure Because dietary exposure to cyanazine is largely theoretical, and because it is much less than occupational exposure, margins of safety and excess oncogenic risk are not given for combined occupational and dietary exposure. For example, for acute exposure (U.S. population), the MOS decreased by only 3%, from 385, for occupational exposure, to 374, adding dietary exposure. The addition of drinking water exposure to combined gave
Table 33.9 Excess Cancer Risk from Theoretical Dietary Exposure to Cyanazine Lifetime riska Dietary exposure
MLE
UB
No drinking water
4.3 x 10-6
7.7 x 10- 6
With drinking water
5.0 x 10- 6
8.5 x 10- 6
aCalculated by multiplying the cancer potency factor Q1 or Qj by the theoretical, annual average dietary exposure (U.S. popUlation), not adjusted for percentage of crop treated.
a MOS of 373, also a decrease of 3% below occupational exposure alone. 33.4.4 CONCLUSIONS A margin of safety of at least 100 is generally considered adequate to protect people from the toxic effects of a chemical when the NOEL is based on toxicology data from animal studies. Mean, short-term worker exposure data resulted in MOS values above 100 for both farmers and commercial applicators when calculated using abnormalities in the rabbit fetus as the toxicological end point. An estimated 95th percentile of acute exposure gave MOS values below 100 for these workers. Longterm occupational exposure data resulted in MOS values above 100 for both farmers and commercial applicators when calculated using weight loss in a rat chronic study as the toxicological end point. Excess lifetime cancer risk was greater than 10-5 (1 in 100,000) but less than 10-4 (1 in 10,000) for commercial applicators and greater than 10-6 (1 in 1,000,000) but less than 10-5 (1 in 100,000) for farmers. These values may be considered overly conservative because they are based on a working lifetime of 40 out of 75 years of exposure to cyanazine. Based on the available toxicity and residue data, it was concluded that the MOS values for potential acute (daily) and chronic (annual) dietary exposure, for all commodities for which D.S. EPA tolerances have been established, were above 100 for all population subgroups studied. The excess lifetime cancer risk for the general population was greater than 10-6 (l in 1,000,000) but less than 10- 5 (l in 100,000). It is considered, from a risk management perspective, that excess risks above 10-6 are excessively high for a single pesticide. The current U.S. cyanazine registrants, DuPont and Griffin, have voluntarily agreed to gradually phase out and eventually cancel the use of cyanazine, under certain conditions, by the year 2002.
33.5 DISCUSSION Risk assessment is the process which is used to evaluate the potential for human exposure and the likelihood that the toxic effects of a substance will occur under specific exposure conditions. In addition, ecological risk assessment may also be required based, in part, on the use patterns of the pesticide
33.5 Discussion and on the results of toxicology tests conducted on nontarget organisms. Every risk assessment has inherent limitations and uncertainties in the application of existing data to estimate the potential risk to human health. Therefore, certain a priori assumptions are incorporated into the hazard identification and exposure assessment processes. These, in turn, result in uncertainty in the risk characterization, which integrates all of this information. Qualitatively, risk assessment for all chemicals has similar types of uncertainty. However, the degree or magnitude of the uncertainty varies depending on the availability and quality of the data and the exposure scenarios being assessed. Varying degrees of uncertainty are involved in the estimation of these parameters, affecting the accuracy of the risk characterization. These have been divided into toxicity-related and exposure-related issues. Specific areas of uncertainty associated with the cyanazine risk assessment are appraised in the following discussion. 33.5.1 TOXICOLOGY
In the evaluation of the chronic toxicity of cyanazine, the most prevalent noncancer toxicological end point in rats, mice, and dogs was loss of body weight and a reduction in body weight gain. This effect was not solely a result of lower food intake due to poor palatability because reduced food intake was reported regardless of the route of administration of cyanazine (e.g., dermal, oral gavage, and inhalation, as well as dietary exposure). The NOEL for this effect in the rat was used to assess chronic exposure. However, the toxicological significance of body weight loss is difficult to assess. Indeed, male rats (though not females) showed significantly increased longevity associated with reduced body weight at the highest dose in the chronic dietary study. Other toxicological effects observed in chronic studies included inanition, poor skin and fur condition, and anemia, which may have all been secondary to poor nutrition resulting from reduced food intake. It is unclear to what extent these and other chronic effects of exposure are reversible or adaptive in nature following the discontinuation of dosing; that is, the effects may dissipate and the animal recover. In the United States, it is standard practice to evaluate cancer risk separately from other forms of chronic toxicity. Oncogenicity for cyanazine was assessed using a linear multi stage model which assumes a nonthreshold mechanism (LNT model). It should be noted that the Global 86 (multistage) model used in this assessment is only appropriate when the tumor(s) in question exhibit a dose-response. For high-do se-only carcinogens, where a threshold is assumed, such linearized models are inappropriate and such pesticides are commonly regulated with a MOS approach, sometimes using an additional safety factor. Evidence for genotoxicity is considered to support the use of a nonthreshold model and cyanazine was active in several types of such assay. On the other hand, evidence supporting a nongenotoxic mechanism is also considered and, when sufficiently strong, it will result in a MOS approach being used for risk assessment in place of a nonthreshold model. For example,
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frank organ toxicity or cytotoxicity may cause mitogenesis or cell proliferation resulting in cancer, usually specific to an organ and species, for example, male rat kidney tumors caused by d-limonene (Dietrich and Swenberg, 1991), and this would be assessed using a MOS approach. Although there was no evidence for toxicity to the rat mammary gland, it is possible that mammary tumors resulting from cyanazine exposure arose from an estrogenic (hormone-receptor-mediated) effect (Stevens et aI., 1994; Tennant et aI., 1994), which might be expected to show a threshold. This has been suggested for atrazine and simazine, where malignant mammary tumors have been found specifically in the same [Sprague-Dawley (SD)] strain and sex of rat (Hauswirth and Wetzel, 1998). It has been proposed (Simpkins et aI., 1998) that atrazine suppresses the surge in luteinizing hormone (LH) levels, leading to premature reproductive senescence, which in this strain of rat is associated with elevated blood levels of estradiol, which in turn results in an elevated mammary tumor incidence. Because there are differences between the aging of the human female reproductive system and that in the SD rat, it is possible that such a mechanism would have little relevance in human health risk assessment. However, as noted for the reduced cancer incidence which is normally associated with reduced body weight in the SD rat (Section 33.4), there is a confounding factor: dietary restriction of undo sed SD rats caused a slight delay in the onset of puberty but a large increase in reproductive life span (McShane and Wise, 1996). In experiments with ovariectomized rats, this increase was determined to be due to an increase in mean LH pulse amplitude which was attributed to an effect on the hypothalamic-pituitary axis. Thus, the effects of atrazine on reproductive aging in the SD rat (earlier persistent estrus, reduced LH), which are associated with reduced body weight, are precisely opposite the (normal) effects of caloric restriction. Other aspects of possible ways to address cancer risk, from a regulatory perspective, are discussed below. There has been considerable debate about the relevance of the use of nonthreshold models in cancer risk assessment. The 1998 Society of Toxicology debate on the subject voted overwhelmingly in favor of dropping nonthreshold, linearized models in favor of a MOS approach. The original mathematical underpinning for the former (Guess and Crump, 1976) was based on radioactive decay. Here, the concept arose that the decay of a single atom of, for example, plutonium in an organ such as the lung could lead to cancer. Thus, the line relating cancer incidence to dose was "forced" to pass through the origin, because even the smallest dose possible was potentially toxic. In this way, the dose associated with the theoretical risk of cancer of, for example, one in a million above control, could be determined. The debate about the precise shape of the cancer dose-response curve, particularly at low doses where exposure is most likely to occur, has been developed for radioactive and chemical carcinogens. A dedicated issue of the BELLE Newsletter [7(3), 1999] reviewed the question of whether radioactive and chemical carcinogens exhibit hormesis (a low dose of a carcinogen gives a lower cancer incidence than control) and/or adaptation (a low
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dose of a carcinogen protects against a high dose). The shape of the dose-response for cyanazine carcinogenicity (Table 33.3) close to the lowest dose (1 ppm) is difficult to assess. It is possible that hormesis or adaptation may be occurring. This would indicate that the LNT model is both inappropriate and overly conservative for dealing with low-dose cancer risk assessment. Several examples of hormesis have been described, for both radiation and chemically induced cancers. However, although several explanations have been proposed, including the stimulation of DNA repair by low doses of a chemical alkylating agent (Samson and Cairns, 1977) or radiation (Olivieri et aI., 1984; Wiencke et aI., 1986), it is unclear how to make use of the information in human cancer risk assessment. It is often felt that dispensing with the LNT model requires a more thorough understanding of mechanism( s) of hormesis and adaptation. Even so, it seems likely that the linearized cancer doseresponse curve may not be valid for extrapolation to low doses in most cases. Instead, a MOS approach may be more appropriate, although generally less conservative. The phenomenon of adaptation should also be considered because the mechanism(s) may be identical to those for hormesis. It should be noted that hormesis and adaptation induced by low-level exposure to radiation can be applicable to reduced risk from chemical carcinogens and vice versa (Cai, 1999). An area of research which could add substantially to the understanding of mechanisms of cancer production by pesticides is that of oncogenes and/or tumor suppressor genes (Purchase, 1994). Mutations occur at specific codons in these genes, to activate (or inactivate) them, when exposed to particular carcinogens. For example, in rat mammary gland tumors induced by a single administration of nitrosomethylurea, tumor cells from all (48) tumor-bearing rats contained the G to A mutation at position 2 of codon 12 of the Ha-ras-1 oncogene, which was quite different from DMBA-induced mammary tumors, which had mutations at positions 2 and 3 of codon 61 (Zarbl et aI., 1985). The proteins produced by oncogenes are often localized in the plasma membrane and have been demonstrated to be involved in regulating cell growth or cell division and proliferation. Cells containing such activated genes could thus have a selective growth advantage over normal cells, leading to tumor progression. It is hoped that future research into mutations in oncogenes and tumor suppressor genes following pesticide exposure will lead to a clearer picture emerging about the mechanisms of cancer induction and progression. For example, the linearized multi stage model (Global 86) is based on the premise that a carcinogen causes tumors by the same mechanism which causes tumors in control, undosed animals (Crump, 1996). The technology is now available to test this supposition, which could lead to improvements in risk assessment and thus in the regulation of carcinogenic pesticides. 33.5.2 OCCUPATIONAL EXPOSURE
Occupational exposure studies using Bladex formulations on cotton, the major-use crop in California, were not available
to DPR. A ground study using Bladex 4L on corn (preemergent) was considered to be a suitable alternative (Sanborn and Mehler, 1996). However, several possible sources of error may exist. For example in 1993, 223,3551b of active ingredient were applied in California to cotton as 4L (51%) and 216,080 lb as 90DF (49%). No calculations were made to estimate possible occupational exposure to the 90DF formulation, although exposure to 90DF could be quite different for two reasons. First, because 90DF is a solid, unlike the 4L formulation, which is a liquid. Second, because unlike for the 4L formulation, chemical-resistant gloves are not required on the label for MLAs using 90DF; only waterproof gloves are currently required. The estimation of long-term and lifetime exposure assume worst-case scenario assumptions regarding repeated exposure, based on the amortization of short-term exposure estimates. It is unlikely that an individual will use the same pesticide for 40 years. However, it is quite possible that an agricultural worker may be exposed to other chemicals with similar toxicological properties, that is, other chloro-triazines. Therefore, these two uncertainties about the duration of exposure could cancel out. The issue of using amortized average lifetime exposure estimates in cancer risk assessment, when exposure is intermittent rather than continuous, has been addressed in the literature. For example, it was concluded by Kodell et al. (1987) that cancer risk estimates based on lifetime average dose rates may be in error, based on whether the exposure is in early, middle, or late life and on whether the carcinogen is early, middle, or late acting. Whenever the exposure and tumor inductia are widely separated temporally, risk will be overestimated to a large degree (several orders of magnitude). In cases where the exposure and susceptibility to neoplas are in synchrony, cancer risk may be underestimated slightly, by :s k / r, where k is the number of stages of carcinogenesis and r is the number of these (stages) affected by the carcinogen. In the case of cyanazine (and other pesticides) it is unclear which time period is critical for exposure to result in cancer. The majority of the occupational exposure is via the dermal route. Few short-term and no long-term dermal toxicity studies were conducted and therefore it is necessary to estimate dermal absorption for addressing risk associated with occupational exposure. Human dermal penetration data were not available and the absorption was assumed to be 2%, the same as for the rat. This value may be an overestimate of dermal penetration because rates in rodents are generally greater than rates in humans (Feldmann and Maibach, 1974; Wester and Maibach, 1985). However, rat laboratory studies involve only a small area of skin, compared with the larger areas that are generally associated with human exposures. Because absorption tends to increase over a larger surface area of exposure (i.e., the rate and total amount of absorption are generally inversely proportional to the concentration of chemical) the rat data may underestimate human dermal absorption. Another assumption, which would tend to increase the occupational exposure estimates, was the use of a maximum
References
number of loads per day. On the other hand, a factor which would reduce occupational exposure was the use of a mean application rate (2 lb a.i. per acre) rather than the maximum label rate (4.5 lb a.i. per acre). The information on application rates was obtained from the California pesticide usage database, 1991-1993. Applications of cyanazine to cotton in California are largely made early postemergence, when application rates are lower than preemergent ones. This justifies the use of the lower application rate in the occupational exposure calculations. Because cotton is the major crop on which cyanazine is used in California, accounting for ~ 90% of the total pounds of active ingredient applied, the majority of occupational exposure to cyanazine will be from applications to cotton. At the time the worker exposure study was conducted, the label for cyanazine in California allowed an open system to be used by the mixer-loader-applicator and an open tractor cab to be used during application. However, 8 of the 12 data points pooled to derive the ADD value were obtained using a closed cab. The data indicated approximately a 3to 4-fold protection factor. Therefore the current ADD value would underestimate the "actual" ADD, based on label requirements at that time. The label for cyanazine was amended to require a closed cab, from January 1, 1998 (U.S. EPA, 1995). 33.5.3 DIETARY EXPOSURE Because cyanazine is used either preemergence or early postemergence, it is unlikely that residues will be found at harvest in any raw agricultural commodities. Therefore, the default residue values used for calculating possible dietary exposure are considered theoretical values which result in a "worstcase" situation. In practice, the actual MOS values for dietary exposure are thus likely to be considerably higher than those calculated. In addition, the residues in drinking water, which were used for calculating MOS values and excess cancer risk, were also default values at the LOD or 50% of LOD. It is unlikely that an individual will consume for a lifetime commodities which have been treated with cyanazine. Pesticide usage reports indicate, for example, that only 5-8% of California corn is treated with cyanazine and 18-20% in the 17 major corn production states. When the chronic dietary exposure values were adjusted using conservative estimates of percentage of crop treated, they were reduced to between 18 and 28% of the dietary exposure values using 100% of crop treated. The chronic MOS values and excess cancer risk were reduced correspondingly. 33.5.3.1 Drinking Water Cyanazine and other triazine herbicides have a long history of being detected in ground water and surface water in the mid-western states, for example, Illinois, Indiana, Kansas, Missouri, Nebraska, and Ohio. Triazines, such as simazine and
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atrazine, along with selected degradates, are also among the most frequently detected pesticides in California groundwater. However, cyanazine has not been found in groundwater in California. The current residue methods used by D.S. EPA and DPR do not identify specific cyanazine degradates. It has been reported that detections of parent cyanazine in mid-Western wells were only 50% as frequent as were detections of the cyanazine amide, a primary soil degradate (Kolpin et aI., 1996). In a DPR report (DPR, 1996) on groundwater testing results for 1994-1995, desisopropyl atrazine (DIPA) was the fourth most frequently detected pesticide (or degradate) in wells. This compound is a common degradate of both cyanazine and simazine, in addition to atrazine. Therefore, it is possible that some detections of DIPA resulted from cyanazine usage. The calculation of MOS and risk from drinking water exposure to DIPA would reduce the former and increase the latter.
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Murnik, M. R., and Nash, C. L. (1977). Mutagenicity of the triazine herbicides atrazine, cyanazine and simazine in Drosophila melanogaster. J. Toxicol. Environ. Health 3, 691-697. Nemec, M. D. (1987). "Two-Generation Reproduction Study of Technical Bladex Herbicide (SD 15418) in Rats." WIL-93001, WIL Research Laboratories Inc. DuPont SRO 15-87. Olivieri, G., Bodycote, J., and Wolff, S. (1984). Adaptive response of human Iymphocytes to low concentrations of radioactive thymidine. Science 223, 594-597. Perocco, P., Colacci, A, and Grilli, S. (1993). In vitro cytotoxic and cell transforming activities exerted by the pesticides cyanazine, dithianon, diflubenzuron, procymidone and vinclozolin on BALB/c3T3 cells. Environ. Mol. Mutagen. 21(1), 81-86. Peto, R., Pike, M. c., Day, N. E., Gray, R. G., Lee, P. N., Parish, S., Peto, J., Richards, S., and Wahrendorf, J. (1980). Guidelines for simple, sensitive significance tests for carcinogenic effects in long-term animal experiments. In "Long-Term and Short-Term Screening Assays for Carcinogens: A Critical Appraisal," IARC Monographs, Supplement 2. pp. 340-345. IARC, Lyon, France. Purchase, I. F. H. (1994). Current knowledge of mechanisms of carcinogenicity: Genotoxins versus non-genotoxins. Hum. Exp. Toxicol. 13, 17-28. Roloff, B., Belluck, D., and Meisner, L. (1992). Cytogenetic effects of cyanazine and metolachlor on human Iymphocytes exposed in vitro. Mutat. Res. 281, 295-298. Samson, L., and Cairns, J. (1977). A new pathway for DNA repair in Escherichia coli. Nature 267,281-283. Sanborn, J. R., and Mehler, L. (1996). "Assessment of Human Exposure to Cyanazine." Report No. HS-1526, Department of Pesticide Regulation, Worker Health and Safety Branch, California Environmental Protection Agency. Seilkop, S. K (1995). The effect of body weight on tumor incidence and carcinogenicity testing in B6C3F 1 mice and F344 rats. Fundam. Appl. Toxicol. 24,247-259. Simpkins, J. w., Eldridge, J. c., and Wetzel, L. T. (1998). Role of strain-specific reproductive patterns in the appearance of mammary tumors in atrazinetreated rats. In "Triazine Herbicides: Risk Assessment" (L. G. Ballantine, J. E. McFarland, and D. S. Hackett, eds.), ACS Symposium Series, Vol. 683, pp. 399-413. Am. Chem. Soc., Washington, DC. Stevens, J. T, Breckenridge, C. B., Wetzel, L. T, Gillis, J. H., and Luempert, L. G., Ill. (1994). Hypothesis for mammary tumorigenesis in SpragueDawley rats exposed to certain triazine herbicides. 1. Toxicol. Environ. Health 43, 139-153. Taets, C., Aref, S., and Rayburn, A. L. (1998). The clastogenic potential of triazine herbicide combinations found in potable water supplies. Environ. Health Perp. 106(4), 197-201. Tannenbaum, A (1948). Effects of varying caloric intake upon tumor incidence and tumor growth. Ann. New York Acad. Sci. 49, 5-17. TAS (l992a). Exposure 4. Detailed Distributional Dietary Exposure Analysis, version 3.2. Technical Assessment Systems, Inc., Washington, DC. TAS (I 992b). Exposure 1. Chronic Dietary Exposure Analysis, version 3.2. Technical Assessment Systems, Inc., Washington, DC. Tennant, M. K, Hill, D. S., Eldridge, J. C., Wetzel, L. T, Breckenridge, C. B., and Stevens, J. T (1994). Chloro-s-triazine antagonism of estrogen action: Limited interaction with estrogen receptor binding. J. Toxicol. Environ. Health 43,197-211. Travis, C. c., and White, R. K. (1988). Interspecies scaling of toxicity data. RiskAnal. 8,119-125. Turnbull, G. J., Lee, P. N., and Roe, F. J. C. (1985). Relationship of body-weight gain to longevity and to risk of development of nephropathy and neoplasia in Sprague-Dawley rats. Food Chem. Toxicol. 23, 355-361. U.S. Department of Agriculture (USDA) (1987-1988). "Data Set: NFCS 87-1-1 Nationwide Food Consumption Survey. 1987-1988." Preliminary report, unpublished, U.S. Department of Agriculture. U.S. Environmental Protection Agency (U.S. EPA) (1984). "Pesticide Assessment Guidelines, Subdivision F. Hazard Evaluation: Human and Domestic Animals." U.S. Environmental Protection Agency, Washington, DC.
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V.S. Environmental Protection Agency (V.S. EPA) (1985). "Cyanazine: Special Review Position Document I." OPP-30000/46, Office of Pesticide and Toxic Substances, Washington, DC. V.S. Environmental Protection Agency (V.S. EPA) (1986a). "Guidance for the Reregistration of Pesticide Products Containing Cyanazine as the Active Ingredient." PB86-175098, Office of Pesticide Programs, Washington, DC. V.S. Environmental Protection Agency (V.S. EPA) (1986b). "Cyanazine: Special Review Technical Support Document." Office of Pesticide Programs, Washington, DC. V.S. Environmental Protection Agency (V.S. EPA) (1986c). Guidelines for carcinogen risk assessment. Federal Register 51(185),33993-34012. V.S. Environmental Protection Agency (V.S. EPA) (1987). "Cyanazine: Intent to Cancel Registrations; Denial of Applications for Registration; Conclusion of Special Review." PB90-261595, Office of Pesticide Programs, Washington, DC. U.S. Environmental Protection Agency (V.S. EPA) (199Ia). "Peer Review of Cyanazine (B1adex)." May 21,1991. V.S. Environmental Protection Agency (V.S. EPA) (199Ib). "Cyanazine (188C), Atrazine (63) and Simazine (740) Quantitative Risk Assessment Comparisons on Malignant Mammary Tumors Only in Rats. Revised Comparisons as of July, 1991." V.S. Environmental Protection Agency (V.S. EPA) (1993). "Cyanazine; Quantitative Estimate of Carcinogenic Risk: Oral Slope Factor." June 14, 1993.
725
V.S. Environmental Protection Agency (V.S. EPA) (1994a). Atrazine, simazine and cyanazine: Notice of initiation of special review. Federal Register 59(225),60412-60443. V.S. Environmental Protection Agency (V.S. EPA) (1994b). Drinking Water Regulations and Health Advisories. EPA 822-R-94-001. V.S. Environmental Protection Agency (V.S. EPA) (1995). "Questions & Answers: Phase-out of Cyanazine." OPPTS 7506C. 8-2-95. Venkat, 1. A., Shami, S., Davis, K, Nayak, M., Plimmer, 1. R., Pfeil, R., and Nair, P. P. (1995). Relative genotoxic activities of pesticides evaluated by a modified SOS microplate assay. Environ. Mol. Mutagen. 25, 67-76. Walker, A. 1. T., Brown, V. K H., Kodama, 1. K, Thorpe, E., and Wilson, A. B. (1974). Toxicological studies with the 1,3,5-triazine herbicide cyanazine. Pestic. Sci. 5, 153-159. Wester, R. c., and Maibach, H. 1. (1985). In vivo percutaneous absorption and decontamination of pesticides in humans. 1. Toxicol. Environ. Health 16, 25-37. Wiencke, 1. K, Afzal, V., Olivieri, G., and Wolff, S. (1986). Evidence that the [3Hl thymidine-induced adaptive response of human lymphocytes to subsequent doses of x-rays involves the induction of a chromosomal repair mechanism. Mutagenesis I, 375-380. Wiles, R. et al. (1994). "Tap Water Blues." Environmental Working Group, Washington, DC. www.ewg.org Zarbl, H., Sukumar, S., Arthur, A. v., Dionisio, M.-Z., and Barbacid, M. (1985). Direct mutagenesis of Ha-ras-l oncogenes by N-nitroso-N-methylurea during initiation of mammary carcinogenesis in rats. Nature 315,382-385.
CHAPTER
34 Pesticides as Endocrine-Disrupting Chemicals* Robert J. Kavlock U.S. Environmental Protection Agency
34.1 INTRODUCTION The endocrine system consists of a number of central and peripheral organs (e.g., hypothalamus-pituitary, thyroid, parathyroid, adrenal, pancreas, ovaries, testes) that synthesize, store, and release hormones (e.g., thyroid hormone, parathyroid hormone, corticosterone, insulin, estrogen, progesterone, and testosterone) into the blood (Griffin and Ojeda, 1988; Hadley, 1996). These hormones, in turn, regulate the function of remote organs and tissues in the body to maintain homeostasis either by inducing (or suppressing) the synthesis of genes or by altering signal transduction pathways within the target cells. Most endocrine organs are linked with the hypothalamus and pituitary gland in classical negative feedback loops, which allow precise control of hormone levels in the blood. Frequently (especially during critical developmental periods) the levels of hormones are regulated within narrow limits by the feedback loops, but at other times (e.g., during the estrous cycle) large fluctuations in hormone levels are the intended result, as they interact with target tissues to either augment or repress the release of other hormones. Given the central role of the endocrine system in regulating homeostasis and controlling developmental processes, it is not surprising that interference with their normal action leads to alterations in either function or morphology. Indeed, pharmaceutical agents are often developed for such properties, be it the regulation of ovulation by birth control pills, the reduction of breast cancer risk by antiestrogens (e.g., tamoxifen), or the reduction of prostate growth by antiandrogens (e.g., finasteride). Although the ability of pesticides to interact with the endocrine function has been known for decades, it was only in the 1990s that interest in this mode of action rose to a high level of prominence. For example, the estrogenic action of some DDT analogues was first reported in 1952 (Fisher et aI., 1952), and a number of publications appeared in the 1970s demonstrat-
ing that the insecticides kepone (e.g., Gellert, 1978; Guzelian, 1982) and methoxychlor (e.g., Bulger et aI., 1978) were estrogenic. With the reports of reproductive tract cancers and other disorders noted in the offspring of women who received diethylstilbestrol (DES, a powerful synthetic estrogen) in the 1950s to help prevent miscarriages (Herbst et aI., 1972) and the followup work in animal models by McLachlan and co-workers from the NIEHS (Korach and McLachlan, 1985; McLachlan and Newbold, 1987), the phrase "environmental estrogens" became prevalent. The NIEHS subsequently organized several meetings beginning in the late 1970s (McLachlan, 1980; 1985; McLachlan and Korach, 1995) which were largely centered around the effects of diethylstilbestrol. Endocrine disruption as a main environmental issue can principally be traced to 1992. At that time, wildlife biologist Theo Colborn of the World Wildlife Fund organized a conference held at the Wingspread Conference Center in Wisconsin that examined a broad range of indicators of adverse health outcomes in humans and in wildlife (Colborn and Clements, 1992). The epidemiologists, toxicologists, and wildlife biologists who attended that meeting issued the following consensus statement that attempted to link a series of diverse observations with endocrine disruption as a common and powerful mode of action:
*This manuscript has been reviewed by the V.S. EPA and approved for publication. Mention of trade names or other commercial products does not constititute official endorsement. Handbook of Pesticide Toxicology Volume 1. Principles
727
We are certain of the following: A large number of man-made chemicals that have been released into the environment, as well as a few natural ones, have the potential to disrupt the endocrine system of animals, including humans. Among these are the persistent, bioaccumulative, organohalogen compounds that include some pesticides (fungicides, herbicides, and insecticides) and industrial chemicals, other synthetic products, and some metals. Many wildlife populations are already affected by these compounds. The impacts include thyroid dysfunction in birds and fish, decreased fertility in birds, fish, shellfish, and mammals; decreased hatching success in birds, fish, and turtles; gross birth deformities in birds, fish, and turtles; metabolic abnormalities in birds, fish, and mammals; behavioral abnormalities in birds; demasculinization and feminization of male fish, birds, and
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mammals; defeminization and masculinization of female fish and birds; and compromised immune systems in birds and mammals.
Since the publication of those proceedings, there has been an explosion of scientific symposia, workshops, and committee efforts to explore the linkage among environmental exposures, altered endocrine function, and adverse health effects. The identification of wildlife populations experiencing either adverse effects on individuals or on populations from areas contaminated with endocrine-disrupting chemicals, combined with the observation of declines in human health indices such as sperm quality and cancers of the endocrine-regulated organs such as the breast, testes, and prostate have further raised concerns (Cooper and Kavlock, 1997; Kavlock and Ankley, 1996). From 1995 to 2000 alone, more than two dozen major activities (Table 34.1) have explored the issue of endocrine disruption on a wide variety of aspects, including specific classes of animals, particular ecosystems, major adverse health outcomes, screening and testing procedures, and overall scientific assessments and research needs. Within the U.S. government, efforts have been directed at developing a national research agenda, establishing an inventory of ongoing research projects, and identifying specific research gaps (Reiter et aI., 1998). In 1996, nearly 400 projects funded by the U.S. government were identified that bore relevance to understanding the nature and magnitude of the problem. The concern about potential adverse effects of endocrine disruptors culminated in the enactment of two laws by the U.S. Congress and included the requirement for screening chemicals for estrogenic and other endocrine activity (i.e., the Food Quality Protection Act of 1996 and the Safe Drinking Water Act of 1996). As a result of this legislation, the U.S. Environmental Protection Agency (U.S. EPA) established the Endocrine Disruptor Screening and Testing Advisory Committee (EDSTAC) to assist in its implementation (see Section 34.7). The EDSTAC-recommended approach to evaluating potential endocrine-disrupting effects involving the estrogen, androgen, and thyroid hormone signaling pathways (U.S. EPA, 1998) is arguably the largest new testing program to be proposed. The Wingspread Conference (Colborn and Clements, 1992) appears to be the earliest use of the phrase "hormone disruptors" to represent the range of potential interactions discussed at that meeting. In 1995 at a workshop organized by the U.S. EPA (Kavlock et aI., 1996), the first formal definition of an "endocrine disruptor" was put forth to emphasize that more than estrogenicity was involved in the findings of adverse effects mediated by exogenous chemicals on the endocrine system (Table 34.2). This definition was subsequently criticized for not explicitly indicating that any effect had to occur as a result of primary interference with the endocrine system, and not requiring characterization of such action by in vivo studies. Subsequent definitions by the European Commission (1996) and the International Programme on Chemical Safety (International Programme on Chemical Safety, 1998) now overtly provide those distinctions. From those efforts, terms like "potential endocrine disruptors" arose to describe chemicals which
might, for example, be able to activate estrogen inducible reporter genes in an in vitro system, but which lack evidence of estrogenicity in vivo. In light of the complexity of the endocrine system and the multiple points at which it can be perturbed by exogenous agents, it is not surprising that endocrine disruption may be caused by a number of different chemical classes and structures, including a number of pesticides. One common underlying theme is that the reproductive system, particularly that of the developing organism, is especially vulnerable to the toxicity manifest by alterations in endocrine function; a number of pesticidal agents have been shown to exert their effects by this mode of action. Some pesticides, such as the insect growth regulators, are specifically developed for those abilities (although these do not appear to be endocrine disruptors in vertebrates), whereas for others the endocrine effects exhibited by target and nontarget organisms are clearly secondary to their primary mode of toxicity. In this chapter, four basic modes of action (Table 34.3) were used to classify endocrine-disrupting pesticides: (1) the ability to interact directly with steroid receptors; (2) the ability to modify steroid hormone metabolizing enzymes; (3) the ability to perturb hypothalamic-pituitary release of trophic hormones; and (4) as yet uncharacterized proximate modes of action. The following text provides an overview of the pesticides that act via these mechanisms and of the types of effects observed in experimental situations. It does not cover other aspects of their toxicity, as that is amply covered in other sections of this handbook. Impacts of altered endocrine function on development and reproduction, particularly in experimental animal models, will be emphasized, but other endocrine effects of pesticides are also covered where there is sufficient information, as are a few examples from wildlife-related studies. Studies using in vitro systems to detect modes of action will generally be mentioned only in conjunction with in vivo applications documenting that the mode of action is operable in an intact multicellular organism.
34.2 STEROID RECEPTOR LIGANDS 34.2.1 ESTROGENS 34.2.1.1 DDT
The estrogenic potential of some DDT analogues, in particular 2,2' -bis-(p-hydroxyphenyl)-l, 1, l-trichloroethane), was described in the ovariectomized rat nearly 50 years ago by Fisher et al. (1952) and confirmed later by Bitman et al. (1968) for the o,p'-DDT isomer. Numerous succeeding publications have shown that the o,p' isomer of DDT, which comprises approximately 20% of the technical grade product, is the active estrogenic moiety (as will be noted in Section 34.2.2.3, however, the persistent p,p'-DDE (1,I-dichloro2,2-bis( 4-chlorophenyl)ethylene) metabolite is also hormonally active, but in this instance the activity is that of an antiandrogen). The binding affinities of several DDT isomers (o,p' -DDT,
34.2 Steroid Receptor Ligands
729
Table 34.1 Major Reports of Workshops and Committee Reports from the 1990s Examining the Issue of Endocrine Disruption Year
Organization
Purpose
Reference
1992
World Wildlife Federation
Examine the commonalities of adverse effects in wildlife, experimental animals,
Colborn and Clements
(WWF) 1995
German Federal Environmental Agency Ministry of Environment and
and humans; produced the "Wingspread Consensus Statement" Discuss the occurrence and impact of substances that have an endocrinic effect, and the potential risks that may arise to humans and the environment
(1992) Umweltbundesamt (1995)
Investigate effects of estrogens on male reproductive development and function
Toppari et al. (1996)
Review existing literature for evidence of changes in human reproductive health and
Institute for Environment
Energy, Denmark UK Medical Research Council U.S. Environmental Protection Agency National Institutes of Environmental Health
effects in wildlife and to determine causal links; also to identify gaps in knowledge Meet the research needs for the carcinogenic, reproductive, immunologic, and neurologic
and Health (1995) Kavlock et al. (1996)
effects of endocrine disrupting chemicals Investigate the cell biology, developmental effects, sources, and health implications of environmental estrogens
McLachlan and Korach (1995)
Sciences U.S. Environmental Protection Agency 1996
European Commission
Develop a research strategy for assessing the ecological risk of endocrine
Ankley et al. (1997)
disruptors Assess the scope of the endocrine disruption problem in Europe, identify gaps in present knowledge and outstanding epidemiological questions; summarize current
European Commission (1996)
activities in Europe and identify research priorities 1997
U.S. Environmental Protection Agency U.S. EPA, CMA, WWF
Provide an overview of the current state of the science relative to environmental endocrine
U.S. EPA (I997a)
disruption in humans, laboratory testing, and wildlife species Evaluate technical merits and limitations of available in vivo and in vitro assays to
Gray et al. (I997a)
detect chemicals that act as (anti)estrogens and (anti)androgens, antithyroid agents SETAC/OECD
Assess the need for tests and procedures which can identify endocrine modulators;
Tattersfield et al. (1997)
critically evaluate available test methods; recommend appropriate hazard identifcation strategies OECD
Critically assess the ability of existing OECD-authorized test methods to detect
Organization of
a chemical sex-hormone-disrupting potential and also review a range of
Economic
nonregulatory model systems for suitability in research
Cooperation and Development (1997)
Arctic Monitoring and Assessment Program
Examine the levels of anthropogenic pollutants in the Arctic environment and assess their effects in all relevant compartments to aid in identification of pollutant assessments
Japan Chemical Industry Association 1998
European Commission
Scientifically evaluate the present situation of endocrine disruption; elaborate research needs for Japanese government; provide basis for scientific cooperation Exchange information on national and international research activities and plans; discuss research coordination; assist in development of global assessment
U.S. Committee on the Environment and Natural Resources SETAC
Develop a national planning framework for endocrine-disruptor research;
Arctic Monitoring and Assessment Programme (1997) Japan Chemical Industry Association (1997) European Commission (1998) Reiter et al. (1998)
analyze the existing federally funded research projects to help identify information gaps Develop a process through which a scientifically based risk assessment can be
Kendall et al. (1998)
conducted for endorcrine disruptors, including both acute and chronic exposures in wildlife U.S. EPA, CMA, WWF
Identification of methods for detecting the effects of (anti-)estrogenic and (anti)androgenic
Ankley et al. (1998)
effects in wildlife, particularly as relates to development and reproduction Japan Environment Agency
Summarize of scientific knowledge and specific approaches to the problem
Japan Environment
Swedish Environmental
Report on basis for differences in sex development between organisms, and the
Swedish Environmental
Agency (1998) Protection Agency
role of hormones; assess role of foreign chemicals with regard to
Protection Agency
endocrine-disrupting properties
(1998) (continues)
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CHAPTER 34
Pesticides as Endocrine-Disrupting Chemicals
Table 34.1 ( Continued) Year
Organization
Purpose
Reference
1999
V.S. EPA, CMA, WWF
Evaluate screening methods for chemicals that alter thyroid function and
DeVito et al. (1999)
homeostasis, including methods to detect receptor binding, alterations in thyroid hormone transport, and catabolism. Marine Mammal Commission 1999
National Research Council
Review of what is known and what needs to be learned about the effects of persistent
O'Shea et al. (1999)
ocean contaminants on marine mammals. Critical review of literature on hormon-related toxicants in the environment; the known
NRC (1999)
and suspected toxicologic mechanisms, significant uncertainties, and to recommend research, monitoring and testing priorities
European Commission
Review of existing literature and scientific opinion of evidence for chemically-induced endocrine disruption, with emphasis on European wildlife, EV testing strategy, and
European Commission (1999)
ecological risk assessment
p,p'-DDT, o,p'-DDE, p,p'-DDE, o,p'-TDE (1,1-dichloro-2,2bis(4-chlorophenyl)ethane), and p,p'-TDE) relative to 17fJestradiol for the ERa (ER = estrogen receptor) and ERfJ receptors are similar (generally < 0.01 compared to estradiol; Kuiper et aI., 1998). The particular susceptibility of the developing organism to hormonally active compounds was clearly evident from the work of Heinrichs et al. (1971) and Clement and Okey (1974). In these studies, either direct injections of o,p'-DDT to neonatal rats or administration via the diet to breeding pairs of rats resulted in conditions of persistent estrus, polycystic ovaries, and infertility as the offspring reached maturity. Injections of neonatal rat pups with doses of o,p'-DDT as low as 0.1 mg per pup on postnatal days 2-4 were effective in inducing persistent vaginal estrus and anovulation in adulthood. The uterine epithelium of adult females given higher (0.5 or 1 mg) injections as neonates consisted of stratified squamous epithelium. Following ovariectomy, females given 0.1 mg and higher failed to show compensatory increases in luteinizing hor-
mone (LH) and follicle stimulating hormone (FSH). Combined, these effects suggest that the hypothalamic-pituitary axis was androgenized. Neonatal male rats treated similarly had normal reproductive organ weights and motile sperm (Gellert et aI., 1974). With the exception of in vitro studies looking at ligandreceptor interaction and monitoring studies of environmental contaminants, there has been relatively little research in the last 10-15 years on the health effects of DDT itself. This is not true, however, for studies on the p,p'-DDE metabolite (see Section 34.2.2.3). Cumulative lifetime exposure to estrogen is a well-known risk factor for breast cancer in women. The estrogenic effect of o,p'-DDT, the long environmental persistence of some metabolites of DDT (particularly p,p'-DDE), and the association between exposure to some organochlorine compounds and the incidence of breast cancer, led to the hypothesis of a cause-andeffect relationship (Davis et aI., 1993). However, a subsequent study, involving a larger study cohort and the ability to measure
Table 34.2 Definitions of an Endocrine Disruptor An exogenous agent that interferes with the production, release, transport, metabolism, binding, action, or elimination of natural
Kavlock et al. (1996)
hormones in the body that are responsible for the maintenance of homeostasis and the regulation of developmental processes An exogenous agent that interferes with the synthesis, secretion, release, transport, binding, action, or elimination of natural
V.S. EPA (1997a)
hormones in the body that are responsible for the maintenance of homeostasis, reproduction, development, and/or behavior An exogenous substance that causes adverse effects in an intact organism, or its progeny, consequent to changes in endocrine function (A potential endocrine disruptor is a substance that possesses properties that might be expected to lead to endocrine
European Commission (1996)
disruption in an intact organism.) A compound that has the ability to alter the homeostatic status of hormones or their interactions with associated receptors;
Kendall et al. (1998)
modified as necessary by the terms "natural" or "anthropogenic" or "direct" "indirect"-versus acting compounds An exogenous chemical substance or mixture that alters the structure or function(s) of the endocrine system and causes adverse
V.S. EPA (1998)
effects at the level of the organism, its progeny, populations, or subpopulations of organisms, based on scientific principles, data, weight-of-evidence, and the precautionary principle An exogenous substance that alters function(s) of the endocrine system and consequently causes adverse health effects in an intact organism, or its progeny, or (sub )populations
International Programme on Chemical Safety (1998)
34.2 Steroid Receptor Ligands
731
Table 34.3 General Modes of Action within the Endocrine System for Various Functional Classes of Pesticides Discussed in This Chapter" Hypothalamic-pituitary-endocrine organ feedback Sex steroid receptor function
Mise. or uncharacterized
perturbation
Steroid
Estrogen
Androgen
synthesis
Thyroid
Leydig ceIl
endocrine modes
receptor
receptor
inhibitors
function
function
of action
Fungicides
Procymidone
Ketoconazole
EBDCs
ETU
EthyIenethiourea
VincIozolin
Tributyltin
Thiram
Etridiazole
FoIpet
Metam sodium
Fenbuconazole
Procymidone
Mancozeb
VincIozolin
RI5I885
Pentachloronitrobenzene Triadimefon Insecticides
o,p'-DDT
p,p'-DDE
Chlordimeform
Clofentezine
Boric acid
Kepone
HPTE
Amitraz
Ethofenprox
Methoxychlor
Methoxychlor
Cypermethrin
Fipronil
o,p'-DDT
pyrethrins
fJ-HCH Herbicides
Linuron
Molinate
Atrazine
Acetochlor
Linuron
Nitrofen
Amitrole Bromacil DCPA Pendimenthalin Prodiamine Proamide Terbutym Thiazopyr Trifluralin Fumigants
DBCP
Mise.
N-OBHD
Piperonyl butoxide Pyrimethanil
aNote that there is a strong reliance on evidence from in vivo studies for characterization of endocrine-disrupting modes of action in this assessment. HPTE = 1.1-dichloro-bis(4-hydroxyphenyl)ethene; fJ-HCH = fJ-hexachlorocycIohexane.
polychlorinated biphenyls (PCBs) and DDE in blood samples taken many years prior to the time of diagnosis of breast cancer, found no association between exposure and incidence (Krieger et al., 1994). Two exhaustive reviews also concluded that the existing evidence does not support a cause-and-effect relationship between exposure to organochlorine compounds and either breast cancer, uterine cancer, or endometriosis, although neither does it provide sufficient grounds to reject such an hypothesis (Adami et al., 1995; Ahlborg et al., 1995). A more recent 17year prospective study of breast cancer found no relationship between total DDT and breast cancer, but it did find a twofold increased risk of breast cancer for exposure to dieldrin (Hoyer et al., 1998). Given the early identification of DDT as an estrogenic compound and the dramatic example of avian eggshell thinning by organochlorines, including o,p-DDT and its stable metabolite, p,p'-DDE, first detected in the 1960s, it is interesting to note that eggshell thinning remains a poorly understood phenomenon at the biochemical level (Feyk and Giesy, 1998). In
part this has occurred because some of the more commonly used laboratory species (the domestic chicken and the bobwhite quail) do not display eggshell thinning following exposure. In addition, the mechanism of action may vary among species. Interestingly, potential mechanisms, which include premature termination of shell formation, premature oviposition, effects on the protein matrix of the shell, effects on the initiation sites of shell formation, and enhancement of shell growth inhibitors are related in various ways to alterations in Ca+2 homeostasis and not to estrogenicity. In this regard, it is important to remember that the banning of DDT by the EPA in 1972 was based largely on environmental persistence and concerns for the status of wildlife populations (see the chapter on DDT). Biochemical mechanisms for eggshell thinning that have been studied with limited success include inhibition of carbonic anhydrase activity, inhibition of a calcium-dependent ATPase, inhibition of a Ca+2 -Mg+ 2 -activated ATPase, inhibition of progesterone binding, and inhibition of prostaglandin synthesis. Interpretation of the findings has been limited by
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Pesticides as Endocrine-Disrupting Chemicals
the lack of in vivo confirmation of in vitro findings and by the inability to accurately reflect known structure-activity relationships in the potency to cause eggshell thinning. Thus, the mechanism(s) responsible for eggshell thinning remain obscure because of potentially multiple effects which may be at least in part species specific. This early example of the difficulties in ascertaining modes of action for a clearly demonstrable population-level effect foreshadows much of the current debate on the role of endocrine disruption in other adverse health outcomes. 34.2.1.2 Kepone (Chlordecone)
Kepone (decachlorooctahydro-1 ,2,4-metheno- 2H -cyclobuta (6d)pentalen-2-one) is a hydrated chlorinated polycyclic ketone insecticide synthesized by the dimerization of hexachlorocyclopentadiene (Huff and Gerstner, 1978). It and the closely related mirex are potent ant killers. Huber (1965) noted that dietary exposure of male and female mice to between 10 and 37.5 ppm kepone for one month prior to mating and then for a lOO-day trial period substantially reduced litter size and offspring survival to weaning. Subsequent studies found that nearly all females fed 40 ppm for 4 weeks were in constant estrus. Histological analysis of the ovary showed normal follicular development, but few corpora lutea. Although a direct estrogenic effect was not hypothesized by Huber, he noted that the hormonal system of the female was severely disturbed. The effects on fertility of female mice was confirmed by Good et al. (1965), who noted decreases in litter sizes and frequency of litters following dietary administration of 5 ppm kepone over a l20-day period. Reproduction was also impaired in F 1 offspring maintained on control diet after weaning. Kepone was shown to be estrogenic to the oviduct of the quail (McFarland and Lacy, 1969) and Palmitar and Mulvihill (1978) showed that kepone bound to the chick estrogen receptor and induced ovalbumin and conalbumin. It also raised serum progesterone levels. Following the 1977 incidence of exposed workers (and secondarily their families) at the Life Sciences Plant in Hopewell, Virginia, concern for the toxicological effects of kepone increased considerably. Guzelian (1982) reviewed the effects of kepone in the Hopewell plant and noted decreased libido in 7 of 28 exposed workers. Although only 8 had normal sperm counts, histological analysis of testicular biopsies found arrested sperm maturation. In 12 of 13 cases from which repeated sperm measures were obtained, an increase in motile sperm in inverse proportion to the blood levels of kepone was observed. These effects were considered consistent with the toxicology literature that indicated an estrogenic effect of kepone, albeit at a potency about 10,000 times less than estradiol. In reviewing the literature up to that point, Huff and Gerstner (1978) noted reports of reductions in the germinal epithelium and number of spermatozoa in the testes of exposed quail, testicular atrophy in rats, and hepatocellular carcinomas in both sexes of rats and mice, but the endocrine effects were not emphasized in the toxicological profile. Likewise, Larson et al. (1979) noted testicular
atrophy in rats fed 25 ppm kepone for 3 months, without comment as to the mode of action. Reversible impairment in the fertility of female rats feed 25 ppm kepone for 3 months was noted by Cannon and Kimbrough (1979), who thought that perhaps the reproductive effects were secondary to alterations in adrenal function. In this study, fertility of treated males was not affected, but testicular histology was not an end point. Consistent with the now well-recognized sensitivity of the developing reproductive tract to endocrine disruption, Eroschenko and Mouse (1979) demonstrated that neonatal exposure of female mice to between 0.015 and 0.125 mg/day for up to 10 days resulted in keratinization of the vaginal epithelium and hypertrophy, hyperplasia, and glandular formation in the uterus. These changes were identical to those observed with estradiol. In a subsequent review of the literature, Eroschenko (1981) clearly demonstrated that the immature female reproductive tract was a target for kepone and that the effects appeared very similar to those noted after sustained estradiol treatment. Gray (1982) extended the effects of neonatal treatment beyond the reproductive tract by noting that exposing hamsters to 0.25-1.0 mg/day at 2 or 4 days of age altered the process of behavioral sex differentiation, with the females becoming masculinized and showing abnormal bisexual behavior. These findings indicate that kepone acts on the central nervous system during the critical period of sex differentiation to masculinize the brain the same way that estrogen does (testosterone is aromatized to estrogen within the neurons) in the brain. Uphouse (1985a, b) subsequently confirmed the weak estrogenicity of kepone (50 mglkg by intraperitoneal injection) in the adult female rat by noting that it induced persistent vaginal estrus in ovariectomized and intact animals, although it failed to mimic estrogen in priming the ovariectomized female for behavioral receptivity. When exposure occurred on the morning of proestrus, mating behavior was reduced at 50 and 75 mglkg, but most females had sperm in the vaginal smear following overnight breeding, and most females at these dose levels delivered offspring (Uphouse, 1985b). Adult mice given 8 mglkg kepone by oral gavage 5 days per week for 6 weeks displayed persistent vaginal estrus within the first two weeks and had significantly lowered ovulatory response to exogenous gonadotropins at 4 and 6 weeks of treatment compared to either controls or 17,B-estradiol treated females, suggesting a direct effect on the ovary in addition to the estrogenic effect (Swartz et aI., 1988). Administration of 30 mglkg chlordecone to ovariectomized immature rats increased the uterotrophic response between 0.1 and 1 !1g, but not following 10 !1g, estradiol benzoate; likewise the uterotrophic response of 15 and 30, but not 45 mglkg kepone was enhanced by 0.1 !1g estradiol benzoate (Johnson, 1996). In the adult male rat, Cochran and Wiedow (1984) failed to show effects on the reproductive tract following administration of chlordecone at sublethal doses either by subdermal implants (600 !1g/day) or by intraperitoneal injections (2.5 mg/day three times per week), whereas l7,B-estradiol (3.6 !1g/day) or testosterone (135 !1g/day) caused reductions in spermatogenesis, sperm motility, and weights of the testes and the epididymides. Dietary exposure of adult male rats to chlordecone
34.2 Steroid Receptor Ligands
for 90 days resulted in reversible decreases in epididymal sperm motility and viability at 15 and 30 ppm, but had no effect on testes weight, testes histology, or fertility (Linder et aI., 1983). In light of the estrogenic effects of kepone and the known role of endogenous estrogens on male brain development, efforts to characterize the long-term effects in perinatal exposure grew in prominence in the 1980s. Exposure of Fischer 344 rats throughout gestation and the first 12 days of postnatal life to 1 or 6 ppm kepone via the maternal diet did not affect maternal weight gains, birth weights, litter sizes, or sex ratios, forelimb or hindlimb grip strength, spontaneous motor activity, acoustic startle response, or tail flick latencies to thermal stimulation, but they did result in hypersensitivity to the motility-increasing effects of apomorphine in male rats at 100 days of age (Squibb and Tilson, 1982). Injection of female Sprague-Dawley rats with 5 or 10 mg/kg kepone from day 18 of gestation to postnatal day 7 resulted in precocial vaginal opening in the offspring, but did not alter neurological development as measured by open field activity, righting response, or eye opening, nor did it alter the volume of the sexually dimorphic medial preoptic nucleus of the brain (Cooper et aI., 1985). Impairment of adrenal function following early life stage exposure has also been noted. Injection of male and female Fischer 344 rats on postnatal day 4 (1 mg per pup) depressed basal and adrenal corticosterone levels in adult males (Rosencrans et aI., 1985). A holistic interpretation of the effects of kepone on the developing fetus, adult male, and adult female has yet to emerge. Clearly some of the effects are directly related to its estrogenic potential, but others, such as the effects on spermatogenesis and long-term effects on the estrous cycle, suggest that other modes of action, potentially on neurotransmitter levels, may be operable.
34.2.1.3 Methoxcyhlor Methoxychlor is undoubtedly the most well characterized pesticidal estrogen, possibly because unlike other well-known environmental estrogens such as o,p'-DDT and kepone, it is still registered for use in the United States. As early as the mid-1970s it was recognized that methoxychlor was activated by metabolism to an estrogenic chemical (Nelson et aI., 1978). Kupfer and Bulger (1987) reported that methoxychlor and MDDE (MDDE = 1,I-dichloro-2,2-bis(4-metoxyphenyl) ethane), an olefinic derivative, are proestrogens, and that monohydroxymethoxychlor and monohydroxy-MDDE are estrogenic both in vitro and in vivo. The bis-hydroxy metabolites were even more potent in vitro. Methoxychlor binds with equal affinity to ERa and ER,8 with a binding affinity relative to 17,8estradiol of < 0.01 (Kuiper et aI., 1998). In addition to the estrogenic potential of methoxychlor and its metabolite, Maness et al. (1998) demonstrated that both methoxychlor and HPTE (HPTE = 2,2-(bis-(p-hydroxyphenyl)-I, 1, I-trichloroethane)) in the range of 10- 8 to 10- 4 M reduced the activity of coadministered dihydrotestosterone, but showed no agonist activity. The metabolite was approximately 10-fold more potent than the parent compound as an antagonist. However, there is no litera-
733
ture at present confirming that this activity is sufficient to induce anti androgenic effects in vivo. Cummings (1997) reviewed the reproductive toxicity literature on methoxychlor, with particular emphasis on the female. Effects on fertility, early pregnancy, and in utero development of females, as well as adverse effects on adult males such as altered social behavior, following prenatal exposure were reported. For example, exposure of female rats on days 1-8 of gestation with doses of methoxychlor of 200 mg/kg resulted in a reduction in implantation sites due to the acceleration of embryo transport to the uterus from the estrogen-stimulated oviduct. This same dose level given for 8 days to adult female rats inhibited the decidualization response, an estrogenand progesterone-dependent response of the uterus that mimics the growth of the endometrium that occurs during pregnancy. A number of estrogen-stimulated biomarkers, including peroxidase, creatine kinase, and epidermal growth factor receptor, have also been observed in the uterus of females receiving methoxychlor. In the male, exposure to 100 or 200 mg/kg methoxychlor by oral gavage for 70 days damaged Sertoli cells and induced degenerative changes in the spermatogonia and spermatocytes, with some seminiferous tubules devoid of all cellular elements except spermatogonia (Ba!, 1984). No endocrine measures were recorded, but it was hypothesized that the estrogenic properties of methoxychlor might be involved with the disruption of spermatogenesis. Using a shorter duration exposure, Linder et al. (1992) administered either 4000 mg/kg for 1 day or 2000 mg/kg for 4 days, and reported degenerating cells in Stage VII seminiferous tubules 2 days after the acute exposure, and similar changes plus sliverlike remnants of condensed spermatid nuclei in Stages VIII-XIV and testicular debris in the caput. Again, no hormonal measures were recorded. Long-term (lO-month) exposure of weanling male rats to methoxychlor at levels between 200 and 400 mg/kg/day delayed puberty by as much as 10 days and reduced fertility and copulatory plug formation, sperm counts, and time to pregnancy (Gray et aI., 1999a). Unlike what was observed for 17,8-estradiol-implanted rats, no effects of methoxychlor exposure were noted on pituitary weight or on serum LH or prolactin, indicating that the central effects of methoxychlor do not resemble those of endogenous estrogen. The only developmental effect of methoxychlor noted in a standard teratology study in which females were exposed on days 6-15 of gestation to doses of methoxychlor between 100 and 400 mg/kg was an increase in wavy ribs at all dose levels (Khera et aI., 1978). However, using other approaches, the finding of heightened sensitivity of the developing organism to estrogens has been confirmed for methoxychlor. For example, mice given 300 mg/kg by oral gavage on gestation days 6-15 were unable to maintain pregnancy, while those given 200 mg/kg had prolonged pregnancies and the offspring had an increased percentage of atretic follicles (Schwartz and Corkern, 1992). Gray et al. (1989) exposed rats continuously from weaning through puberty and gestation to day 15 of lactation with 25, 50, 100, or 200 mg/kg/day methoxychlor. Treated females displayed accelerated age at vaginal opening
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Pesticides as Endocrine-Disrupting Chemicals
and first estrus at all dose levels. Cyclicity was accelerated at 25 mg/kg/day, normal at 50 and 100 mg/kg/day, and abolished at 200 mg/kg/day. An increase in cornified vaginal smears was present at 50 mg/kg/day and above. Puberty (preputial separation) was delayed in males at the two highest dose levels, whereas growth was reduced at all dose levels. Sex accessory gland weights and caudal sperm counts were reduced at doses as low as 50 mg/kg/day as were pituitary weights at 100 mg/k/day and above. The alterations in pubertal development of both sexes demonstrate the sensitivity of the developing organism to estrogenic substances. These findings were extended by Chapin et al. (1997), who exposed female rats to 0, 5, 50, or 150 mg/kg/day for the week before and after parturition, and then directly dosed the offspring until postnatal day 21. Dose-dependent amounts of methoxychlor and metabolites were present in milk and in the plasma of dams and pups. In the high-dose females, levels of methoxychlor and its monohydroxy metabolite reached 466 and 1004 ng/ml in the plasma and milk, respectively. Litter size was reduced at the high dose by 17%. Vaginal opening was accelerated in all dose groups. Preputial gland separation was delayed at the two high exposure levels. Adult estrous cyclicity was also altered and fewer ova released at 50 mg/kg and above; all groups of females showed uterine dysplasias, less mammary alveolar development, and reduced estrous FSH levels. Relatively minor effects were found on immune (decreased antibody plaque-forming response in males) and neurological (high-dose males were more excitable) function. Collectively these studies show both central and peripheral effects of developmental exposure to estrogenic chemicals, and reinforce the concept of sensitivity of the developing female in particular to such exposures.
34.2.1.4 Lindane Following reports of lindane-induced estrogen-mediated effects on female sex behavior in the rat (Uphouse and Williams, 1989), Cooper et al. (1989), and Laws et al. (1994) explored the interaction of y-hexachlorocyclohexane (y-HCH) with the estrogen receptor of female Sprague-Daw ley rats. In the immature (21-day-old) female, administration of 40 mg/kg y-HCH for 7 days significantly attenuated the estrogen-dependentincrease in uterine weight period. Dose-dependent delays in vaginal opening and onset of ovarian cyclicity were also observed. However, it did not alter the concentration of the estrogen receptor in the nuclei of uterine cells 30 hours after administration of y-HCH alone, or in combination with estradiol, compared to the relevant controls. In addition, combination with estradiol did not alter the induction of progesterone receptors in the uterus of the immature female, nor did it modify estradiol levels in the serum of estradiol-treated ovariectomized adult females. Thus, y-HCH does not appear to exert an estrogenic effect in the intact rat. The effects of premating, pregnancy, and lactational exposure to y-HCH (1 mg/kg/day) on reproduction in the ewe was evaluated by Beard and Rawlings (1999). Although pregnancy rate was significantly decreased by lindane, there were
no effects on serum leuteinizing hormone, follicle stimulating hormone, thyroxin, or cortisol either under basal conditions or following administration of gonadtropin-releasing hormone, thyroid-stimulating hormone (TSH), and adrenocorticotropin. When mink were exposed to 1 mg/kg/day lindane from the time they were weaned for three generations (Beard and Rawlings, 1998), nO overt signs of toxicity were noted. However, lindane treatment reduced the proportion of mated mink that subsequently whelped, and the litter size of mink that did whelp. Testes size was reduced in third-generation males. No effects were noted on serum concentrations of cortisol, testosterone, or estradiol during the study. ,B-HCH was shown to be uterotrophic in ovariectomized adult mice following three daily doses, and sufficient levels had been stored in body tissues such that a 2-day fast 14 days after the ,B-HCH exposure also increased water inhibition by the uterus (Bigsby et aI., 1997). This response may have at least partially resulted from continual exposure of the uterus to the chemical during the 14 days, but similar potential release from body stores was not seen with o,p'-DDT. Using a human breast cancer cell line, ,B-HCH stimulated proliferation of ER positive cell lines MCF-7 and T47D, but not ER negative lines MDAMB231, MDA-MB468, and HS578T (Steinmetz et aI., 1996). The estrogen-inducible protein pS2 was increased in ,B-HCHtreated MCF-7 cells, a response that could not be inhibited by the antiestrogen ICI-64384. However, it did not displace estradiol in a competitive binding assay, nor was nuclear retention of ER altered. Another experiment showed that ,B-HCH did not activate a luciferase reporter gene containing a minimal promotor and a COnsenSUS estrogen response, whereas it did in the presence of a 2500-base-pair promotor from the prolactin gene. Thus ,B-HCH appears to induce estrogenic responses via a complex promotor, and this suggests that they are not mediated through the classical estrogen-receptor signaling pathway.
34.2.2 ANTIANDROGENS 34.2.2.1 Linuron Linuron is a chlorinated urea-based herbicide with structural similarity to the nonsteroidal anti androgen ftutamide. It induces Leydig cell adenomas in male rats in chronic bioassays although it lacks genotoxic action in a number of in vitro assays. Cook et al. (1993) exposed adult male rats to 200 mg/kg linuron for 2 weeks to study effects on sex accessory gland weights and the function of the hypothalamic-pituitary axis. Linuron decreased accessory sex gland weights in sexually immature rats and adult treated rats. Increased estradiol and LH levels were seen in adult treated males. These effects were consistent with the effects of ftutamide, although linuron did not elevate serum testosterone as did ftutamide. Linuron also competed with eH]testosterone for binding to the androgen receptor. The binding affinity was approximately 3.5 times less potent than ftutamide in this study. The antiandrogenic effects of developmental exposure to linuron have recently been described by Gray et al. (1999c) (see Section 34.2.2.5).
34.2 Steroid Receptor Ligands
34.2.2.2 Dicarboximide Fungicides (Vinclozolin and Procymidone) Although antiandrogens alter adult male reproductive function, the true impact of their toxicity is not observed until exposures occur encompassing the critical developmental periods when androgens play crucial roles in the differentiation of the reproductive tract and other tissues. One of the first demonstrations of a pesticide displaying antiandrogenic effects following developmental exposures was reported by Gray et al. (1994) with the dicarboximide fungicide vinclozolin. Exposure of rats to 100 or 200 mg/kg by oral gavage from gestation day 14 to postnatal day 3 yielded marked demasculinizing effects on male offspring. In both dose groups, male anogenital distance at birth was female-like, and prominent nipple development was evident at 2 weeks of age. As adults, treated male offspring were unable to attain intromission due to cleft phallus with hypospadias; mounting behaviors were normal. Other abnormalities observed included suprainguinal ectopic testes, vaginal pouches, epididymal granulomas, and small to absent sex accessory glands. The only change noted in female offspring was a reduced anogenital distance during the neonatal period. The phenotypic appearance in males is consistent with inhibition of both testosterone-dependent (Wolffian duct differentiation) and dihydrotestosterone-dependent (urogenital sinus and external genitalia) tissues, as expected of an androgen receptor antagonist. This activity was subsequently confirmed by Ke1ce et al. (1994), who reported that neither vinclozolin nor two principle metabolites [designated Ml (2-[[(3,5-dichlorophenyl)-carbamoyl]oxy]-2-methyl3-butenoic acid) and M2 (3',5'-dichloro-2-hydroxy-2-methylbut-3-enanilide)] inhibited 5-reductase activity, but that the metabolites (particularly M2) were able to competitively inhibit binding of [3H]RI881 to the androgen receptor and block androgen-dependent gene activation. Peripubertal exposure of male rats to vinclozolin at doses of 30 and 100 mg/kg/day retarded sex accessory gland and epididymal growth, but no effects on testes weight or sperm maturation were observed in adulthood (Monosson et al., 1999). A lower dose (10 mg/kg/day) induced significant increases in serum LH and testosterone concentrations. Analysis of serum levels ofMl and M2 suggested that these effects occurred in conjunction with only a low percentage of androgen receptors being occupied, as the levels were below the Ki values from in vitro binding assays. In studies situated following observation of Leydig's cell tumors in exposed rats procymidone, a related dicarboximide fungicide, was also shown to bind to the rat and mouse androgen receptor in a study triggered by the observation of hypergonadotropism after 2 weeks of dietary exposure (Hosokawa et al., 1993). Pituitary LH levels were increased after 2 weeks of exposure to 700 ppm in the rat, and 5000 ppm in the mouse. Smaller, nonsignificant increases in serum testosterone and LH were noted in both species at the higher exposure concentrations. Scatchard analysis of rat and mouse prostate androgenreceptor binding showed that procymidone had less than 0.07%
735
of the binding affinity of dihydrotestosterone. This affinity was similar to that of fiutamide and is sufficient to produce the same spectrum of phenotypes as seen in vinclozolin-exposed male offspring and to inhibit dihydrotestosterone-induced transcriptional activity in CV-l cells cotransfected with the human androgen receptor and a luciferase reporter gene (Gray et al., 1999b; Ostby et al., 1999). In vivo, procymidone appeared to have approximately half the potency of vinclozolin. 34.2.2.3 p,p' -DDE
Concern for anti androgenic effects of pesticides was considerably broadened with the report of Ke1ce et al. (1995) that p,p'-DDE, the persistent metabolite of DDT, inhibited androgen binding to the androgen receptor, androgen-induced transcriptional activity, and androgen action in developing, pubertal, and adult male rats. The inhibitor concentrations necessary for 50% displacement (lCsos) of eH]R1881 from the androgen receptor were 75, 5, 95, and 90 IJ.M for p,p'-DDT, p,p'-DDE, o,p'-DDT, and p,p'-DDD, respectively. Of these, only o,p'-DDT showed appreciable affinity for the estrogen receptor (ICso = 5 IJ.M). When rats were treated with 100 mg/kg/day o,p'-DDE on gestation days 14-18, male offspring displayed reduced anogenital distance and retained nipples at 13 days of age. Exposure of weanling (21-day-old) males to 100 mg/kg/day through postnatal day 57 delayed the onset of puberty by 5 days. Adult males given 200 mg/kg/day for 4 days had significantly reduced seminal vesicle and ventral prostate weights. Messenger RNA levels for an androgen-repressed protein (TRPM-2) in the ventral prostate was elevated by p,p'DDE treatment, whereas that of an androgen-induced protein (C3) was elevated. Assessment of the in vitro antiandrogenic effects of the DDT isomers has been confirmed in HepG2 cells by Maness et al. (1998), who noted that p,p'-DDE, p,p'-DDT, p,p'-DDD, and o,p'-DDT were antiandrogenic. The ICso for p,p'-DDE, the most potent isomer, was 1.86 IJ.M. The antiandrogenic effects of prenatal p,p'-DDE exposure were further studied by You et al. (1998), who observed that treatment of either Sprague-Dawley or Long-Evans rats with 100 mg/kg/day on gestation days 14-18 induced a reduction in male anogenital distance, an increase in the retention of male thoracic nipples, and alterations in the expression of the androgen receptor. Although tissue and body fluid concentrations of p,p'-DDE were similar in the two strains, there were indications that the LongEvans was more sensitive to the anti androgenic effects. 34.2.2.4 Pyrethroids Pyrethroids have been reported to bind to the androgen receptor with Km values in the micromolar range. Cypemethrin was given to pregnant rats by subcutaneous injections at dose levels between 0.25 and 25 mg/kg/day for the last 7 days of gestation, and pups continued to be dosed postnatally until 30 days of age. Anogenital distance was reduced at birth and at 85 days of age in males. There was a dose-response decrease in relative prostate weight at 55, but not 85, days of age, and there were no
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Pesticides as Endocrine-Disrupting Chemicals
effects on epididymal sperm counts at that time (Ronis et aI., 1995). These results suggest a weak antiandrogenic effect.
34.2.2.5 Comparative Aspects The phenotypes induced by prenatal exposure of male offspring to a number of antiandrogenic pesticides and toxic substances have been analyzed using a "pseudohermaphrodism index" (PHI) which measures the degree of feminization recorded on a number of testosterone (seminal vesicle, epididymal and testicular weights) and dihydrotestosterone (anogenital distance, nipple and areolas retention, ventral prostate weights, and incidence of hypospadias and vaginal pouch) sensitive tissues (Gray et aI., 1999c). Pesticides studied included vinclozolin, procymidone, chlozolinate, iprodione, p,p' -DDE, linuron, and ketoconazole following exposure on gestation day 14 to postnatal day 3. Male pups exposed to 100 mg/kg/day vinclozolin were about 90% female-like in their dihydrotestosterone (DHT) phenotype compared to 25% female-like in the measures of T-sensitivity. Procymidone and p,p' -DDE produced similar profiles of effect as vinclozolin, but with lower potencies. However, the related dicarboximide fungicides chlozolinate (100 mg/kg/day) and iprodione (100 mg/kg/day) were ineffective in altering the phenotype of male offspring. In contrast to DHT-end-point effective dicarboximides, the toxic substances dibutylphthalate and diethylhexylphthalate, produced higher incidences of testicular and epididymal effects than anticipated based on the overall PHI. Linuron (100 mg/kg/day) yielded somewhat unexpected results. While DHT-sensitive tissues were affected as expected by a low-affinity androgenreceptor antagonist, more pronounced effects were seen on the epididymis and testes, suggesting some additional mechanism of action beyond androgen receptor blockade may be operational. Treatment with ketoconazole (12-50 mg/kg/day) delayed parturition by as much as 3 days, consistent with effects on inhibition of steroid synthesis, but apparently fetal testosterone synthesis was not impaired, as offspring phenotype was normal in treated litters.
34.3 INHIBITORS OF STEROID SYNTHESIS 34.3.1 KETOCONAZOLE Ketoconazole, an imidazole antifungal agent first introduced in 1981, is a widely administered oral treatment for systemic mycoses (Fromtling, 1988). Its antifungal action is due to its inhibition of the synthesis of ergosterol via the P-450-dependent enzyme, 14-demethylase (Como and Dismukes, 1994). Ketoconazole is also known to inhibit P-450 enzymes of the steroidogenesis system, resulting in adverse endocrine effects in humans (Como and Dismukes, 1994). In clinical studies, ketoconazole was shown to inhibit both adrenal and testicular steroidogenesis (Pont et aI., 1982a, 1982b) with short-term decreases in serum androstenedione and testosterone following a single oral dose (DeCoster et aI., 1985). Discovery of
these endocrine effects sparked a flurry of interest in other potential clinical applications for this drug. Ketoconazole has been successfully used as a treatment for Cushing's syndrome and prostate cancer to decrease steroid hormone production (Sonino, 1987). Numerous animal studies have been conducted to explore its potential use as a male contraceptive. WaIler et al. (1990) evaluated the effects ofketoconazole on male rat fertility following three consecutive daily oral doses of either 200 or 400 mg/kg/day. Ketoconazole at a dose of 200 mg/kg/day significantly reduced fertility compared to control animals and resulted in a complete loss of fertility at a dose of 400 mg/kg/day. Sperm motility was reduced at the high dose and forward progression was reduced at both doses. In a similar study with mice, Joshi et al. (1994) also found a significant decline in sperm motility as well as reductions in sperm density at an oral dose of 400 mg/kg administered for a period of 60 days. Fertility in these mice was greatly reduced compared to that of controls. Research has also shown adverse effects on female reproduction. In a study of the effect of ketoconazole on early pregnancy, Cummings et al. (1997) treated rats with 10-100 mg/kg ketoconazole on days 1-8 of pregnancy. Evaluations at gestational day 9 showed a significant reduction in the number of implantation sites and serum progesterone levels as well as increases in uterine body weight. Further test results from pseudopregnant, ovariectomized rats and in vitro ovary culture indicate that ketoconazole directly interferes with uterine function by inhibiting ovarian steroidogenesis. This study confirms earlier research by Buttar et al. (1989), who found intrauterine growth retardation, delayed parturition, and postnatal developmental effects such as late descent of testes and vaginal opening in both rats and mice.
34.3.2 MOLINATE Molinate is a thiocarbamate herbicide that has been shown to reduce serum testosterone levels with resulting testicular toxicity (delayed release of spermatids) and impaired fertility in exposed male rats (Minor et aI., 1984). In a time course study, Sprague-Dawley rats received a single exposure to 100-400 mg/kg of molinate or 55-200 mg/kg molinate sulfoxide (a major metabolite found in rats) by intraperitoneal injection and were followed for up to 3 weeks. Testicular damage was dose and time dependent following molinate exposure. Histopathological changes (Sertoli cell vacuolation, failed spermiation, and phagocytosis of spermatids at Stages X and XI of spermatogenesis), were evident at 2 days after 400 mg/kg, and 1 week after 200 mg/kg. With additional time, the lesion progressed until germ cells were virtually absent from the seminiferous tubule. Similar effects were observed with lower doses of the sulfoxide (Jewell et aI., 1998). Additional experiments using 14C-labeled molinate, molinate sulfoxide, and molinate sulfone found extensive and tight binding to a protein of 180 kDa, subsequently identified as Hydrolase A, a carboxylesterase present in liver and testis (Jewell and Miller, 1998). They hypoth-
34.4 Hypothalamic-Pituitary Feedback Loops
esized that inhibition of the esterase could alter the mobilization of cholesterol esters from high-density lipoproteins, thus affecting testosterone biosynthesis. It was subsequently demonstrated that administration of molinate to rats (40-140 mg/kg/day for 7 days) caused a marked decrease in serum and testicular testosterone. In addition, 3H-molinate accumulated in the Leydig cells, and esterase activity in those cells was inhibited. In vitro, molinate sulfonate, and molinate sulfone, but not molinate, were potent inhibitors of the esterase activity in Leydig cells (Ellis et aI., 1998). In a risk assessment of molinate, it was noted that testicular toxicity has not been seen in exposed primates, and epidemiological studies in exposed workers showed no effect, although limitations of those studies did not preclude potential risks to human reproduction (Cochran et aI., 1997). In another review, it was noted that spermatoxicity was not seen in molinateexposed rabbits, dogs, or monkeys, whereas it was in mice and rats (Wickramaratne et aI., 1998). The relative order of sulfur oxidation as measured by analysis of urinary metabolites was reported as dog > rat > mouse '"'-' monkey » rabbit > human. For thiocarbamate cleavage, the rank order was rat > dog > mouse » rabbit > monkey (no data were available for humans). Wickramaratne et al. (1998) argued that the metabolic differences, combined with the unique role of high-density lipoproteins in cholesterol mobilization in rodents (which is inhibited by metabolites of molinate) as opposed to other mammals which rely on low-density lipoproteins (whose esterase, acetyl-CoA, is not inhibited by molinate metabolites) as their primary source of cholesterol, suggest that the rodent data on testicular toxicity is not relevant to humans. 34.3.3 DIBROMOCHLOROPROPANE The nemacide dibromochloropropane (DBCP) reached notoriety in 1977 when it was discovered that workers in a pesticide manufacturing plant had become oligospermic. Only 7 of 26 DBCP-exposed workers had normal sperm counts 11 years later (Lahdetie, 1995). Although the effects of DBCP on the adult testes appear to be related to direct effects on spermatogenic cells, most likely through damage to DNA (Lag et aI., 1989), there is some evidence that the mode of action on the developing fetus might have a endocrine basis. Determining which is a cause and which is an effect in such situations, however, is complicated. Warren et al. (1988) exposed pregnant rats to 25 mg/kg DBCP on days 14.5-19.5, 16.5-19.5, or 18.5-19.5 of gestation. Treatment for 6 days reduced intratesticular testosterone concentrations on day 20.5 by 50%. In adulthood, all exposure durations reduced male body weights, whereas testes weights were reduced 75% by 2 days of exposure and 90% following 4 or 6 days of exposure. In the brain, the volume of the normally sexually dimorphic nucleus of the preoptic area in males treated for 6 days in utero was not different from that of control females, and these males all displayed female lordosis behavior. A few of the males treated for 4 or 6 days had testes lacking seminiferous tubules.
737
34.3.4 TRIAZOLES Inhibition of aromatase activity, the enzyme which converts testosterone to estrogen, in the ovarian granulosa cell by the antifungal triazole 1, I-di -( 4-ftuorophenyl)-2-(1 ,2,4-triazoll-yl)-ethanol (RI51885) has been linked to a blockade of ovulation, particularly when given during diestrus I or 11 of the estrous cycle (Middleton et aI., 1986; Milne et aI., 1987). Doses as low as 5 mg/kg by gavage completely suppressed ovulation. When given at midday of diestrus 11, there were no effects on serum LH, FSH, or progesterone until the afternoon of proestrus. However, plasma estradiollevels were reduced by nearly 50% within 12 hours of treatment and remained low for an additional 12 hours. The agent was not directly uterotrophic in the ovariectomized mature rat, but doses of 25 mg/kg were able to reduce estradiol-stimulated uterine weight increases. 34.3.5 TRIBUTYLTIN One of the clearest examples of pesticide-induced populationlevel effects in wildlife via alterations of the endocrine system is that of altered reproductive development and subsequent population declines in marine snails brought about by exposure to tributyltin, an antifouling agent (Ankley and Giesy, 1998). In sensitive species, tributyltin has been shown to inhibit the metabolism of testosterone to estrogen by aromatase, with subsequent development of male reproductive organs in females, condition referred to as "imposex." Marine snails appear to be especially sensitive to this mechanism of action. Effective concentrations are in the range of nanograms per milliliter and population declines are a global observation. Despite the dramatic nature of the effect in marine snails, tributyltin-induced inhibition of aromatase has not been reported in higher organisms, including mammals. However, it is toxic to the mammalian immune system by a nonsteroid-receptor mode of action.
34.4 HYPOTHALAMIC-PITUITARY FEEDBACK LOOPS 34.4.1 FEMALE NEUROENDOCRINE REGULATION 34.4.1.1 Dithiocarbamates Dithiocarbamates are a broad chemical class including fungicides such as the ethylenbisdithiocarbamates, metam sodium and thiram. They are also metal chelating agents and are known to inhibit the synthesis of neurotransmitters, particularly norepinephrine, via chelation of the copper-containing portion of the enzyme dopamine-,B-hydroxylase. Norepinephrine plays a critical role in the release of gonadotropin-releasing hormones (GnRH) from the hypothalamus. During a short time period (between 1400 and 1600 hours on the day of proestrus), the sequential feedback of estrogen and then progesterone stimulates
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Pesticides as Endocrine-Disrupting Chemicals
the activity of a-adrenergic neurons which induces a phasic release of GnRH. This, in turn, triggers the anterior pituitary to release a surge of LH. Concentrations of LH rapidly rise in serum from levels below 1 ng/ml to 5-10 ng/ml and ovulation is induced. There is a minimum concentration for this surge to be effective in inducing ovulation in spontaneous ovu1ators such as the rat and the human. Because of the critical timing of events, and the multiple steps which are susceptible to disruption, analysis of the control of ovulation has proven a particularly useful tool in understanding neuroendocrine toxicology. Ovariectomized, estrogen-primed female rats given a single injection of 50 or 100 mg/kg thiram at 1100 hours lacked the expected elevation in LH measured over the succeeding 4 hours. A dose level of 25 mg/kg blocked the surge in some animals, and attenuated it in others (Stoker et aI., 1993). Administration of 50 mg/kg to intact females on the afternoon of vaginal proestrus delayed ovulation by 1 day. When mated the following evening, there were no differences in the proportion of sperm-positive females compared to controls mated the previous evening, but there was a significant decrease in embryo viability between gestation days 7 and 11, indicating that ovulation delayed for 24 hours was deleterious to the ripening oocytes (Stoker et aI., 1996). Doses of 50 and 100 mg/kg metam sodium induced nearly identical effects, and the surge effect was reversed by the a-adrenergic agonist c1onidine. In addition, anterior and posterior hypothalamic norepinephrine levels fell 3 hours after injection and were accompanied by a rise in dopamine (Go1dman et aI., 1994). Using sodium dimethyldithiocarbamate and comparing systemic versus intrabursal injections, Goldman et al. (1997) went on to demonstrate that, although the effect on hypothalamic catecholamine synthesis may underlie the ovulatory blockade, there is also a local ovarian response that is independent of ovarian norepinephrine concentrations. Thus, there appear to be multiple, albeit uncharacterized, modes of action by which the dithiocarbamates may disrupt the regulation of ovulation. 34.4.1.2 Formamidines The formamidines are a class of insecticides which include amitraz and chlordimeform, the latter of which is no longer marketed in the United States due to its carcinogenic potential. Their mode of insecticidal activity is based on mimicking the insect neurotransmitter octopamine, but they are also capable of binding to and inhibiting a-adrenergic receptors in mammals. As seen with the dithiocarbamates, interference with norepinephrine action in the hypothalamus can lead to significant alterations in female reproductive function. For instance, both amitraz and chlordimeform block the LH surge (Cooper et aI., 1994), which is mediated in part by norepinephrine. 34.4.1.3 Atrazine Concern for the endocrine-disrupting effects of atrazine, a chlorotriazine herbicide, arose following the observation of increased incidence of mammary tumors in a chronic bioassay in female Sprague-Dawley (SD) rats exposed to 400 ppm atrazine
in the diet for 104 weeks. These tumors also appeared in control females, but occurred earlier in the treated females. No other tumors were present in the treated Sprague-Dawley female rats, nor in male Sprague-Dawley rats or male and female Fischer 344 rats (Stevens et aI., 1994; Thakur et aI., 1998). The finding of an earlier onset of mammary tumors led to an investigation into the estrogenicity of atrazine, but under equilibrium conditions, atrazine was not able to compete with estradiol for binding to rat uterine estrogen receptors. A weak competition was noted if the cytoso1s were preincubated at 25°C prior to incubation with the tracer (Tennant et aI., 1994a). Somewhat conflicting results have been seen in other studies. Daily exposure of adult Fischer rats to 120 mg/kg for 7 days resulted in fewer treated females displaying normal estrous cycles, and the number of days in diestrus increased significantly. Fertility was reduced in females during the first week after exposure, but pregnancy outcome was not affected in those that became inseminated (Simic et aI., 1994). However, treatment of adult, ovariectomized SD rats with up to 300 mg/kg atrazine by oral gavage for 3 days did not result in an increase in uterine weight, nor were there increases in uterine progesterone levels, suggesting the lack of an estrogenic potential. Indeed, when estradiol (2 !J.g/kg subcutaneously) was given in conjunction with 300 mg/kg or orally administered atrazine, there was a weak inhibition (~25%) of the uterotrophic response (Tennant et al., 1994b). In a similar study, immature female SD rats were dosed with 0, 50, 150, or 300 mg/kg atrazine by gavage for 3 days. Uterine weight was not increased, but decreases in uterine progesterone receptors and peroxidase activities were noted; however, when combined with estradiol, antiestrogenic effects of atrazine including decreases in uterine progesterone receptor binding and uterine peroxidase was not noted on the uterus (Connor et aI., 1996). In this same study, atrazine did not affect basal or estradiol-induced MCF-7 cell proliferation, nor did it display agonist or antagonist action against estradiolinduced 1uciferase activity in MCF-7 cells transfected with a Ga14-regulated human estrogen receptor chimera. To further evaluate effects on reproductive function, female Long Evans (LE) and SD rats that had been screened for regular 4-day estrous cycles, received 0, 75, 150, or 300 mg/kg/day atrazine by gavage for 21 days. In both strains, atrazine disrupted the regular 4-day estrous cycles. For the LE rats, all dose levels were effective, whereas SD rats required a higher dose (150 mg/kg/day) for a longer time for this effect to appear. The increased time spent in vaginal diestrus was associated with elevated serum progesterone and low estradiol concentrations, indicative of a repetitive pseudopregnant condition. This hormonal condition was not considered by the authors to be conducive to the development of mammary tumors, although there was some indication of prolonged estrous at the lowest dose tested (Cooper et aI., 1996). The strain difference noted in the premature onset of mammary tumors (insensitive Fischer 344 rats versus sensitive SD rats) has been attributed to differences in the normal aging of the reproductive tract in these strains (Eldridge et aI., 1994; Stevens et aI., 1994; summarized in Chapin et aI., 1996). Re-
34.4 Hypothalamic-Pituitary Feedback Loops
productive cycling in the female SD rat begins to decline in animals less than a year of age, presumably due to the loss of sensitivity of adrenergic neurons in the hypothalamus that control GnRH release to the pituitary. This loss of stimulation reduces FSH and LH release, and ultimately ovulatory failure. In turn, the ovaries contain many follicles but no corpora lutea. In contrast, adrenergic neurons of female Fischer 344 rats do not seem to lose their sensitivity to estrogen stimulation, and regular cycling is maintained for a much longer time. Also in contrast to the onset of persistent estrus, reproductive aging in the Fischer 344 is believed to be due to an inability to control daily prolactin surges, a prolonged activity of the corpora lutea (i.e., repetitive pseudopregnancy), and a higher level of progesterone release. Hence, the endocrine milieu of the aging SD rat, but not the Fischer 344 rat, favors development of mammary tumors and helps explain the difference in incidence of spontaneous tumors as females of these strains age. How atrazine accelerates the neuroendocrine aging of the reproductive axis in the SD rat, however, has not been determined. Although the induction of mammary tumors by atrazine may not be relevant to humans (International Agency for Research on Cancer, 1999), the action of atrazine on the hypothalamus may be of some significance to human health. 34.4.2 THYROID TUMORS Endocrine disruption of the pituitary-thyroid axis is a relatively well understood process by which endogenous chemicals induce thyroid follicular cell neoplasia. The physiological regulation of thyroid cell growth and function involves a complex interactive network of trophic factors that are mediated by a number of second messenger systems (Hard, 1998). TSH is the main growth factor for follicular cells, with insulin-like growth factor 1 (IGF-l), epidermal growth factor (EGF), basic fibroblast growth factor (bFGF), and transforming growth factor f3 (TGF-f3) also involved in various ways. Activation of TSH receptors stimulates G protein-dependent rises in cAMP and phospholipase C, with resulting consequences of iodine uptake and release, thyroid peroxidase (TPO) generation, thyroid hormone synthesis and release, and thyroid cell growth and division. Relative to metabolism, T4 is secreted by the thyroid, but must be converted to T3 via either Type I 5' -diodinase in the liver or Type 11 5' -diodenase in the brain, pituitary, and brown adipose tissue. There are three main carrier proteins for thyroid hormones, thyroxine binding protein (65%), transthyretin (20%), and albumin (10%); only about 5% of the hormone is unbound (in the rat, thyroxine binding protein is absent during most of adult life). Further metabolism occurs in the liver, intestines, and kidneys and involves inactivation of biological activity by conjugation with glucuronic acid or sulfate. Whether by reduced synthesis due to inhibition of TPO, reduced peripheral de-iodination, or by elevated turnover via induction of conjugating enzymes, sustained release of TSH in response to decreased circulating levels of thyroid hormones is intimately involved in thyroid gland neoplasia. This suggests that nonlinear thyroid cancer dose-response considerations can be applied
739
to chemicals that reduce thyroid hormone levels, increase TSH and thyroid cell division, and are judged to lack mutagenic activity (Hill et aI., 1998). Although much of thyroid gland physiology is similar across experimental animals and humans, there are, as noted above, some important differences that may reduce the sensitivity of humans relative to rodents (Hard, 1998). Interestingly, childhood radiation is the only known exogenous risk for thyroid gland carcinogenesis in humans. A total of 240 pesticides have received an in-depth review for potential carcinogenicity by the US EPA, with evidence of induction of thyroid follicular tumors in appropriate chronic tests present for (Hurley, 1998). Thyroid tumors were second only to liver tumors in the frequency with which they were observed. Three pesticides (amitrole, ethylene thiourea, and mancozeb) induced a high incidence (>0.48) at relatively low daily doses (3.5-30.9 mg/kg/day). All but 2 of the 24 pesticides induced thyroid tumors only in rats; none induced tumors only in mice; 8 induced tumors only in males; none were positive in females only. Sixteen induced tumors in at least one other site, the most frequent being the liver; other sites of preponderance being the glandular stomach, mammary gland, parathyroid, pancreas islet cells, testis, and thyroid C cell. Based on gene mutation and chromosomal aberration tests, a direct mutational mode of action appears possible for only 3 of the 24 pesticides: acetochlor, ethylene thiourea, and etridiazole. None of the pesticides has had a complete evaluation of all potential sites of antithyroid action (inhibition of iodide uptake, inhibition of thyroid peroxidase, damage to thyroid follicular cells, inhibition of thyroid hormone release, inhibition of 5' -monodeiodenase activity, and enhancement of metabolism and excretion by the liver), but 12 have sufficient information to infer a mode of action. Three (amitrole, ethylene thiourea, and mancozeb) inhibit thyroid peroxidase; four inhibit the iodide pump (amitrole, ethiozin, ethylene thiourea and pentachloronitrobenzene), nine stimulate thyroid hormone metabolism and excretion (acetochlor, clofentezine, fenbuconazole, fipronil, pendimethrin, pentachloronitrobenzene, prodiamine, pyrimethanil, and thiazopyr). 34.4.3 LEYDIG CELL TUMORS Leydig cells, which are situated in the interstitial space between seminiferous tubules in the testes, serve as the primary source for synthesis of androgens in mammals. As in the thyroid gland, alterations in endocrine feedback loops in the hypothalamic-pituitary-gonadal axis have been linked to Leydig cell hyperplasia-and, ultimately, neoplasia (reviewed in Cook et aI., 1999). The primary trophic influence over Leydig cells is exerted by pulsatile release of LH by the anterior pituitary following stimulation by GnRH originating in the preoptic and medial basal areas of the hypothalamus. In the Leydig cell, LH activates adenyl cyclase via a G protein, which initiates a cascade of events leading to increased steroidogenesis. Testosterone completes the feedback loop by providing inhibitory signals to the hypothalamus and pituitary to lower LH secretion. Also similar to observations regarding the thyroid
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Pesticides as Endocrine-Disrupting Chemicals
gland, a number of other trophic factors affecting the Leydig cells have been identified, including IGF-l, TGF-a, TGF-,B, bFGF, interleukin I (IL-l), inhibin, and activin. Because Leydig cell hyperplasia is frequently noticed in areas proximate to seminiferous tubule damage, a local endocrine imbalance has been assumed to play a key role, but in nearly all cases a sustained elevation in serum LH is a key mediator of the response. Almost all Leydig cell tumors are benign adenomas, and the distinguishing feature between hyperplasia and tumors is the size of the nodule. There are species differences in the susceptibility to Leydig cell hyperplasia produced by exposure to pesticides. For example, Murakami et al. (1995) studied the effect of procymidone, an antiandrogenic dicarboximide fungicide, in inducing testicular interstitial cell tumors in rats and mice. Male SpragueDawley rats and ICR mice were exposed via the diet to 7006000 ppm and 1000-10,000 ppm, respectively, for 3 months. In the rat, serum and intratesticular testosterone and serum and pituitary LH levels were increased in treated animals throughout the exposure. The endocrine effects were more pronounced at 4 versus 13 weeks. Interstitial cells were also hypersensitive to hCG in vitro at 2, 4, and 13 weeks, as evidenced by the rate of testosterone release. In contrast, although serum and pituitary levels of LH were elevated after 4 weeks of exposure, no significant changes in testosterone were detected in mice, either in vivo or in vitro. For a reason not yet understood, rat Leydig cells appear more sensitive to chronic stimulation by LH than the mouse cells, an effect which correlates with observations of tumors. As noted by Cook et al. (1999), Leydig cell tumors are exceptionally rare in human populations.
34.5 MISCELLANEOUS OR UNKNOWN MODES OF ACTION 34.5.1 NITROFEN Nitrofen (2,4-dichlorophenyl-p-nitrophenyl ether) is a preemergent herbicide removed from the U.S. market in the early 1980s due to concerns over its developmental toxicity. Administration during gestation to both rats and mice induces a constellation of developmental alterations that in many respects resemble those of altered thyroid hormone function. Consistent with the similarity in effects, nitrofen also bears structural similarity to thyroid hormone. These effects include delayed maturation of the lungs and pulmonary surfactant, small or missing Harderian glands, growth deficits, delayed opening of the eyes, and altered neurobehavioral development. In addition, malformations of the cardiovascular system and urogenital track and diaphragmatic hernias have been observed (Gray et aI., 1982; Ostby et aI., 1985). Developmental toxicity was observed at doses as low as 4.17 mg/kg/day given to rats on gestation days 8-16. In subsequent studies to examine whether nitrofen interferes with the hypothalamic-thyroid axis, adult mice were given 500 or 1000 mglkglday for 3 days (Gray and Kavlock, 1983). Serum thyroxine was reduced by 60% in the high-dose
group and by 20% at the low dose, but no changes in serum T3 levels were found. In this study, there was no effect on body weight, but liver weights were slightly increased. In support of the role of altered thyroid function induced by nitrofen on the developing organism, Manson et al. (1984) reported that a single dose of 250 mg/kg administered to pregnant rats on gestation day 11 significantly depressed TSH. The decrease was most evident at 6 and 24 hours after dosing, but were still present at term. Maternal serum T4 levels levels were likewise depressed at 8 and 24 hours after treatment, but were normal at term. No effects were seen in T3. Although confounded by the effects ot thyroidectomy, administration of T4 (4 Ilg per 100 g body weight) with 25 mg/kg nitrofen (days 9-11 of gestation) resulted in a 70% reduction in malformations, with the heart most protected and the kidney least protected. Competitive radioimmunoassay binding studies indicated that the 4-hydroxy-2,5-dichloro-4'-aminophenyl ether metabolite could displace T3. Collectively these results indicate that some of the nitrofen-induced developmental toxicity is potentially mediated by alterations in the thyroid status of the maternal organism and her fetus.
34.6 IMPACT ON TESTING GUIDELINES Assessment of the potential developmental and reproductive risks of environmental contaminants is generally determined through application of testing guidelines that are established by regulatory agencies such as the U.S. EPA or by international coordinating bodies such as the Organization of Economic Cooperation and Development (OECD). In 1991, the U.S. EPA began a process to update its developmental and reproductive testing guidelines to ensure that they incorporated contemporary scientific methodology. Traditionally, these tests had been very apical in nature; that is, they relied on end points which were diagnostic of adverse biological outcomes, but did not provide clarification of potential modes of action, target organs, or most sensitive life stage or gender. For example, the multi generation reproductive test guidelines require groups of animals (generally rats) to be exposed to the test chemical beginning shortly after weaning and continuing until production of the second generation (thus concluding with examination of offspring of animals exposed from fertilization and through reproduction). In the past, the primary end points that were evaluated in such tests included fertility (are the animals capable of reproducing?), fecundity (how many offspring are produced?), and growth of the offspring. A large body of evidence suggests, especially for endocrine-disrupting chemicals, that these end points are neither very sensitive to reproductive disturbance nor indicative of the underlying biological effect. Both these issues raise concern regarding the suitability of previously issued test guidelines to satisfactorily detect and characterize reproductive hazard. As part of this effort to revise the testing guidelines, particular emphasis was placed on improving the ability to detect the action of chemicals that may act via the endocrine system to
34.6 Impact on Testing Guidelines
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Table 34.4 Summary of Relative Sensitivity of End Points in the Traditional Multigenerative Reproduction Study (Fertility, Fecundity, and Somatic Growth) Compared with Alternate Measures of Reproductive Function or Capacity for Chemicals that Work through the Estrogen, Androgen, and Ah Receptorsa Weak estrogenb Female
Male
Fertility
+
+
Fecundity
+
Male
Growth ofF1
++
Physiology
Male
Female + (TTP)
+ -(AGD)
-(AGD)
+++ (YO)
++ (PS)
+ (YO)
++(EC)
+ + + (EO)
+++(AGD) + (PS)
Female
++
Sex differentiation Puberty
Ah agonistd
Antiandrogen C
-(AGD)
Gamete number
++ (CSC)
+ (ESC)
+ + + (ESC)
Accessory sex gland weights
++
+++
+
Gonad weight Pituitary hormones
+ +++(PL)
Steroid hormones
-(isT)
+
Malformations
-(T) + (HS)
++(YT)
QThe lowest dose level at which a statistically significant effect was observed is noted by symbols (the lower the dose within a chemical, the number of "+"s, see footnotes b-d for chemical, exposure duration, and dose levels). Reprinted with permission from Kavlock (1999). bMethoxychlor, GDI5-PD21; dose levels: -, 200 mg/kg/day; +, 100 mg/kg/day; ++,50 mg/kg/day; + + +, 25 mg/kg (Gray et aI., I988a, 1989). Fertility and fecundity data are for the parental generation. cYinc!ozolin, GDI4-PD3; dose levels: -, 100-200 mglkg/day; +,50 mg/kg/day; + + + 12.5-25 mg/kg/day; + + +, 3-6 mg/kg/day (Gray et aI., 1994; Ostby et aI., 1997). dDioxin, GD 15; dose levels; -, I j.lg/kg; +,0.8 j.lg/kg; ++, 0.2 j.lg/kg; + + +,0.05 j.lglkg (Gray et aI., 1995, 1997b, c; Gray and Ostby, 1995). Abbreviations: AGD, anogenital distance; CSC, caudal sperm count; EC, estrous cyC!icity; EO, age at eye opening; ESC, ejaculated sperm count; GD, gestation day; HS, hypopsadias; isT, in vitro stimulated testosterone release from testes; PD, postnatal day; PL, prolactin; PS, age at preputial gland separation; YO, age at vaginal opening; YT, vaginal thread; T, serum testosterone; TTP , time to pregnancy.
perturb reproduction. Data from multigeneration studies (summarized in Table 34.4), obtained from the same laboratory so that comparison across end points and chemicals is relatively straightforward, indicate that, for chemicals that act via the estrogen receptor (e.g., methoxychlor), the androgen receptor (e.g., vinclozolin), or the Ah receptor (e.g., dioxin), the traditional end points of fertility, fecundity, and growth do tend to pick up effects, but confirm that there are a lack of sensitivity and a poor ability to characterize the overall impact. For example, one of the most sensitive indicators of developmental exposure to an estrogen is accelerated puberty in the female (Gray et aI., 1988a), whereas diminished anogenital distance and accessory sex gland weights are most sensitive to developmental exposure to an anti-androgen (Gray et aI., 1994), and decreased ejaculated sperm counts are the most sensitive to chemicals that act via the Ah receptor (Gray and Ostby, 1995). None of these effects would have been identified by the traditional multigeneration reproductive test. The effects of methoxychlor on the growth of the parental males present an interesting example of where additional information can help interpret the data (Gray et aI., 1988b). In this instance, the impairment of growth might be considered a manifestation of systemic toxicity if it were not known that estrogens will decrease appetite in males and hence result in decreased food consumption and body growth. How the emerging finding (Maness et aI., 1998) that methoxychlor interacts with both the estrogen receptor and the androgen (as an agonist and antago-
nist, respectively), remains to be reconciled with the pattern of effects observed in vivo, where the estrogenic effects appear to prevail. The newly harmonized multigeneration reproductive testing guidelines (U.S. EPA, 1997b) now include a number of end points to monitor reproductive performance and health. These include assessments of the following: female estrous cyclicity; sperm parameters (total number, percentage progressively motile, and sperm morphology in both the parental and FI generations); the age at puberty in the FI generation (vaginal opening in the female, preputial separation in the males); an expanded list of organs for pathology, gravimetric analysis, and/or histopathology to identify and characterize effects at the target organ; some triggered end points including anogenital distance in the F2 generation and primordial follicu1ar counts in the parental and FI generations. For the new prenatal developmental toxicity test guidelines (U.S. EPA, 1997b), one important modification related to the improved detection of endocrine disruptors was the expansion of the period of dosing from the end of organogenesis (i.e., palatal closure) to the end of pregnancy in order to include the developmental period of urogenital differentiation. Collectively these modifications of the test guidelines should markedly improve the characterization of endocrine-mediated effects during reproduction and development.
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34.7 IMPACT OF FOOD QUALITY PROTECTION ACT Over the past several years, two environmental laws enacted by the United States Congress specifically require that pesticides and other chemicals found in or on food or in drinking-water sources be tested for their potential to cause "estrogenic or other endocrine effects in humans." The Food Quality Protection Act of 1996 (FQPA) and the Safe Drinking Water Act Amendments of 1996 (SDWA) require the U.S. EPA to, within 2 years of enactment, develop a screening program using appropriate, valid test systems to determine whether substances may have estrogenic or other endocrine effects in humans. The screening program must undergo a public comment period and peer review and be implemented within 3 years. The laws require that the manufacturers, registrants, or importers of the pesticides and other substances conduct the testing according to the program the U.S. EPA develops. At joint workshops, cosponsored by U.S. EPA, the Chemical Manufacturers Association, and the World Wildlife Fund (Ankley et aI., 1998; DeVito et aI., 1999; Gray et aI., 1997a), a number of assays potentially suitable for assessing endocrine-disrupting chemicals (EDCs), particularly those for detecting (anti-)estrogenic, (anti-)androgenic, and (anti-)thyroidogenic effects, were identified and critiqued. Based upon input from these and other workshops and its own deliberations, the Endocrine Disruptor Screening and Testing Advisory Committee (EDSTAC), an advisory committee to the U.S. EPA on implementation of the FQPA and SDWA, has recommended a battery of assays for both screening and testing potential EDCs that will be used to address the mandates of the FQPA and SDWA (U.S. EPA, 1998). The assays are intended to detect potential interaction both with the sex steroids (estrogen and testosterone) and with thyroid hormone function, and they include assessment of both potential human health effects and effects in wildlife. To help prioritize chemicals for screening and testing, the EDSTAC recommended a high-throughput screening (HTPS) cellular-based, receptor-mediated gene transcription assay for chemicals which act as either agonists or antagonists for estrogen, androgen, or thyroid receptors. It has been estimated that perhaps 15,000 chemicals would be evaluated in the HTPS. The EDSTAC recommendation for the "Tier 1" screening (TIS) battery includes three in vitro assays and five in vivo assays. The in vitro assays in TIS include an estrogen receptor binding or transcriptional activation assay; an androgen receptor binding or transcriptional activation assay; and a steroidogenesis assay using minced testis. The five in vivo screens recommended include the rodent 3-day uterotrophic assay, a rodent 20-day pubertal female assay for effects on thyroid function, a male rodent 5- to 7-day Hershberger assay, a frog metamorphosis assay for thyroid effects, and a fish gonadal recrudescence assay. It is estimated that perhaps as many as 1500 chemicals would enter the TIS, and positive chemicals would move into a second level (T2T), where more defined toxicological responses would be characterized. In the U.S. EPAs Reports to Congress in 2000 (U.S. EPA, 2000) it stated that it is proceeding on two fronts to imple-
ment the screening and testing program. The first element is establishing a method for setting priorities for screening. For commercial chemicals and environmental contaminants other than pesticides, this method will include use of a database and software that EPA is developing. Prioritization of pesticidal active ingredients will be done in conjunction with a review of existing data on health and environmental effects. In the second element, it is ensuring that the Tier 1 and 2 assays are scientifically valid. Validation includes developing protocols to conduct specific assays, evaluating their effectiveness, and ensuring that the assays can be performed reliably and consistently in different laboratories. This effort is being conducted in liaison with the Interagency Coordinating Committee for the Validation of Alternative Methods (ICCVAM) and is following ICCVAM principles. In the Report to Congress, EPA estimated that the Tier 1 screens would be validated by the end of 2002, and the Tier 2 by 2004. Updates on progress of the screening and testing program can be found at: http.//www.epa.gov/ endocrine/.
ACKNOWLEDGMENTS Without the most helpful assistance of many members of the Reproductive Toxicology Division, including Janice Brown, Jerome Goldman, Susan Laws, Earl Gray, and Ralph Cooper, this effort would not have been possible.
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CHAPTER 34
Pesticides as Endocrine-Disrupting Chemicals
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Minor, J. L., Knapp, H. F., Stuart, B.O., Killinger, J. M., Zwicker, G. M., and Freudenthal, R. I. (1984). Evaluation of male rat fertility following inhalation exposure to ordram. Toxicologist 4, 80. Monosson, E., Kelce, W. R., Lambright, c., Ostby, J., and Gray, L. E., Jr. (1999). Peripubertal exposure to the antiandrogenic fungicide, vinclozolin, delays puberty, inhibits the development of androgen-dependent tissues, and alters androgen receptor function in the male rat. Toxicol. Ind. Health 15(1-2),65-79. Murakami, M., Hosokawa, S., Yamada, T., Harakawa, M., Ito, M., Koyama, Y., Kimura, J., Yoshitake, A., and Yamata, H. (1995). Species-specific mechanism in rat Leydig cell tumorigenesis by procymidone. Toxico!. Appl. Pharmacol. 131, 244-252. National Research Council (NRC) (1999). "Hormonally Active Agents in the Environment." National Academy Press, Washington, DC. Nelson, J. A., Struck, R. F., and James, R. (1978). Estrogenic activities of chlorinated hydrocarbons. J. Toxico!. Environ. Health 4, 325-339. Organization of Economic Cooperation and Development (1997). "Appraisal of Test Methods for Sex Hormone Disrupting Chemicals." Draft detailed review paper. Environment Directorate, OECD, Paris. Available at www.oecd.org/ehs. O'Shea, T., Reeves, R. R., and Long, A. K. eds. (1999). "Marine Contaminants and Persistent Ocean Contaminants: Proceedings of the Marine Mammal Commission Workshop," Keystone Colorado, 12-15 October 1998. Marine Mammal Commission, 4340 East-West Highway, Room 905, Bethesda, MD 20814. Ostby, J. S., Gray, L. E., Kavlock, R. J., and Ferrell, J. M. (1985). The postnatal effects of prenatal exposure to low doses of nitrofen (2,4) in SpragueDawley rats. Toxicology 34, 285-297. Ostby, J., Kelce, w., Lambright, c., Wolf, c., Mann, P., and Gray, L. E. (1999). The fungicide procymidone alters sexual differentiation in the male rat by acting as an androgen-receptor antagonist in vivo and in vitro. Toxico!. Ind. Health 15,80-93. Palmitar, R. D., and MulvihiII, E. R. (1978). Estrogenic activity of the insecticide Kepone on the chicken oviduct. Science 201, 356-358. Pont, A., WiIIiams, P. L., Loose, D. S., Feldman, D. S., Reitz, R. E., Bochra, c., and Stevens, D. A. (1982a). Ketoconazole inhibits adrenal steroid synthesis. Ann. Intern. Med. 97, 370-372. Pont, A. P., WiIIiams, P. L., Azhar, S., Reitz, R. E., Bochra, c., Smith, E. R., and Stevens, D. A. (1982b). Ketoconazole blocks testosterone synthesis. Arch. Intern. Med. 142, 2137-2140. Reiter, L. W., DeRosa, C., Kavlock, R. J., Lucier, G., Mac, M. J., Melillo, J., Melnick, R. L., Sinks, T., and Walton, B. T. (1998). The V.S. federal framework for research on endocrine disruptors and an analysis of research programs supported during fiscal year 1996. Environ. Health Perspect. 106(3), 105-113. Ronis, M. J., Barger, T. M., Gandy, J., Bell, L. M., and Green, K. (1995). Antiandrogenic effects of perinatal cypermethrin exposure in the developing rat. Neurotoxicology 16, 673. Rosencrans, J. A., Squibb, R. E., Johnson, J. H., Tilson, H. A., and Hong, J. S. (1985). Effect of neonatal chlordecone exposure on pituitary-adrenal function in adult Fischer 344 rats. Neurobehavioral Toxico!. Teratol. 7, 33-37. Schwartz, W. J., and Corkem, M. (1992). Effects of methoxychlor treatment of pregnant mice on female offspring of the treated and subsequent pregnancies. Reprod. Toxicol. 6, 431-437. Simic, B. S., Kniewald, J., and Kniewald, Z. (1994). Effect of atrazine on reproductive performance in the rat. J. App!. Toxico!. 14(6),401-404. Sonino, N. (1987). The use of ketoconazole as an inhibitor of steroid production. New England J. Med. 317, 812-818. Squibb, R. E., and Tilson, H. A. (1982). Effects of gestational and perinatal exposure to chlordecone (Kepone) on the neurobehavioral development of Fischer 344 rats. Neurotoxicology 3, 17-26. Steinmetz, R., Young, P. C. M., Caperall-Grant, A., Gize, E. A., Madhukar, B. v., Ben-Jonathan, N., and Bigsby, R. M. (1996). Novel estrogenic action of the pesticide residue ,B-hexachlorocyclohexane in human breast cancer cells. Cancer Res. 56, 5403-5409. Stevens, J. T., Breckinridge, C. B., Wetzel, L. T., Gillis, J. H., Luempert, L. G., III, and Eldridge, J. C. (1994). Hypothesis for mammary tu-
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morigenesis in Sprague-Dawley rats exposed to certain triazine herbicides. 1. Toxieo!. Environ. Health 43, 139-153. Stoker, T. E., Cooper, R. L., Goldman, J. M., and Andrews, J. E. (1996). Characterization of pregnancy outcome following thiram-induced ovulatory delay in the female rat. Neurotoxieo!' Teratol. 18,277-282. Stoker, T. E., Goldman, J. M., and Cooper, R. L. (1993). The dithiocarbamate fungicide thiram disrupts the honnonal control of ovulation in the female rat. Reprod. Toxieol. 7, 21I-218. Swartz, W. J., Eroschenko, V. P., and Schutzman, R. L. (1988). Ovulatory response of chlordecone (kepone)-exposed mice to exogenous gonadotropins. Toxicology 51,147-153. Swedish Environmental Protection Agency (1998). "Endocrine Disrupting Substances-Impainnent of Reproduction and Development" (P.-E. OIsson, B. Borg, B. Brunstrom, H. Hakansson, and E. Klasson-Wehler, eds.). Elanders Gotab, Stockholm. Tattersfield, L., Matthiessen, P., CampbeIl, P., Grandy, N., and Lange, R., eds. (1997). "Workshop on Endocrine Modulators and Wildlife: Assessment and Testing." SETAC-Europe, Brussels. Tennant, M. K., HilI, D. S., Eldridge, J. c., Wetzel, L. T., Breckenridge, C. B., and Stevens, J. T. (1994a). Chloro-s-triazine antagonism of estrogen action: Limited interaction with estrogen receptor binding. 1. Toxieo!. Environ. Health 43,197-211. Tennant, M. K., HilI, D. S., Eldridge, J. c., Wetzel, L. T., Breckenridge, C. B., and Stevens, J. T. (l994b). Possible anti-estrogenic properties of chloro-striazines in rat uterus. 1. Toxieo!. Environ. Health 43, 183-196. Thakur, A. K., Wetzel, L. T., VoeIker, R. w., and Wakefield, A. E. (1998). Results of a two-year oncogenicity study in the Fischer 344 rats with atrazine. In "Triazine Herbicides Risk Assessment." (L. G. Ballentine, J. E. McFarland, and D. S. Hackett, eds.), ACS Symposium Series, Vo!. 683, pp. 384398. Am. Chem. Soc., Washington, DC. Toppari J., Larsen, J. c., Christiansen, P., Giwercman, A., Grandjean, P., Guiliette, L. J., Jr., Jegou, B., Jensen, T. K., Jouannet, P., Keiding, N., Leffers, H., McLachlan, J. A., Meyer, 0., Muller, J., Rajpert-De Meyts, E., Scheike, T., Sharpe, R., Sumpter, J., Skakkebaek, N. E. (1996). Male reproductive health and environmental xenoestrogens. Environ. Health Perspeet. 104(Supp!. 4), 741-803.
Umweltbundesamt (1995). "Endocrinically Active Chemicals in the Environment," Texte Series 3, Vo!. 36, Expert Round, 9-10 March 1995, Berlin, Gennany. Uphouse, L. (I 985a). Effects of chlordecone on neuroendocrine function of female rats. Neurotoxieology 6, 191-210. Uphouse, L. (1985b). Single injection with chlordecone reduced behavioral receptivity and fertility of adult rats. Neurobehavioral Toxieol. Teratol. 8, 121-126. Uphouse, L., and WilIiams, J. (1989). Diestrous treatment with lindane disrupts the female reproductive cycle. Toxieo!. Lett. 48, 21. U.S. Environmental Protection Agency (U.S. EPA) (l997a). "Special Report on Environmental Endocrine Disruption: An Effects Assessment and Analysis." EPAl6301R-012, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA) (l997b). Toxic Substances Control Act test guidelines; final rule, Federal Register 62(158), 4381943861 (August 15). U.S. Environmental Protection Agency (US. EPA) (1998). "Endocrine Disruptor Screening and Testing Advisory Committee (EDSTAC) Final Report," Vo!. 1. U.S. Environmental Protection Agency (U.S. EPA) (2000). Endocrine Disruptor Screening Program-Report to Congress, August 2000. WaIler, D. P., Martin, A., Vickery, B. H., and Zaneveld, L. J. (1990). The effect ofketoconazole on fertility of male rats. Contraception 41(4), 41I-417. Warren, D. w., Ahmad, N., and Rudeen, P. K. (1988). The effects of fetal exposure to 1,2-dibromo-3-chloropropane on adult male reproductive function. Bio!. Reprod. 39, 707-716. Wickramaratne, G. A., Foster, J. R., Ellis, M. K., and Tomenson, J. A. (1998). Molinate: Rodent reproductive toxicity and its relevance to humans-a review. Regul. Toxieo!. Pharmaeol. 27, 1l2-1I8. You, L., Casanova, M., Archibeque-Engle, S., Sar, M., Fan, L. Q., and Heck, H. A. (1998). Impaired male sexual development in perinatal Sprague-Dawley and Long-Evans hooded rats exposed in utero and lactationally to p,p'-DDE. Toxieol. Sci. 45(2),162-173.
CHAPTER
35 Genetic Toxicity of Pesticides David A. Eastmond and Sharada Balakrishnan University of California, Riverside
35.1 INTRODUCTION Pesticides are biologically active compounds selected and used for their toxic properties. In many cases, these agents are highly specific in their pesticidal effects, acting on a unique molecular target or affecting a narrow range of organisms. Other pesticides can affect a much broader range of targets and organisms, including humans. As a result, there exist ongoing concerns about the health effects of pesticide exposure in humans. These concerns have been heightened by pesticide-related poisoning episodes that have occurred over the past 50 years such as those involving hexachlorobenzene (Schmid, 1960), methylmercury (Bakir et aI., 1973), malathion (Baker et aI., 1978), dibromochloropropane (Whorton et aI., 1979), aldicarb (Green et aI., 1987), and methylparathion (Rehner et aI., 2000). In addition to acute effects, substantial concerns exist about chronic effects such as cancer and heritable diseases that might stem from pesticide exposure. An association between pesticide exposure and cancer has been suspected for more than 40 years following reports of the occurrence of elevated levels of skin and lung cancer in European farmers using arsenical insecticides in grape production (Jungmann, 1966; Roth, 1958; Thiers et aI., 1967). In a few cases, the association between pesticide exposure and cancer has been confirmed (Blair and Zahm, 1995; IARC, 1987a, 1994; IOM, 1999; Zahm et al., 1997). However, in many cases, these concerns remain unsubstantiated either due to an underlying lack of an association or because of the difficulties in conducting epidemiological studies in these exposed populations. Even where associations have been seen or suspected, identifying the specific agent responsible has been difficult for many reasons, including poorly defined and variable exposure levels, exposure to multiple pesticides as well as other potentially carcinogenic agents, long latency periods, small study populations, and other confounding factors. As additional limitations, human epidemiological studies are costly and can only take place following exposure-an approach that is not considered protective of public health. Because of these difficulties, regulatory agencies and other organizations have turned to chronic animal bioassays, short-term tests, and other relevant data, in addition to human epidemiological studies, to evaluate the carcinogenicity and potential carcinogenicity of Handbook of Pesticide Toxicology Volume 1. Principles
pesticides and other agents. These agencies, such as the International Agency for Research on Cancer (IARC),* the D.S. Environmental Protection Agency (EPA), and the National Toxicology Program (NTP) have adopted a weight-of-the-evidence approach to make decisions on the carcinogenicity of an agent. For example, after reviewing the human, animal, and relevant biological data for one class of pesticides, IARC concluded that the spraying and application of nonarsenical insecticides entail exposures that are probably carcinogenic to humans (IARC, 1991). To date, a relatively small number of pesticides ( < 10) have been recognized by one or more of these organizations as human carcinogens (Goldman, 1998; IARC, 2000; NTP, 2000). It should be noted that, in these cases, the primary evidence has come, not from agricultural uses, but from studies of exposed workers manufacturing the agent for other industrial uses or, in the case of inorganic arsenic, from therapeutic and industrial uses as well as environmental exposures (IARC, 1987a). For instance, most of the evidence for the carcinogenicity of agents such as inorganic arsenic, benzene, cadmium, and chromium(VI), which historically were used as pesticides,t as well as agents currently registered for use such as ethylene oxide and coal tar, have been obtained from studies involving nonagricultural uses. In some cases, it is believed that the agent responsible for the toxic effects seen in the pesticideexposed individuals is a contaminant or an "inert" ingredient in the pesticide formulation rather than the active ingredient itself. For example, many of the adverse effects proposed as *The International Agency for Research on Cancer, part of the World Health Organization, produces authoritative evaluations of the carcinogenic risks of chemicals and other agents to humans. Following a critical review of both human and animal studies, IARC classifies agents as exhibiting sufficient evidence of carcinogenicity, limited evidence of carcinogenicity, inadequate evidence of carcinogenicity, or evidence suggesting a lack of carcinogenicity in animals or humans. As a final step, the IARC considers the entire body of evidence, including mechanistic information, to reach an overall evaluation of the carcinogenicity of the agent to humans. Other regulatory agencies such as the U.S. Environmental Protection Agency and the U.S. National Toxicology Program use similar approaches to evaluate the carcinogenicity of chemical agents. t Some forms of arsenic are still registered in the United States for use under severely restricted conditions. Arsenicals continue to be used as insecticides and wood preservatives in other countries (Zahm et aI., 1997).
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Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
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being associated with the chlorophenoxyacetic acid herbicides are believed to be due to contamination by low levels of2,3,7,8tetrachlorodibenzo-para-dioxin (TCDD), a potent animal and human carcinogen (IARC, 1997; IOM, 1999, 2000). Furthermore, it is conceivable that other cancers such as the leukemias and non-Hodgkin's lymphomas that have been attributed to pesticide exposure may in part be due to the use of benzene and other solvents as ingredients in the formulation products (Blair and Zahm, 1995; Petrelli et ai., 1993). As indicated previously, chronic testing in animals is also used by regulatory agencies to evaluate the carcinogenic effects of chemical agents. Animal bioassays have been conducted for a considerable number of individual pesticides and a significant number have been reported as tumorigenic in one or more animal tissues. According to Zahm and Ward (1998), of the 51 pesticides evaluated prior to 1990 by the U.S. National Cancer Institute and the National Toxicology Program, 24 exhibited carcinogenicity in chronic animal bioassays. These authors further reported that, as of 1997, the IARC had classified 26 pesticides as having sufficient evidence of carcinogenicity in animals and 19 as having limited evidence. However, because of their cost, lengthy duration, and concern that the results may not be directly relevant to humans, these bioassay results are considered as less than ideal and are usually evaluated in conjunction with additional types of biological information. In addition to human and chronic animal studies, regulatory agencies often rely on other relevant biological data to assist in the evaluation of carcinogenicity. These other data may include information on preneoplastic lesions, tumor pathology, genetic and related effects, structure-activity relationships, metabolism and pharmacokinetics, physicochemical parameters, and mechanisms of action (IARC, 199ge). In particular, short-term tests evaluating the genetic toxicity of the agent are often relied on in the decision-making process. The development and interpretation of these short-term tests have stimulated the development of the field of genetic toxicology. As a subspecialty of toxicology, genetic toxicology is concerned with the adverse effects of chemicals and other physical agents on the deoxyribonucleic acid (DNA) and other genetic components of living organisms. The primary focus of this discipline is to identify the agents and mechanisms involved in the formation of mutations-heritable genetic alterations in cells. When broadly defined, mutagenesis encompasses the induction of DNA damage as well as all types of genetic alterations, ranging from a single nucleotide change in the DNA sequence to large-scale changes in chromosome structure and number. The recognition that cancer is fundamentally a genetic disease, combined with the close association that has been seen between mutagenicity and carcinogenicity, has led to the use of mutation and genotoxicity assays as screens to identify agents likely to be carcinogenic or cause other genetic diseases. Over the past 30 years, a large number of short-term tests have been developed as screening tools to identify genotoxic and mutagenic chemicals. These short-term tests may employ bacteria, yeast, plants, insects, isolated mammalian cells, or whole animals and can be performed for a fraction of the cost and time required for a long-
term cancer bioassay. In addition, a number of these assays have been modified for use in detecting genetic alterations occurring in human populations exposed to genotoxic and carcinogenic agents. The objective of this chapter is to provide an overview of the methods of genotoxicity testing and their application to identifying pesticides capable of inducing genetic damage. The initial section will focus on the most common short-term tests that are employed for detecting the genotoxicity of pesticides in model systems and the use of these assays to detect genetic alterations in exposed humans. This will then be followed by an overview of the results of genotoxicity studies that have been performed on individual agents and studies of genetic damage in pesticideexposed workers. The last section will briefly address the value and interpretation of this information in the safety evaluation and risk assessment process.
35.2 GENOTOXICITY TESTS Hundreds of short-term tests have been developed to screen chemicals for potential mutagenic and carcinogenic effects. These assays measure effects ranging from DNA adduct formation to mutations induced in transgenic animals. A listing of representative short-term tests as well as a brief description of how these effects are measured is presented in Tables 35.1 and 35.2. Each of these genotoxicity assays has its own unique characteristics and measures only a subset of the possible heritable alterations involved in cancer and other genetic diseases. As a result, combinations of short-term tests are often used to increase the likelihood of detecting genotoxic effects. Over the years, requirements for genotoxicity testing have been established in the United States and other developed nations, and agents being proposed as new pesticides must undergo testing prior to registration. Using the EPA requirements as an example, the initial test battery includes (1) a gene mutation assay in bacteria, typically the Salmonella typhimurium reverse mutation assay; (2) one of several gene-inactivating (forward) mutation assays using mammalian cells in culture; and (3) an in vivo assay for chromosomal effects in mammalian bone marrow cells using either metaphase analysis for structural aberrations or the micronucleus assay (Auletta et aI., 1993; Dearfield et aI., 1991). Depending on the results of the initial battery, as well as other relevant information, additional shortterm testing or a chronic animal bioassay may be required. As a general principle, agencies such as the EPA place greater weight on tests conducted in eukaryotes than in prokaryotes and in mammalian species rather than in submammalian species when conducting a hazard evaluation of a chemical (Auletta et aI., 1993). For heritable noncancer risks, the results from studies in germ cells are accorded more weight than those obtained using somatic cells. Because of their prominent role in the testing of pesticides, the principal required assays will be described in more detail. For more detailed reviews of these and other short-term genotoxicity tests, the reader is referred to more
35.2 Genotoxicity Tests
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Table 35.1 Representative Short-Term Tests for Genotoxicity
Type of test
Specific test
DNA adduct formation
Covalent binding of radiolabeled chemicals 32p-postlabeling of adducts Measurement of oxygen radical-derived adducts Immunological detection of adducts
DNA damage in microorganisms
Pol A test rec test Mitotic recombination, mitotic crossing over, or mitotic gene conversion in yeast (D3, D4, D5, or D7 assays)
DNA damage in mammalian cells
Unscheduled DNA synthesis (UDS) Single-cell gel electrophoresis (Comet) assay Sister chromatid exchange (SCE)
Gene mutation in bacteria and fungi
Salmonella microsome reversion assay (Ames test) WP2 assay Yeast "forward" and "reverse" assays Miscellaneous
Gene mutation in higher systems
HPRT, TK, and Na/K-ATPase assays in vitro Sex-linked recessive lethal assay Tradescantia or maize waxy locus plant tests HPRT assay in vivo Mutation in lac IlIac Z-bearing transgenic animals
Chromosomal effects in isolated cell systems
In vitro cytogenetics assays In vitro micronucleus test Aneuploidy assays
Chromosomal effects in whole organisms
In vivo cytogenetics Micronucleus test Nondisjunction assay Heritable translocation assay Dominant lethal assay Alterations in germ cells
Oncogenic transformation
Transformation assays (clonal or focus)
Modified from U.S. EPA (1979).
comprehensive sources (IARC, 1980; IPCS, 1985; Rice et aI., 1999b). 35.2.1 BACTERIAL REVERSE MUTATION ASSAY
The bacterial reverse mutation assay uses specially engineered amino acid-requiring strains of Salmonella typhimurium (S. typhimurium) or, less frequently, Escherichia coli (E. coli) to detect point mutations, which involve the substitution, deletion, or insertion of one or a few DNA base pairs (IPCS, 1985; D.S. EPA, 1998c). The widely used Salmonella assay was developed by Ames, McCann, and Yamasaki and is commonly known as the Ames test (Ames et aI., 1975; McCann and Ames, 1976). The basis for the assay is as follows: Following exposure
to a mutagenic chemical, mutations are detected that reverse existing gene-inactivating mutations present in the Salmonella test strains, thereby restoring the ability of the bacteria to synthesize the essential amino acid. The bacteria carrying the reverse mutations (called revertants) are detected by their ability to grow in the absence of the amino acid required by the parental test strain. Many of the test strains have also been engineered to increase the sensitivity of the assay. These enhancements include a modification of the cell wall to be more permeable to lipophilic chemicals, inactivation of a gene involved in DNA excision repair, and addition of another gene coding for an error-prone DNA repair gene. In addition, the assay is conducted in the presence and absence of a mammalian metabolic system to increase the sensitivity of the assay to chemicals requiring metabolic activation for genotoxicity. Most
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Genetic Toxicity of Pesticides
Table 35.2 The Measurement of Genotoxic Effects in Short-Term Tests DNA binding (32 P-postIabeIing, 8-0H-dG, and others)
DNA damage in bacteria (Pol A test, rec test)
The covalent binding of a chemical to DNA is used as a measure of its
Two strains of bacteria are used that are identical except in their
reactivity and potential for genotoxicity. The DNA adducts can be
ability to repair DNA damage; one strain can repair damage
detected and quantitated by using a radiolabeled chemical,
whereas the other cannot. Both strains are exposed to the test
by labeling the adducted nucleotide after formation, through
substance, and the extent to which cells are killed is measured
an analysis for specific adducts, or by immunological techniques.
for each. If the repair-deficient strain has a greater degree of cell killing, DNA damage is assumed to have occurred.
DNA damage in yeast (mitotic recombination, mitotic gene conversion, or
Chromosomal effects in isolated cells or whole organisms (cytogenetic
mitotic crossing over)
assays or in vitro micronucleus assay)
Special strains of yeast cells are used to test for these effects.
Treated cells (or cells from treated organisms) are stained and
When the cells change color from white to either pink or red, DNA
then examined under the microscope for various chromosomal
damaging potential is indicated.
abnormalities. Lost, broken, or misarranged chromosomes or the formation of micronuclei indicate genotoxicity
Gene mutation in bacteria or fungi (Ames test, WP2 assay, yeast assays, and others)
Oncogenic transformation (transformation assays) When certain types of mammalian cells are treated in vitro
Special strains of bacteria are used that cannot grow without
with carcinogens, they undergo cancer-like transformation.
a nutritional supplement. Certain types of mutations will permit
If these cells are injected into appropriate experimental animals,
these bacteria to grow in unsupplemented media. By treating the
tumors will appear. Most frequently, transformed cells are
cells and then seeing if they can grow in unsupplemented media,
distinguished by their unusual growth patterns in culture,
mutagenicity can be measured. Distinguishing mutated bacteria from
such as abnormal piling-up and disorientation of cells.
nonmutated bacteria is not necessary using this procedure, because only mutant cells can grow and form visible colonies. DNA damage in mammalian cells (unscheduled DNA synthesis, sister chromatid exchange, and single-cell gel electrophoresis 'Comet' assay)
Micronucleus test in vivo Animals are treated with a chemical, and their red blood cells are removed, stained, and examined under the microscope. If small
Abnormal distribution of a dna marker indicates whether DNA damage
fragments of the genetic material (micronuclei) are observed,
has occurred. Microscopic examination, photographic measurements,
chromosomal damage is indicated. Normal red blood
and computerized image analysis systems are used to detect the DNA
cells will not contain any genetic material or fragments of
damage.
genetic material.
Gene mutation in mammalian cells or plants Mammalian cells (HPRT, TK, and Na/K-ATPase assays) In these
Drosophila melanogaster (sex-linked recessive lethal test for mutations: nondisjunction and heritable translocation assays for chromosomal effects)
systems, mutations that confer resistance to a poison are measured. Cells
Drosophila have a variety of "marker" traits that can be used
are first treated with a test chemical and then exposed to the poison.
to signal whether gene mutations or chromosome disturbances
Because only mutant cells can survive and grow, mutagenicity can be
have occurred. Specially "marked" male or female flies are treated
measured simply by observing the extent of growth in the poisonous
with a substance, mated, and then their offspring are observed to see
Plant cells: environment. (Tradescantia and maize waxy locus)
if they have certain specific features, such as unusual eye color or shape.
Mutations in these plants are detected by looking for color changes in the stamen hairs (Tradescantia) or pollen grains (maize). Modified from U.S. EPA (1979).
commonly, the metabolic system is a cofactor-supplemented post-mitochondrial fraction (S9) prepared from the livers of rats treated with enzyme-inducing agents such as Arochlor 1254. By using these specially engineered and rapidly growing bacteria, chemically induced mutations occurring at low frequencies (1 x 10-6 ) in tens to hundreds of millions of bacteria can be rapidly and inexpensively detected. However, the targets for reverse mutations in the test strains are very small in relationship to the bacterial genome and only a narrow range of point mutations can be detected. To increase the range of point mutations that can be detected, regula-
tory guidelines recommend that five different test strains of bacteria be used (U.S. EPA, 1998c). This assay has been shown to be moderately to highly efficient at predicting carcinogenicity with predictive values generally ranging from 0.5 to 0.8, depending on the study and the characteristics of the chemicals being tested (Brusick, 1987). However, due to fundamental differences between prokaryotic and eukaryotic organisms, these assays are not able to detect mutations induced by some types of chemical agents, for example, topoisomerase inhibitors and nucleoside analogs (U.S. EPA, 1998c).
35.3 Genotoxicity Testing of Pesticides
35.2.2 IN VITRO MUTATION ASSAY IN MAMMALIAN CELLS The in vitro mammalian cell gene mutation assay can be used to detect gene-inactivating mutations induced by chemicals (U.S. EPA, 1998d). Mouse, Chinese hamster, or human cell lines are exposed to the test chemical and mutations occurring in endogenous genes such as thymidine kinase (TK), hypoxanthineguanine phosphoribosyl transferase (HPRT), and a transgene of xanthine-guanine phosphoribosyl transferase (XPRT) are measured. Using the assay for mutations in thymidine kinase as an example, cells with a mutation converting the TK heterozygote (T K+I-) to cells lacking a functional TK allele (T K-I-) are resistant to the cytotoxic effects of the nucleotide analog, trifluorothymidine (TFr). Thymidine-proficient cells are sensitive to TFr, which inhibits cellular metabolism and halts cell division. As a result, mutant cells are able to proliferate in the presence of TFr, whereas normal cells that contain the functional TK allele are unable to grow. Similarly, cells deficient in HPRT or XPRT are selected based on their resistance to 6thioguanine or 8-azaguanine, respectively. In these assays, cells are exposed to the test chemical both in the presence and in the absence of metabolic activation for a suitable period of time, then subcultured to allow phenotypic expression prior to mutant selection with the toxic nucleotide analog. Mutant frequency is then determined, after an appropriate incubation period, by seeding known numbers of cells in a medium containing the selection agent to detect mutant cells and in a medium without the selection agent to determine the cloning efficiency. The principal advantage of this assay is that it allows rare mutations occurring in mammalian cells to be detected simply and relatively inexpensively. Moreover, because this assay measures gene-inactivating mutations in eukaryotic cells, it is capable of detecting a much broader range of mutagenic events (i.e., large deletions, recombination, etc.) than a bacterial mutation assay.
35.2.3 IN VIVO CYTOGENETIC ASSAY The in vivo chromosome aberration assay is used for the detection of structural chromosome aberrations induced by test chemicals in the bone marrow of mammals, typically rodents (U.S. EPA, 1998e). In this assay, animals are administered the test substance by an appropriate route of exposure and are sacrificed at selected times (typically 12-36 h) after treatment. Prior to sacrifice, the animals are treated with a spindledisrupting agent to arrest rapidly dividing bone marrow cells in the metaphase stage of the cell cycle. Chromosome preparations are made from the bone marrow cells and, following staining, the metaphase cells are analyzed for structural damage to the chromosomes. Some information on changes in chromosome number (aneuploidy and polyploidy) can also be obtained. Because a chromosome break can occur within most, if not all, DNA sequences throughout the genome, this assay is believed to be highly sensitive at detecting agents inducing doublestranded breaks in DNA. In previous studies, it has been shown
751
that, when tested, most human cancer-causing agents induce increased levels of chromosome aberrations in the bone marrow of rodents (Ashby and Paton, 1993). Moreover, this assay is thought to be particularly valuable in that chromosomal alterations are the underlying cause of many genetic diseases and play an important role in carcinogenesis. The in vivo aberration assay is considered particularly useful for assessing mutagenic hazards in that it allows normal in vivo metabolism, toxicokinetics (absorption, distribution, and excretion), and DNA repair processes to occur (Auletta et aI., 1993). 35.2.4 MICRONUCLEUS ASSAY The micronucleus assay is similar to the in vivo aberration assay in that both measure chromosome alterations in treated mammals and, according to most regulatory guidelines, either can be used in the initial testing (Auletta et aI., 1993; Dearfield et aI., 1991). The micronucleus assay detects chromosome breakage and loss occurring following chemical treatment. Although micronuclei can be formed in any dividing tissue of any species following treatment, for regulatory purposes the assay is almost always conducted in the bone marrow or, less frequently, the peripheral blood erythrocytes of rodents (U.S. EPA, 1998f). As a bone marrow erythroblast develops into a newly formed ribonucleic acid (RNA)-containing (polychromatic) erythrocyte, the main nucleus is extruded. In a damaged cell, the micronucleus that has been formed remains behind in the anucleate cytoplasm. Using a stain such as acridine orange that differentially stains RNA and DNA, the DNA-containing micronucleus can easily be visualized in the cytoplasm of the newly formed RNA-containing erythrocytes. An increase in the frequency of micronuclei following treatment with a test chemical indicates that an increase in chromosome damage has occurred. The assay can be performed in one of two ways: with a single dose followed by two or more sampling times or with two or more sequential doses followed by a single harvest. As with the in vivo aberration assay, this in vivo assay allows normal metabolism, toxicokinetics, and DNA repair to occur. In addition, many human and animal carcinogens when tested have shown positive results in this assay (Ashby and Paton, 1993).
35.3 GENOTOXICITY TESTING OF PESTICIDES As indicated previously, genotoxicity testing is required for the registration of new pesticides in the United States and most developed nations. Testing has also been performed for many of the pesticides that were registered prior to the current testing requirements. It should be noted, however, that often the results of these tests are considered proprietary and are not published in the public domain. Published genotoxicity test results for many pesticides and other agents evaluated by the EPA and IARC are available in both graphical and tabular forms in the Genetic Activity Profile (GAP)
752
CHAPTER 35
Genetic Toxicity of Pesticides
database (Waters et aI., 1991, 1999). This program can be downloaded without charge at www.epa.gov/gapdb. Two valuable sources of the summary results of unpublished tests on pesticides are the toxicological summaries compiled by the California Department of Pesticide Regulation (available at www.cdpr.ca.gov/docs/toxsums/toxsumlist.htm) and the toxicological evaluations performed as part of the joint meeting of the Food and Agricultural Organization panel of experts on pesticide residues in food and the environment [see FAOIWHO (1999) for a recent example]. A representative listing of specific pesticides, along with their activity in various genotoxicity tests and evaluations for carcinogenicity, is shown in Table 35.3. As is evident from the table, a variety of patterns of responses can be seen. Some agents are clearly genotoxic and carcinogenic, whereas others have shown activity in the genotoxicity assays without showing an increase in tumors in the cancer bioassays. Other pesticides have primarily exhibited negative results in short-term genotoxicity assays but have shown increases in tumors in chronic animal testing. Other agents have demonstrated no genotoxic or carcinogenic effects in in vitro or in vivo studies. Finally, many agents have given mixed or equivocal responses in genotoxicity or carcinogenicity tests. Interpretation of this latter pattern of responses is particularly challenging due to the likelihood of false positive results when many short-term assays are conducted or assays are performed under conditions (high concentrations, increased osmolality, pH, oxygen tension, etc.) that may differ significantly from those likely to be encountered in vivo. In addition, the pathological evaluation of many different tissues and organs also increases the likelihood of false positives in a chronic animal cancer bioassay. For illustration, examples of each of the preceding patterns of response will be presented.
35.4 PATTERNS OF RESPONSE 35.4.1 PESTICIDES EXHIBITING BOTH GENOTOXICITY AND CARCINOGENICITY 35.4.1.1 Ethylene Oxide
Ethy lene oxide, or epoxy ethane, is an insecticidal fumigant used for stored food products, bedding, carpets, and clothing (Gehring et aI., 1991). It is also used to sterilize heat-sensitive medical devices and as an intermediate in the synthesis of other chemicals, particularly ethylene glycol (Dellarco et aI., 1990). Structurally, it is a reactive chemical that exerts its cytotoxic effects by alkylating a broad range of critical cellular macromolecules such as DNA and proteins (Dellarco et aI., 1990; Gehring et aI., 1991). Given its ability to alkylate DNA, it is not surprising that it exhibits genotoxic effects in most genotoxicity assays. Ethylene oxide has been shown to be mutagenic in bacterial and mammalian cells,
to increase chromosome aberrations and micronuclei in the bone marrow of rodents, and to exhibit positive responses in a series of other genotoxicity assays (Dellarco et aI., 1990; IARC, 1994). In reviewing the evidence, the IARC has concluded that ethylene oxide is both an animal and a human carcinogen (IARC, 1994). In addition to affecting somatic cells, ethylene oxide is also an established germ cell mutagen, which has been shown to induce dominant lethal mutations and translocations in rodents (Dellarco et aI., 1990). Ethylene oxide is one of the few agents for which heritable risks to humans have been evaluated (Rhomberg et aI., 1990). 35.4.1.2 Ethylene Dibromide
Ethylene dibromide, EDB or 1,2-dibromoethane, has been used as a fumigant for stored grain, fruits, and vegetables (Gehring et al., 1991). It has also been used as a soil treatment for nematodes and as a scavenger in tetraethyl lead-containing gasoline. Ethylene dibromide is metabolic ally activated through both microsomal- and glutathione transferase-dependent pathways to form reactive DNA and protein-binding metabolites (Gehring et aI., 1991). EDB has been shown to be mutagenic in bacteria and mammalian cells, to bind to DNA, to induce DNA strand breakage, and to increase unscheduled DNA synthesis (U.S. EPA, 1997). Although genotoxic in the majority of in vitro tests and in vivo assays for DNA breakage, EDB has shown largely negative results in in vivo assays of chromosome damage and dominant lethal mutations (IARC, 1999b). These somewhat differing results may reflect the target organ specificity, as well as the types of DNA damage induced by this agent. EBD has been shown to exhibit carcinogenic effects in multiple animal species. However, the evidence for carcinogenic effects in humans is considered inadequate (IARC, 1999b; U.S. EPA, 1997). Both the IARC and the EPA consider EDB to be a probable human carcinogen. Similar patterns can be seen for other pesticides or "inert" ingredients such as chromium(VI), arsenic, formaldehyde, benzene, and creosote. In each of these cases, the agent is carcinogenic in either animals and/or humans and is positive in most genetic toxicity assays. t The lack of activity in a few short-term tests suggests that the agent acts through a specific genotoxic mechanism, that target organspecific effects or metabolism may be occurring, or that the genotoxicity result is in error (i.e., a false negative). Based on the strongly positive results observed, most of these types of agents have been banned for use as pesticides or are registered for use under highly restricted conditions. +Benzene and arsenic have consistently exhibited negative results in gene mutation assays but have been positive for the induction of chromosomal alterations in vivo. The critical genetic alterations in the carcinogenicity of these agents appear to be chromosomal in nature.
35.4 Patterns of Response
753
Table 35.3 Short·Term Genotoxicity Results and Evaluation of Carcinogenic Risk for Selected Pesticides Chromosomal aberrations/micronuclei
Mutation
Pesticide
Salmonella
Mammalian
(Ames test)
cells
in vitro
in vivo
lARC classification Human
Animal
carcinogenicity
carcinogenicity
Inorganic metals
+a +a, _lh
+la. +Ib
SE
LE
+
eb, ea
SE
SE
+
+a, +ah
+b, +a
SE
SE
_a, eb
NR
NR
ab
ND
lE
+Iah
+1'
ND
lE
Arsenic compounds Cadmium chloride Chromium(Vl)
+
compounds Carbamates
Propoxur* +a, +lh
Carbaryl Aldicarb
+, +lh
Chlorinated hydrocarbon insecticides
0
0
lE
SE
-1
0 ea,_lah
0 +a, (+)ah
lE
SE
lE
SE
0
+lah
-lb, +Ia
lE
LE
NR
NR
(+)
+a
0
lE
lE
0 +lah
+a, _ b
ND
lE
ND
lE
0
+a, +lb _la
NR
NR
0
-la
ND
lE
+a, -lah
lE
SE
0
lE
ND
ND
ESL
Chlordane Heptachlor DDT
Aldrin Endosulfan* Endrin Pyrethroids
Deltamethrin Fenvalerate
0 -I
Cypermethrin Permethrin Organophosphate insecticides
Dichlorvos
+
+1
Parathion Methyl parathion Malathion
0
Diazinon
e
Chlorpyrifos* lsazofos
+
+ -I
-I
-I
0
0
0
+a, +bh, +Iah, +Ib ea, -Iah
e _a, -lb b
ND
lE
NR
NR
NR
NR
NR
NR
ND
SE
Fungicides
Captafol Pentachlorophenol Thiram
+
e
Ziram
+
0
ortho-Phenylphenol*
+1
Chlorothalonil Hexachlorobenzene*
0 +
+1
lE
SE
lE
lE
+Ia, +Ib, +Iah
ND
LE
-, +Ib a
lE
LE
lE
SE
0
lE
SE
a
ND
SE
+Ia, +lah
+Ib
lE
SE
_Ib
lE
SE
-la _Iah
+b, +Ia
ND
LE
-1' _Ib
ND
LE
lE
LE
+a
lE
LE
-1', _Iah, _Ib, +Ibh _a, + Ib, _Ibh
1,4-Dichlorobenzene* Propylene oxide
+1' +1',+lah ea ea _ah
0 +Iah +b, +Ia
Herbicides
Atrazine
-I
Monuron
e
Picloram
0 -I
Simazine Trifluralin
0 eah
(continues)
754
CHAPTER 35 Genetic Toxicity of Pesticides
Table 35.3 (continued) Mutation
Pesticide
Salmonella
Mammalian
(Ames test)
cells
MCPA 2,4-D
0
Chromosomal aberrations/micronuclei in vitro
in vivo
rARC classification Human
Animal
carcinogenicity
carcinogenicity
+1 ab
-1', _b,_ ah ab b
lE
ND
LE
lE lE
+1
0
0
lE
Bentazon*
-I
0
-lb
NR
NR
2,4,5-T
0
0
-lb,+1'
LE
lE
Amitrole
e
-1'
lE
SE
+
+a, +Ib, _lab
_a, _lb
lE
SE
+Ih
LE
SE
lE
SE
Methyl chloride
+
Fumigants and nematocides Acrylonitrile
+
Ethylene oxide*
+
+
Ethylene dibromide*
+
+
+a, +Ib
+a, +b -la, _b
+
+a, +ah
_ab, _a, _b, +bhf
LE
SE
+ e
ea
+b, -1"
lE
LE
+a
SE
0
-lab, eb
+b, +1"' - a -
lE
Carbon tetrachloride
lE
SE
Tetrachloroethylene
0
0
-lab
lE
LE
1,3 Dichloropropene
e
lE
SE
Formaldehyde
+
Methyl bromide*
+
DBCP
+
+
+
Solvents and others Xylene
-I", -lab
lE
lE
Benzene
+a
+a, +b
SE
SE
Piperonyl butoxide
0
0
ND
lE
a, chromosomal aberrations; b, micronucleus; h, human cells; s, spermatogonia; (+), weakly positive; e, equivocal/inconclusive; 0, no test results were located; + I, positive in one study; -I, negative in one study; +, positive in more than one study or the majority of studies; -, negative in more than one study or the majority of studies; f, micronucleus formation was positive in buccal mucosal cells in humans whereas it was negative in peripheral blood Iymphocytes, possibly due to the high reactivity of formaldehyde at the primary site of exposure; lE, inadequate evidence for carcinogenicity; LE, limited evidence for carcinogenicity; SE, sufficient evidence for carcinogenicity; ESL, evidence suggesting lack of carcinogenicity; ND, no adequate data were available; NR, not reviewed by the rARe. *See text for additional details. This table was compiled primarily from five sources: (1) the rARC Monographs on the Evaluation of Carcinogenic Risks in Humans; (2) the Environmental Health Criteria series published by the International Programme on Chemical Safety; (3) the toxicological evaluations performed as part of the Joint meeting of the FAO panel of experts on pesticide residues in food and the environment; (4) the Genetic Activity Profile (GAP) database generated jointly by the EPA and rARC; and (5) the toxicological data review summaries prepared by the California Department of Pesticide Regulation.
35.4.2 PESTICIDES EXHIBITING GENOTOXICITY WITH LIMITED OR NO EVIDENCE OF CARCINOGENICITY 35.4.2.1 Methyl Bromide Methyl bromide, or bromomethane, has been widely used as a fumigant for control of insects, nematodes, fungi, and weeds (Gehring et aI., 1991; IPCS, 1985). Although methyl bromide has been shown to react with both DNA and proteins, its mechanism for toxicity remains to be elucidated (IPCS, 1995). Methyl bromide has been shown to be genotoxic in most short-term genotoxicity tests (IARC, 1999c; IPCS, 1995): It induced mutations in bacteria and mammalian cells, increased the incidence of micronuclei in vivo in mouse and rat bone marrow erythro-
cytes, and was shown to bind covalently to the DNA in several rat and mouse organs. In contrast, methyl bromide has produced mixed, largely negative responses in chronic animal bioassays. In a short 13week study in which methyl bromide was administered by oral gavage, it was reported to produce squamous cell carcinomas of the forestomach (IARC, 1999c; D.S. EPA, 1990). However, this result was questioned by other investigators, and, upon reexamination of histological slides, a group of National Toxicology Program pathologists concluded that the lesions were hyperplasia and inflammation rather than neoplasia (U.S. EPA, 1990). In inhalation studies, the most relevant route of human exposure, methyl bromide was reported to be largely negative, although there was some limited evidence for tumorigenicity in various tissues. According to the IARC, no significant increase in tumors was observed in two inhalation studies in mice and one in rats (IARC, 1999c). In another rat inhalation study, a signif-
35.4 Patterns of Response
icant increase in pituitary gland adenomas was seen in males treated at the highest dose. However, a detailed examination of two of the inhalation studies described as negative led some reviewers to suggest that methyl bromide was capable of inducing tumors in some tissues (CDPR, 1999b). Based on its evaluation of the literature, the IARC concluded that there is limited evidence in experimental animals for the carcinogenicity of this agent (lARC, 1999c), whereas the EPA considered the data inadequate to reach any conclusion (U.S. EPA, 1990). Both the IARC and the EPA stated that there is inadequate evidence to make conclusions about the carcinogenicity of methyl bromide in humans (IARC, 1999c; U.S. EPA, 1990). As indicated previously, methyl bromide is genotoxic in most in vitro and in vivo assays. Although there is some evidence for the carcinogenicity of methyl bromide, it has not exhibited consistent carcinogenic effects in most studies. The reason for the discrepancy between the short-tenn tests and the animal bioassay results is not clear. These negative test results could be false negatives, reflecting inadequacies of the animal bioassays. However, several bioassays have been conducted with similar results and no carcinogenic effects were seen even in the comprehensive mouse bioassay conducted by the National Toxicology Program (NTP, 1992). Alternatively, methyl bromide may alkyl ate DNA in vivo at sites that are readily repaired or lead directly to celllethality rather than heritable mutations. 35.4.3 PESTICIDES EXHIBITING CARCINOGENICITY WITHOUT APPRECIABLE GENOTOXICITY 35.4.3.1 Propoxur
Propoxur, or Baygon, is an important carbamate insecticide used primarily against household insects and pests of domestic animals. It is considered among the top 10 most widely used home and garden pesticides in the United States (Grossman, 1995). Similar to other carbamate insecticides, propoxur inhibits acetylcholinesterase, an enzyme involved in neurotransmission, producing neurotoxic effects in insects and nontarget organisms. Propoxur has yielded negative results in the majority of short-term genotoxicity tests that have been conducted (FAOIWHO, 1990). It was negative in bacterial and mammalian mutation assays and in bacterial DNA repair assays. It was reported to be negative in most chromosome aberration and micronucleus assays in vitro and in vivo (FAOIWHO, 1990), although two positive studies have recently been published (Agrawal and Mehrotra, 1997; Wei et aI., 1997). In chronic studies, no evidence of carcinogenic effects was seen in mice treated with propoxur for 24 months or hamsters treated for 53 weeks (FAOIWHO, 1990). However, in a series of studies conducted in rats, highly significant increases in hyperplasia and bladder tumors were seen at high doses of propoxur. No hyperplastic effects in the bladder were seen in short-tenn studies employing mice, dogs, or monkeys, whereas effects
755
were seen in short-tenn studies in the bladders of SpragueDawley rats. Dietary studies have indicated that, in addition to high doses, the urothelial effects of propoxur are dependent on high urinary pH. The carcinogenic effects of propoxur in rats have been proposed to be due to chronic mitogenic stimulation of propoxur or a metabolite on the urothelium rather than from a direct genotoxic or mutagenic effect (Cohen et aI., 1994). 35.4.3.2 Hexachlorobenzene
Historically, hexachlorobenzene (HCB) was commonly used as a seed treatment for prevention of fungal growth on crops such as wheat, barley, oats, and rye (IPCS, 1997). Concern for human health and the environment resulted in its discontinued use as a pesticide in many countries during the 1970s. Hexachlorobenzene is currently found as an unintentional by-product in several high-volume chlorinated solvents (carbon tetrachloride, trichloroethylene, and perchloroethylene) and in various pesticides, including pentachloronitrobenzene, chlorothalonil, dimethyl 2,3,5,6-tetrachlorotereohthalate (DCPA), picloram, and pentachlorophenol (ATSDR, 1997). In general, studies investigating the genotoxicity of HCB have indicated that it exhibits weak or no genotoxic activity (Brusick, 1986; Gorski et aI., 1986; IPCS, 1997). In most studies, hexachlorobenzene exhibited no detectable mutagenic activity in Salmonella either with or without microsomal activation. No increase in structural chromosome aberrations was seen in Chinese hamster lung cells (Ishidate et aI., 1988). Canonero and associates evaluated the in vitro genotoxicity of HCB in primary cultures of rat and human hepatocytes (Canonero et aI., 1997). An induction of micronuclei but not DNA strand breaks was seen in rat hepatocytes treated with hexachlorobenzene. In the studies with human hepatocytes, the authors reported that hexachlorobenzene induced a weak but significant increase in the frequency of both DNA breaks and micronuclei. Low levels of DNA binding were seen following the in vivo treatment of rats with hexachlorobenzene (Gopalaswamy and Nair, 1992). Additionally, no increase in sister chromatid exchanges (SCEs) in the bone marrow of male mice or DNA fragmentation in the liver of rats was observed in hexachlorobenzene-treated animals (Gorski et aI., 1986). Hexachlorobenzene also failed to induce dominant lethal mutations in male rats (Simon et aI., 1979). In contrast with the largely negative results in the genotoxicity studies, hexachlorobenzene exhibited carcinogenic effects in a series of animal studies, increasing the incidence of tumors in rats, hamsters, and mice (lARC, 1987b; IPCS, 1997; U.S. EPA, 1985, 1996). Increased tumor fonnation was seen in the liver and kidney as well as the adrenal, parathyroid, and thyroid glands of the treated animals. To date, the mechanisms underlying carcinogenesis in these organs remain unclear. Several theories have been proposed to explain the basis for certain tumors induced by hexachlorobenzene. For example, it has been proposed that the liver tumors occur as a secondary effect resulting from chronic toxicity to this organ (Carthew and Smith, 1994). It has been postulated that the male kidney tumors were due
756
CHAPTER 35
Genetic Toxicity of Pesticides
to an accumulation of the male rat-specific protein alpha 2uglobulin in the proximal renal tubular cells, resulting in a sustained cell proliferation and eventually neoplasia in this organ (Bouthillier et aI., 1991). Finally, others have proposed that the thyroid tumors were the result of a chronic stimulation of cell proliferation in the thyroid gland due to a chronic imbalance in thyroid hormones resulting from an induction of glucuronosyl transferases by hexachlorobenzene (DFG, 1998). All of these theories indicate that hexachlorobenzene exerts its carcinogenic effects through indirect or "nongenotoxic" mechanisms. Assuming that these mechanisms are correct, the difference in the observed genotoxicity and carcinogenicity results would be expected. Following a review of the data, the IARC and the EPA have determined that there was sufficient evidence to conclude that HCB induces cancer in laboratory animals (lARC, 1987b; U.S. EPA, 1996). The evidence in humans is inadequate to draw definite conclusions. However, for regulatory purposes, the EPA considers hexachlorobenzene to be a probable human carcinogen (U.S. EPA, 1996), whereas the IARC considers it to be a possible human carcinogen (IARC, 1987b).
35.4.4 NONGENOTOXIC AGENTS WITHOUT EVIDENCE OF CARCINOGENICITY 35.4.4.1 Endosulfan Endosulfan, or thiodan, is a chlorinated insecticide used on a wide variety of food and non-food crops, including grapes, cantaloupes, lettuce, tomatoes, alfalfa, and cotton. Although a few positive responses have been reported in short-term tests (Smith, 1991), endosulfan is generally viewed by regulatory bodies as being nongenotoxic (CDFA, 1988; FAOIWHO, 1999). Endosulfan has primarily exhibited negative results in both bacterial and mammalian cell gene mutation assays. It was also negative in inducing chromosome aberrations or micronuclei in vitro as well as in vivo. In addition, it has been reported to be negative in other genotoxicity assays. Endosulfan did not exhibit carcinogenic effects in chronic bioassays conducted using mice or rats (FAOIWHO, 1999). Epidemiological studies of cancer in humans have not been conducted.
35.4.4.2 Chlorpyrifos Chlorpyrifos, or Dursban, is a broad-spectrum organophosphate insecticide with widespread usage on food commodities, turf, and ornamental plants. It has been commonly used indoors and for structural pest control. It is one of the most widely used pesticides in the United States and has been one of the top five insecticides used in residential settings (U.S. EPA, 1999). In common with other organophosphate insecticides, upon bioactivation, chlorpyrifos inhibits acetylcholinesterase, an enzyme involved in neurotransmission, producing neurotoxic effects in insects and nontarget organisms. Consequently, genotoxic effects would not be expected nor are they seen (CDPR, 1999a; U.S. EPA, 1999). Chlorpyrifos did not induce gene mutations in either bacterial and mammalian systems, although it was reported to induce slight increases in genetic alterations in yeast
as well as DNA damage in bacteria. No increase in chromosome aberrations was seen in an in vitro study using rat lymphocytes or in two in vivo studies evaluating micronuclei in the mouse bone marrow. It was ineffective at inducing unscheduled DNA synthesis in isolated rat hepatocytes. Chlorpyrifos was evaluated for carcinogenic potential in both rats and mice with no evidence of carcinogenicity (CDPR, 1999a; U.S. EPA, 1999).
35.4.4.3 Bentazon Bentazon, 3-( 1-methylethy 1)-1 H -2,1 ,3-benzothiadiazin-4(3 H)one-2,2-dioxide, is a herbicide used in agriculture for control of broadleaf weeds in crops such as soy beans, rice, corn, peanuts, and lima beans (U.S. EPA, 1998g). As summarized from EPA reports (U.S. EPA, 1998b, g), Bentazon is not chemically reactive and no highly reactive species have been identified during its metabolism. Bentazon was negative in bacterial mutation assays, in a mammalian cell assay, in the unscheduled DNA synthesis assay, and in the mouse micronucleus assay in vivo. In chronic animal bioassays, no increases in tumors were seen in the rat. A slight dose-related increase in hepatocellular tumors was seen in the mouse studies. However, upon reexamination, it was concluded that the incidence did not differ significantly from the controls. In its evaluation of the toxicity of Bentazon, the EPA concluded that bentazon was essentially noncarcinogenic in animals and was not likely to cause cancer in humans.
35.4.5 PESTICIDES EXHIBITING MIXED RESULTS IN GENOTOXICITY OR CANCER TESTS
35.4.5.1 ortho- Phenyl phenol ortho-Phenylphenol (OPP) and its sodium salt, sodium o-phenylphenate (SOPP), are broad-spectrum fungicides and disinfectants with widespread agricultural, industrial, and domestic usage. OPP has historically been among the most widely used home and garden pesticides (Grossman, 1995). Investigations into the genotoxic effects of SOPP and OPP have indicated that these compounds are inactive or weakly active in bacterial mutation assays (NTP, 1986). Some evidence for the mutagenicity of OPP has been seen in mammalian cell assays. A weak increase in mutations was seen at the TK locus in treated CHO cells (NTP, 1986), whereas a strong increase in ouabain-resistant mutants was reported to occur in an ultraviolet-sensitive human Rsa cell line following treatment with OPP (Suzuki et aI., 1985). Negative results were also observed when measuring unscheduled DNA synthesis in rat hepatocytes following exposure to SOPP (Reitz et aI., 1983). In cytogenetic studies, several reports indicate that OPP and its metabolite phenylhydroquinone have induced sister chromatid exchanges and structural chromosomal aberrations in CHO cells in the presence of exogenous metabolic activation (NTP, 1986; Tayama and Nakagawa, 1991; Tayama et al., 1989; Tayama-Nawai et al., 1984), whereas others have reported negative or ambiguous results (Ishidate, 1988;
35.4 Patterns of Response
NTP, 1986). Phenylhydroquinone was also shown to induce chromosome-containing micronuclei upon prostaglandin[H] synthase-mediated activation in V79 cells (Lambert and Eastmond,1994). Following the in vivo administration of radiolabeled OPP and SOPP to male F344 rats, no increases in the covalent binding of these compounds to rat bladder DNA were observed using either liquid scintillation counting (Reitz et al., 1983) or a highly sensitive accelerator mass spectrometric technique (Kwok and Eastmond, 1997). Binding to bladder proteins was seen in both studies. Contradictory results have been reported for DNA binding using the 32p postlabeling technique with one group reporting detectable OPP-derived adducts (Ushiyama et aI., 1992) whereas another, focusing on adduct formation in the target urothelial cells, reported negative results (Smith et aI., 1998). A modest increase in DNA breakage in the bladder was detected in rats (Morimoto et aI., 1989) and mice (Sasaki et aI., 1997) following treatment with OPP or SOPP. In addition, a significant increase in micronucleated bladder cells was reported in rats administered a high dose of OPP in the diet (Tadi-Uppala et aI., 1996). OPP and SOPP have been tested for carcinogenicity in both mice and rats by administration in the diet. Increases in bladder tumors were seen in multiple rat studies following treatment with OPP and SOPP (CDPR, 1997; IARC, 1999d). SOPP appears to be more potent and consistent in inducing carcinogenic effects, and it has been proposed that urinary pH plays an important role in the bioactivation and carcinogenesis of these compounds (Fujii et aI., 1987; Kwok and Eastmond, 1997). The effects appear to be specific to the rat as little evidence of carcinogenicity was observed in chronically treated mice (IARC, 1999d), and bladder toxicity was not seen in short-term studies in mice, guinea pigs, hamsters, and dogs (Co see et aI., 1992; Hasegawa et aI., 1990). Upon review of the data, the IARC concluded that OPP was not classifiable as to its carcinogenicity to humans and that SOPP was possibly carcinogenic to humans (IARC, 1999d). The mechanisms underlying the carcinogenic effects of OPP remain to be fully elucidated. It has been proposed that OPP acts as a bladder carcinogen in rats by inducing cytotoxicity and hyperplasia without directly binding to DNA (Smith et aI., 1998). In this case, the genotoxicity may be indirect, occurring through the formation of oxygen radicals, through an enhancement of spontaneous mutations, or through an interaction with protein targets (Appel, 2000; Kwok and Eastmond, 1997). The inconsistent results seen in the short-term tests may, in part, be a reflection of this indirect mechanism of genotoxicity. 35.4.5.2 1,4-Dichlorobenzene l,4-Dichlorobenzene, or para-dichlorobenzene (p-DCB), is commonly used to control moths, molds, and mildew, and as a bathroom deodorizer. p-DCB is also used as an intermediate in the synthesis of polyphenylene sulfide (PPS) resin. The genotoxicity of p-DCB has been investigated with mixed, largely negative results (IARC, 1999a). p-DCB was not mutagenic in bacteria or mammalian cells in vitro but did exhibit
757
some evidence of DNA damage and mutagenicity in yeast. pDCB produced mixed results in in vitro cytogenetic assays with both positive and negative reports for micronuclei and sister chromatid exchanges. It was negative in inducing DNA strand breaks and chromosome aberrations in vitro. p-DCB failed to exhibit genotoxic effects in vivo, exhibiting negative responses in unscheduled DNA synthesis, in the chromosome aberration assay, in the dominant lethal assay, and in the in vivo micronucleus assay. It was reported as positive in one DNA strand breakage assay and in one in vivo micronucleus assay. p-DCB bound to DNA in the liver, lung, and kidney of mice but not in that of male rats (IARC, 1999a). It also induced DNA damage in the liver and spleen but not in the kidney, lung, or bone marrow of mice. The IARC stated that no conclusion could be drawn from the few data on genotoxicity in vivo (IARC, 1999a). In contrast to the negative genotoxicity results, p-DCB induced carcinogenic effects in both rats and mice. Following oral administration, p-DCB increased the incidence of liver tumors in male and female mice as well as the incidence of renal carcinomas in male rats (IARC, 1999a). In evaluating the significance of these tumors, the IARC concluded that the evidence did not support a mechanism of renal cell tumor formation that involved a direct interaction between p-DCB or its metabolites with DNA. The male kidney tumors induced by p-DCB were due to an accumulation of the male rat-specific protein alpha 2u-globulin in the proximal renal tubular cells that eventually resulted in neoplasia in this organ. This mechanism is widely accepted as not being relevant to humans (U.S. EPA, 1991; IARC, 1999a; Rice et al., 1999a). However, the IARC Working Group had more concern for the liver tumors that were seen at a high incidence in the male and female mice. Because p-DCB was reported to cause DNA damage in the liver and spleen of mice and bound weakly to DNA, the tumors in the liver were thought to be potentially relevant to humans. IARC concluded that p-DCB was an animal carcinogen and possibly carcinogenic to humans (lARC, 1999a). As illustrated in the preceding examples, different patterns of genotoxic and carcinogenic effects can be seen in short-term tests and in animal bioassays. In many instances, the outcome of the studies and interpretation of their relevance to humans is relatively straightforward, indicating that these agents pose or do not pose significant carcinogenic risks to humans. However, in other cases, the interpretation of the results can be quite challenging. In almost all cases, scientists and regulators rely on a weight-of-the-evidence approach, where the number, consistency, and quality of the studies is combined with mechanistic, structure-activity, and other information to reach conclusions about the genotoxicity and likely human carcinogenicity of the agent. In addition to the short-term tests and animal results, information about the genotoxic effects of the pesticide in humans can contribute significantly to the risk assessment process.
758
CHAPTER 35
Genetic Toxicity of Pesticides
35.5 HUMAN BIOMONITORING To identify pesticides and other agents capable of inducing genotoxicity in humans and to identify groups at elevated risk for cancer or other genetic diseases, biological markers of exposure and effect have been developed to measure genetic changes in exposed humans (Albertini and Hayes, 1997; Albertini et aI., 2000; Sorsa et aI., 1992; Tucker et aI., 1997; Wild and Pisani, 1997). These biomarkers range from early premutagenic lesions such as covalent adducts between the chemical and DNA to heritable mutations in endogenous genes such as HPRT. Although these studies have primarily been conducted using somatic cells, a few have been performed using germ cells in which chromosomal changes in human sperm have been monitored. Among the most commonly used biomarkers is the measurement of structural and numerical alterations in lymphocyte chromosomes. In this assay, the frequencies of chromosome changes occurring in metaphase preparations of stimulated peripheral blood lymphocytes from individuals in an exposed group are measured and compared with those of an appropriate control. Increased frequencies of genetic alterations are believed to indicate that an exposure has occurred that is biologically significant and mechanistically related to cancer and other genetic diseases (Sorsa et aI., 1992). Consistent with this, recent studies have shown that individuals with elevated frequencies of structural chromosomal aberrations in their peripheral blood lymphocytes are at increased risk for the development of cancer (Bonassi et aI., 1995; Hagmar et aI., 1994, 1998). It should be noted that for one frequently measured endpoint, sister chromatid exchanges, such an association was not seen (Hagmar et aI., 1994,1998). A considerable number of studies have been conducted using various biomarkers to measure genetic alterations in the cells of pesticide-exposed workers. A list of genetic biomarker studies obtained primarily from a search of MED LINE (and references cited therein) is shown in Table 35.4. Although these studies represent only a fraction of the studies that have been conducted, the results and patterns of response are probably representative of those commonly seen in pesticide biomonitoring studies. As can be seen from the table, numerous reports from many countries have been published on genotoxic effects in pesticideexposed workers. In many of these, higher frequencies of genotoxic effects have been seen in the exposed workers. However, most studies have been conducted on agricultural workers who have been exposed to many different pesticides. As a result, it is difficult to identify the actual genotoxic agent involved. For example, in the studies conducted in southeast India by Rupa and associates, the cotton field applicators reported having used 11 different pesticides in the period preceding the study (Rupa et aI., 1989b). Even in cases where the exposed workers were exposed primarily to a single pesticide, the reported outcome may be influenced by other confounding factors such as tobacco smoking, age, exposure to solvents, inert ingredients, etc.
As is also apparent from the table, studies have been performed for only a small portion of the thousands of pesticides currently being used. In addition to the paucity of information on most pesticides, interpreting the results of biomonitoring studies such as these and their significance for workers exposed at lower levels or the general public exposed at much lower levels can be difficult. For example, ethylene oxide has exhibited positive responses in the majority of biomonitoring studies and endpoints measured. This is consistent with the known reactivity of this agent and its results in the short-term genotoxicity tests. This would indicate that, at high exposure levels, ethylene oxide poses a genotoxic and carcinogenic risk. However, these studies provide little information about the risk at lower exposure levels, requiring an extrapolation of risk to be made from high exposures to lower exposures. In contrast, negative results were seen in two biomonitoring studies of ethylene dibromide, an agent that yielded positive results in most short-term tests and was carcinogenic in animals. Although one might interpret these results as indicating that EDB is not genotoxic in humans, this conclusion could easily be in error. In this case, the negative results could simply be due to a combination of low exposures and a limited sample size. Quantitative measures of pesticide exposure are infrequently performed in these types of studies. Other results, such as those reported for dichlorodiphenyltrichoroethane (DDT), dimethoate, deltamethrin, and cypermethrin, are also challenging to interpret. Based on the short-term test results, one would not expect these agents to be genotoxic. For example, DDT was negative in 133 out of the143 short-term genotoxicity tests listed in the EPAJIARC Genetic Activity Profile database. This suggests that the positive results seen in these types of biomonitoring studies might be due to other factors such as solvent exposure, tobacco use, etc. (Petrelli et aI., 1993) or may simply be false positives. However, the overall number of positive studies in Table 35.4 far exceeds a reasonable estimate of false positives and indicates that pesticide exposure is frequently associated with genotoxic effects in exposed workers. Although the majority of biomonitoriing studies have been conducted using somatic cells, a small number of studies have been conducted to measure effects in germ cells. Interestingly, positive effects have been reported in three of the four studies conducted to date. Significant increases in aneuploid sperm were seen in agricultural workers exposed to 1,2-dibromo-3chloropropane (DBCP) (Kapp et aI., 1979), in Chinese factory workers exposed to organophosphates (Padungtod et aI., 1999) and in Indian applicators and sprayers exposed to a variety of pesticides (predominantly organophosphate insecticides) (Rupa et aI., 1997). Moreover, increases in breakage/exchanges affecting the lcen-lql2 region of chromosome 1 were also detected in the sperm of the Indian cotton field workers (Rupa et aI., 1997). Notably in earlier studies by Rupa and associates, the Indian group of applicators involved in the sperm and lymphocyte aberration studies had previously been reported to exhibit significant decreases in reproductive performance (fertility, pregnancy loss, and birth anomalies) (Rupa et aI., 1991b). These initial reports indicate that exposure to certain pesticides can
35.5 Human Biomonitoring
759
Table 35.4 Summary of Results of Genotoxicity Studies of Pesticide-Exposed Workers Study group
Location
Pesticide
Endpoint
Result
Reference Linnainmaa, 1983
Foliage sprayers
Finland
2,4-D and MCPA
SCE
Negative
Workers
United States
DBCP
Sperm aneuploidy
Positive
Kapp et aI., 1979
Workers in insecticide plants
Brazil
DDT
Cs aberrationsa
Positive
Rabello et al., 1975
Sprayers
Syria
Deltamethrin and
Cs aberrations
Positive
Mohammad et aI., 1995
SCE
Positive
Larripa et aI., 1983 Steenland et aI., 1986
cypermethrin Brazil
Dimethoate
Papaya workers
Hawaii
Ethylene dibromide
SCE
Negative
Papaya workers
Hawaii
Ethylene dibromide
Cs aberrations
Negative
Steenland et aI., 1986
Pesticide sprayers
United States
Ethylene dibromide
SCE
Negative
Steenland et aI., 1985
Accidental exposure of firefighters
Pesticide sprayers
United States
Ethylene dibromide
Cs aberrations
Negative
Steenland et aI., 1985
Factory workers
Sweden
Ethylene oxide
SCE
Negative
Hogstedt et aI., 1983
Factory workers
Sweden
Ethylene oxide
Cs aberrations
Positive
Hogstedt et aI., 1983
Factory workers
Sweden
Ethylene oxide
Micronuclei
Negative
Hogstedt et al., 1983
Factory workers
Sweden
Ethylene oxide
MicronucIeib
Positive
Hogstedt et aI., 1983
Sanitary workers
Italy
Ethylene oxide
Cs aberrations
Positive
Sarto et aI., 1984
Sanitary workers
Italy
Ethylene oxide
SCE
Positive
Sarto et aI., 1984
Sterilizer operators
United States
Ethylene oxide
Cs aberrations
NegativeC
Galloway et aI., 1986
Malathion workers
United States
Malathion
Micronuclei
Negative
Titenko-Holland et aI., 1997
Workers (production)
Czechoslovakia
Mancozeb-
Cs aberrations
Positive
lablonicbi et al., 1989
SCE
Positive
lablonicka et aI., 1989
Calvert et aI., 1998
containing fungicide Novozir Mn80 Workers (production)
Czechoslovakia
Mancozebcontaining fungicide Novozir Mn80
Fumigation workers
United States
Methyl bromide
HPRT mutations
Negative
Fumigation workers
United States
Methyl bromide
Micronucleid
Equivocal
Calvert et aI., 1998
Pesticide plant workers
Brazil
Methyl parathion
Cs aberrations
Negative
de Cassia Stocco et aI., 1982
Pesticide-preparing workers
Hungary
Monochlorinated benzene
HPRT mutation
Negative
Major et aI., 1992
Patients (attempted suicide or
Hungary
Organophosphates
Cs aberrationse
Positive
van Bao et aI., 1974
exposed during work) Fumigant applicators
United States
Phosphine
Cs aberrations
Positive
Garry et aI., 1989
Fumigant applicators
United States
Phosphine
SCE
Negative
Garry et aI., 1989
Pesticide applicators
United States
Phosphine
Cs rearrangements
Positive
Garry et aI., 1992
Pesticide sprayers
Hungary
Pyrethroids
Cs aberrations!
Positive
Nehez et aI., 1988
Workers (fitters, packers,
Former Soviet
Zineb
Cs aberrations
Positive
Pilinskaya, 1974
Ziram
Cs aberrations
Positive
Pilinskaya, 1970 Yoderetal.,1973
truck drivers) Store workers and packers
Union Former Soviet Union
Pesticide applicators
United States
Herbicides
Cs aberrations
Positive
Pesticide applicators
United States
Insecticides
Cs aberrations
Positive
Yoder et aI., 1973
Farmers
Denmark
Fungicides
Aneuploid sperm
Negative
Harkonen et aI., 1999
Pesticide applicators
United States
Pesticides
Cs aberrations
Positive
Yoder et aI., 1973
Sprayers
New Zealand
Pesticides
SCE
Negative
Crossen and Morgan, 1978
Pesticide workers
Sweden
Pesticides
Cs aberrations
Negative
Hogstedt et aI., 1980
Exposed workers
Hungary
Pesticides
Cs aberrations
Positive
Nehez et aI., 1981
Agricultural workers
Former Soviet
Pesticides
Cs aberrations
Positive
Volnjanskaya, 198 I
Union (continues)
760
CHAPTER 35
Genetic Toxicity of Pesticides
Table 35.4 (continued) Study group
Location
Pesticide
Endpoint
Result
Reference
Floriculturists
Argentina
Pesticides
SCE
Positive
Dulout et aI., 1985
Floriculturists
Argentina
Pesticides
Cs aberrationsg
Negative
Dulout et aI., 1985
Greenhouse pesticide sprayers
Hungary
Pesticides
Cs aberrations!
Positive
Desi et aI., 1986
Ornamental plant breeders
Argentina
Pesticides
Cs aberrations
Negative
Dulout et aI., 1987
Pesticide workers
Mexico
Pesticides
Cs aberrations
Positive
Gayon et aI., 1987
Mixers and field sprayers
Hungary
Pesticides
Cs aberrations
Positive
PaIdy et aI., 1987
Pesticide sprayers in vineyards
India
Pesticides
Cs aberrations
Positive
Rita et aI., 1987
Pesticide sprayers
Hungary
Pesticides
Cs aberrations!
Positive
Nehez et aI., 1988
Vegetable garden workers
India
Pesticides
Cs aberrations
Positive
Rupa et aI., 1988
Fumigant applicators
United States
Pesticides
Cs aberrations
Positive
Garry et aI., 1989
Pesticide sprayers
India
Pesticides
Cs aberrations
Positive
Rupa et aI., 1989b
Pesticide mixers and sprayers
India
Pesticides
Cs aberrations
Positive
Rupa et aI., 1989a
Pesticide applicators
Canada
Pesticides
Micronuclei h
Positive
San et aI., 1989
Agricultural workers
Spain
Pesticides
SCE
Negative
Carbonell et aI., 1990
Workers in flower industry
Italy
Pesticides
Cs aberrations
Positive
De Ferrari et aI., 1991
Workers in flower industry
Italy
Pesticides
SCE
Positive
De Ferrari et aI., 1991
Cotton field workers
India
Pesticides
Cs aberrations
Positive
Rupa et aI., 1991a
Pesticide applicators
India
Pesticides
SCE
Positive
Rupa et aI., 1991c
Pesticide applicators
United States
Pesticides
Cs rearrangements
Positive
Garry et aI., 1992
Workers in plastic greenhouses
Greece
Pesticides
Cs aberrations
Positive
Kourakis et aI., 1992
Pesticide workers
Mexico
Pesticides
SCE
Negative
G6mez-Arroyo et aI., 1992
Floriculturists
Italy
Pesticides
Micronuclei
Positive
Bolognesi et aI., 1993a
Floriculturists
Italy
Pesticides
Micronuclei
Positive
Bolognesi et al.. 1993b
Agricultural workers
Spain
Pesticides
Cs aberrations
Positive
Carbonell et al.. 1993
Agricultural workers
Spain
Pesticides
SCE
Negative
Carbonell et aI., 1993
Pesticide packers
Egypt
Pesticides
Cs aberrations
Positive
Anwar,1994
Pesticide packers
Egypt
Pesticides
SCE
Negative
Anwar,1994
Farm workers
Spain
Pesticides
Cs aberrations
Positive
Carbonell et aI., 1995
Greenhouse sprayers
Scandinavia
Pesticides
SCE
Negative
Lander and Ronne, 1995
Pesticide applicators
India
Pesticides
Cs aberrations i
Positive
Rupa et aI., 1995
Dealers and controllers
Syria
Pesticides
Cs aberrations
Positive
Mohammad et aI., 1995
Farmers
Colombia
Pesticides
SCE
Negative
Hoyos et aI., 1996
Farmers
Colombia
Pesticides
Cs aberrations
Negative
Hoyos et aI., 1996
Pesticide sprayers
Greece
Pesticides
SCE
Negative
Kourakis et al., 1996
Pesticide sprayers
Hungary
Pesticides
Cs aberrations
Positive
Nehez and Desi, 1996
Farmers
Italy
Pesticides
SCE
Negative
Pasquini et aI., 1996
Farmers
Italy
Pesticides
Micronuclei
Positive
Pasquini et aI., 1996
Greenhouse floriculturists
Italy
Pesticides
DNA adducts
Positive
Peluso et aI., 1996
Greenhouse floriculturists
Italy
Pesticides
Cs aberrations
Negative
Scarpato et aI., 1996
Greenhouse floriculturists
Italy
Pesticides
Micronuclei
Negative
Scarpato et aI., 1996
SCE
Negative
Scarpato et aI., 1996
Positive
Joksic et aI., 1997
Greenhouse floriculturists
Italy
Pesticides
Cs
aberrations j
Vineyard growers
Yugoslavia
Pesticides
Pesticide sprayers
France
Pesticides
DNA damage
Increase
Lebailly et aI., 1998
Mixers and applicators
India
Pesticides
Cs aberrations in
Positive
Rupa et aI., 1997
sperm i (continues)
35.6 Genotoxicity and Risk Assessment
761
Table 35.4 (continued) Pesticide
Endpoint
Chile
Pesticides
Micronuclei
Negative
Venegas et aI., 1998
Italy
Pesticides
Micronuclei
Positive
Falck et aI., 1999
Pesticide industry workers
India
Pesticides
SCE
Positive
Padrnavathi et aI., 2000
Farm workers
Canada
Pesticides
Micronuclei
Equivocal
Davies et aI., 1998
Greenhouse workers
Spain
Pesticides
Micronuclei
Negative
Lucero et aI., 2000
Factory workers
China
Pesticides
Aneuploid spenn
Positive
Padungtod et aI., 1999
Study group
Location
Pesticide sprayers Greenhouse workers
Result
Reference
Cs aberrations refer to both chromosome and/or chromatid aberrations. Micronuclei in bone marrow cells. C Negative at low and moderate exposures but positive at high exposures. d Micronuclei in oropharyngeal cells. e Breaks; unstable and stable chromosomal aberrations. f Numerical chromosomal aberrations. g Structural chromosomal aberrations, but exchanges showed a statistically significant increase in exposed over controls.
a
b
Micronuclei in exfoliated urothelial cells. Affecting the lcen-Iql2 region. j Unstable chromosomal aberrations during prespraying period. All studies were perfonned on peripheral blood Iymphocytes unless otherwise noted.
h i
induce chromosome alterations in the sperm of the exposed workers and may contribute to decreased reproductive performance of the workers.
35.6 GENOTOXICITY AND RISK ASSESSMENT As described previously, short-term tests for genotoxicity are required by regulatory agencies for pesticide registration and play an important role in the safety evaluation and risk assessment process. For the few agents that have been evaluated for heritable risks, genotoxicity assays, particularly those assessing heritable effects in germ cells, have played a critical role. Historically, in cancer risk assessment, the short-term test results and human biomonitoring studies have been used to alert agencies and the public to pesticides with potential cancercausing properties as well as to provide valuable supplemental information for the positive or negative results seen in animal bioassays. In recently implemented or proposed regulatory strategies, genotoxicity information plays an increasingly important role in the risk assessment process. DNA reactivity and mechanisms of genotoxicity are being used to provide in sights into an agent's mode of action and, as a result, may play a pivotal role in determining whether linear or nonlinear (apparent threshold) models will be used for extrapolation from high animal doses to lower exposure levels. In the EPA approach, genotoxic effects may also be modeled as precursor events to provide the basis for the selection of a certain extrapolation procedure (Wiltse and Dellarco, 1996). The use of mechanistic or mode-of-action information plays an important role in the cancer risk assessment guidelines proposed by the EPA (Wiltse and Dellarco, 1996) as well as in those implemented by other national and international regulatory groups such as the IARC
(IARC, 199ge) and the German Commission for the Investigation of Health Hazards of Chemical Compounds in the Work Area of the Deutsche Forschungsgemeinschaft (Neumann et aI., 1998). The evaluation of ethylene oxide provides a recent example of the contribution of genotoxicity data to the cancer risk assessment process. Upon reviewing the literature on the carcinogenicity of ethylene oxide in humans and animals, the IARC Working Group concluded that there was limited evidence for the carcinogenicity of ethylene oxide in humans but sufficient evidence in animals (IARC, 1994). However, in its overall evaluation, the IARC Working Group concluded that ethylene oxide is carcinogenic to humans. In making this conclusion, the IARC took into consideration evidence that "ethylene oxide is a directly acting alkylating agent that: (i) induces a sensitive, persistent dose-related increase in the frequency of chromosomal aberrations and sister chromatid exchanges in the peripheral lymphocytes and micronuclei in the bone-marrow cells of exposed workers; (ii) has been associated with malignancies of the lymphatic and haematopoietic system in both humans and experimental animals; (iii) induces a dose-related increase in the frequency of haemoglobin adducts in exposed humans and dose-related increases in the numbers of adducts in both DNA and haemoglobin in exposed rodents; (iv) induces gene mutations and heritable translocations in germ cells of exposed rodents; and, (v) is a powerful mutagen and clastogen at all phylogenetic levels." In a similar fashion with differing conclusions, the IARC recently evaluated data on the relevance of rodent tumors of the urinary bladder, renal cortex, mammary gland, and thyroid gland induced by agents such as atrazine, chlorothalonil, OPP, p-DCB, and saccharin and their relevance to carcinogenic risk in humans (Rice et aI., 1999a). In a number of cases, the lack of genotoxicity exhibited by these agents or their metabolites
762
CHAPTER 35
Genetic Toxicity of Pesticides
played an important role in its conclusions that the mechanisms by which agents such as atrazine and saccharin induced cancer in rodents were not relevant to humans (Rice et aI., 1999a). Other governmental groups have reached similar conclusions (NTP, 2000; U.S. EPA, 1991, 1998a). It should be emphasized that, in all cases, critical evaluation should be used in the interpretation and application of short-term test results in the risk assessment process. Given the large number of tests that can be performed in different cells or strains and at multiple dose levels, positive results should be expected in some tests by random chance alone. As a result, reproducibility and consistency become particularly important in evaluating genotoxicity test results. Short-term tests can also be performed in vitro or in vivo under conditions that will produce positive test results but that are unlikely to pose significant genotoxic risks to humans. For example, there is increasing recognition that positive responses in the in vitro chromosome aberration assay can be caused by mechanisms such as endonuclease activation that are not likely to occur at lower doses (Galloway, 2000; Scott et aI., 1991). These tend to occur more frequently at high test concentrations under conditions in which high osmolality, extremes of pH, or excessive cytotoxicity are seen. Similarly, genotoxic effects may occur at concentrations in vitro that most likely would not occur in vivo as other types of toxic effects such as neurotoxicity would be dose limiting. A comparison of in vitro concentrations or in vivo animal plasma concentrations with expected plasma levels in humans under conditions of normal (and above normal) usage can assist in the interpretation of the test data. Conversely, negative results in short-term genotoxicity tests should not be given undue weight as they do not exclude the possibility that an effect occurred in tissues that were not examined, that inadequate bioactivation was used, that the test was improperly conducted, or that the agent induces another type of genetic damage (IARC, 199ge; Proctor et aI., 1986). Additionally, negative results in these assays cannot be considered to rule out the carcinogenicity of agents that act through other mechanisms (e.g., receptor-mediated effects, cellular toxicity with regenerative proliferation, or peroxisome proliferation) (IARC, 199ge). By using a weight-of-evidence approach to evaluate the data, the likelihood of error (both false positives and false negatives) can be minimized. In a similar fashion, to confidently use human biomonitoring studies to evaluate risk, one should ensure that the biomarker of interest was sufficiently sensitive to detect changes at the exposure levels of interest, that the number of exposed and control individuals in the study was adequate, that an acceptable number of measurements was collected, and that major confounding variables were controlled. In addition, the identification of the specific pesticide and information on the exposure levels, although frequently difficult to obtain, can add significantly to the evaluation. Although it is uncommon for all of the preceding conditions to be fulfilled, results of human biomonitoring studies, when the specific agent is known, can play a valuable role in the risk assessment process (see the previous example for ethylene oxide).
In conclusion, a significant number of pesticides have exhibited genotoxic effects in short-term genotoxicity assays and may pose significant risks to humans. Consistent with this, chromosomal alterations have been seen in many studies monitoring genotoxic effects in pesticide-exposed workers. However, these studies often involve exposures to multiple pesticides and potential confounding factors and at levels much higher than those experienced by the general pUblic. The ongoing challenge for researchers, regulators, and those interested in environmental health is to effectively use genotoxicity data to distinguish noncarcinogenic and nonmutagenic pesticides from those capable of inducing cancer and heritable mutations in humans, to determine which of the latter pose significant risks at human exposure levels, and to identify safe methods and levels for the use of these agents or to eliminate their usage altogether.
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Genetic Toxicity of Pesticides
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CHAPTER
36 Immunotoxicity of Pesticides Kathleen E. Rodgers University of Southern California
36.1 INTRODUCTION
36.1.2 PESTICIDES
36.1.1 IMMUNE SYSTEM The immune system, the system by which foreign invaders are controlled or eliminated, contains two components: innate and acquired immunity. Innate immunity provides protection in an nonspecific manner and does not have a lag time before protection is conferred. Innate immunity consists of barriers, such as skin and mucous membranes, neutrophils, cells of the macrophage lineage, and natural killer cells. The last three elements are involved in the release of inflammatory mediators, such as enzymes and bioactive lipids, cytokines, which can stimulate cellular function or act to eliminate the invader directly, and reactive oxygen and nitrogen intermediate, which act together to inhibit cellular proliferation or are cytotoxic. The specific immune response requires a lag time for the generation of the response and has memory that allows a more intense, rapid, and specific response upon reexposure. There are two arms of the specific immune response, cellular immunity and humoral immunity. The effectors of cellular immunity are cytotoxic T cells that kill virally infected cells or tumor cells by direct contact. The effectors of humoral immunity are antibodies, which are generated by B cells and eliminate antigens by the formation of immune complexes, by complement fixation, or by enhancement of phagocytosis (opsonization) or antibody-directed cellular cytotoxity. The generation of cellular and humoral immune responses is regulated by suppressor T cells and helper T cells. Subsets of helper T cells have been defined which elaborate different cytokines, express different receptors, and are instrumental in the support of either cellular or humoral immunity depending upon the cell type. The optimal functioning of the immune system is under exquisite regulation that allows elimination of the foreign invader without a great deal of detriment to the host. Immune disorders arise when this balance is disrupted. Handbook of Pesticide Toxicology Volume 1. Principles
Pesticides are xenobiotics which are by definition biocidal and are designed to be selective to the species to be killed through metabolism or targeting of a site that is specific to the target organism. However, this is not accomplished for most pesticides and the safety of the pesticide is based on the quantity of chemical used and the method of application. Therefore, there are reasons for unique concern regarding the toxicity of pesticides and consideration as to the dose of a pesticide to which a test animal is exposed. For most pesticides, a marker for poisoning has been established based on the most sensitive parameter of toxicity in laboratory animals that is readily measured in humans. Although exposure to a dose of compound sufficient to cause acute poisoning may have an effect on the immune system, many studies have examined the effects of pesticides on the immune system, with particular emphasis on the administration of a nontoxic dose of compound. By administration of a nontoxic dose of pesticide, one can avoid the complications of the effects of stress and toxicities to other physiologic systems on the function of the immune system. The influence of pesticides on the immune response in humans has largely been ignored. The immune-mediated complications of pesticide exposure to humans that are most often noted are allergic reactions, especially contact dermatitis. This is probably due to the fact that an allergic reaction is readily observed and can be attributed to the causative agent through measurement of immune components specific for this agent. Although a great deal of data are available from animal studies which show that pesticides are immunosuppressive, there is no good evidence at this time for immune suppression in the general popUlation as a result of environmental exposure. This may be due to the lack of well-designed longitudinal studies to examine this potential effect of pesticides. To fully delineate the effects of pesticides on the generation of immune responses, sensitive tests (i.e., those that would measure an alteration in a functional immune response and not simply basal immune function) with low interassay, day-to-day, and person-to person variability should be conducted on persons occupationally ex-
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Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
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Immunotoxicity of Pesticides
posed to well-known levels and compositions of pesticides or those exposed to high levels after accidental exposures. These studies should be longitudinal to allow the determination of the kinetics of potential recovery and to eliminate the possibility that initial observations of variations from the norm were simply due to intrinsic variabilities within the measurements. The roadblocks to conducting such experiments include (1) lack of appropriate control groups, (2) inherent day-to-day and personto-person variability in commonly used assays of immune responsiveness, and (3) lack of knowledge of the physiologic significance of alterations in these parameters.
36.2 ORGANOCHLORINE PESTICIDES Organochlorine insecticides include many classes of compounds, such as chlorinated ethane derivatives (DDT), cyclodienes (chlordane, aldrin, heptachlor), and hexachlorocyclohexanes (lindane). These insecticides were widely used from the mid-1940s to the mid-1960s in the control of insects in agriculture, soil, and structures. This class of compound is less acutely toxic than anticholinesterase pesticides, but has a greater potential for chronic toxicity. 36.2.1 DDT The effects of dichlorodiphenyltrichloroethane (DDT) on the immune system were studied in several species using a variety of functional parameters. In most of these studies described here, DDT was given repeatedly to mimic the persistence of the chemical in the environment. Administration of 100 ppm orally of DDT to chickens for 40 days led to a decrease in the weight of lymphoid organs, including the spleen, thymus, and bursa of Fabricus (Subba Rao and Glick, 1977). In this same study, the humoral immune response to sheep red blood cells (SRBCs), a T-cell-dependent antigen, was also studied, but no effect on this parameter was noted. In another study, the humoral immune response to bovine serum albumin (BSA) after oral exposure to 100-400 ppm DDT orally for 5 weeks was shown to be decreased (Glick, 1974). One study examined the effects of DDT on the immune system of the guinea pig following a single intraperitoneal administration of 15 mg/kg DDT. Following this treatment, the antidiphtheria hemagglutinin titer was unchanged, but anaphylactic shock was decreased (Gablicks et aI., 1973). Studies also showed that the level of protein in the diet affected the ability of exposure to DDT to result in immunosuppression with greater than 3% protein being protective (Banerjee et aI., 1995). Further, combined exposure to DDT and stress will enhance the observed immunosuppression (Banerjee et aI., 1997). Two studies examined the effects of DDT on the cellular and humoral immune response in rats. Administration of 200 ppm DDT orally for 35 days decreased the response to ovalbumin (Wassermann et aI., 1969). In contrast, administration of 40 mg/kg/day DDT orally for 60 days increased both the humoral immune response and the delayed-type hypersensitivity
(DTH) response to BSA (Luki et aI., 1973). Administration of 200 ppm DDT or DDT metabolites (DDE and DDA) in their diet for 5 weeks suppressed the generation of both humoral immune and cell-mediated responses to ovalbumin (Banerjee et aI., 1996). DDT (0.25 mg/kg/day orally for 1 month) had no effect on the phagocytosis of bacteria by polymorphonuclear neutrophils (PMN) (Crocker et aI., 1969). Administration of up to 150 ppm DDT orally to rabbits for 4 weeks had no effect on the humoral immune response to SRBC. Only the high dose of 150 ppm affected the DTH response to tuberculin in this study (Andre et al., 1983). Administration of 20 or 200 pm DDT orally for 1 month decreased anaphylactic shock symptoms (Gablicks et aI., 1975). Only a few studies have examined the effects of DDT on the human immune response. In vitro exposure of human peripheral blood mononuclear cells (PBMC) or PMN had no effect on the mitogenic response to phytohemagglutinin, but slightly decreased the chemotactic response (Lee et aI., 1979). On the other hand, occupational exposure to DDT is thought to depress the PMN function as measured by chemotaxis, nitro blue tetrazolium (NBT) reduction test, and phagocytosis (Hermanowicz et aI., 1982). In addition, there was an increase in the incidence of infections in the occupationally exposed group. On the other hand, no correlation was found between DDT blood levels and the ability to respond to diphtheria immunization in children (Costa and Schvartsman, 1977). Others have suggested that DDT may induce allergic contact dermatitis in humans (Vanat and Vanat, 1971). In summary, DDT has been shown to increase, have no effect on, or decrease the immune system depending upon the dose, route, timing of administration, species, and antigen. 36.2.2 CHLORDANE Most studies of the effects of chlordane on the immune system were performed in mice. Adult mice are relatively resistant to the immunotoxic effects of chlordane. The generation of a humoral immune response was unaffected by administration of 0.1-8 mg/kg/day for 14 days. In this same study, the generation of some cell-mediated immune responses studied was elevated by this treatment regime (John son et aI., 1986). On the other hand, in vitro exposure of murine splenocytes from adult mice to chlordane resulted in a dose-dependent suppression of immune response at doses that reduced cell viability in culture (Johnson et aI., 1987). In contrast, the murine immune system was found to be most sensitive to the immunotoxic effects of chlordane when exposure occurred in utero. BALB/C mice were exposed in utero to 0.16 or 8 mg/kg/day chlordane given to the dam throughout gestation. These treatments did not affect the generation of a humoral immune response to SRBC, but the high dose did inhibit the DTH response to oxazolone (Menna et aI., 1985; Spyker-Crammer et aI., 1982). In studies using a similar treatment regime, alterations were found in the following immune parameters at one or more of the doses administered: natu-
36.2 Organochlorine Pesticides
ral killer (NK) activity, bone marrow hematopoiesis [as measured by colony-forming units-granulocyte macrophage (CFUGM) and colony-forming units-stem cell (CFU-S)], fetal liver hematopoiesis (as measured by CFU-GM and CFU-S), and the DTH response to influenza (Barnett et al., 1985a, b, 1990a, b; Blaylock et al., 1990; Chuang et al., 1992). Some of these alterations occurred up to 200 days after birth. No effects on the generation of a cytotoxic T-lymphocyte (CTL) response or the proliferative responses to T- and B-cell mitogens were observed (Barnett et al., 1985a, b). Further studies have shown that in utero exposure to chlordane led to enhanced peritoneal macrophage function similar to that observed with inflammatory stimuli (i.e., decreased 5' nucleotidase and transferrin receptor expression and protein synthetic pattern similar to inflammatory macrophages) (Theus et al., 1992). In this study, no additional effects on macrophage function were observed in chlordane-treated mice upon stimulation with thioglycollate. In vitro exposure of guinea pig PMN to chlordane resulted in the stimulation of many membrane-related events, including respiratory burst, membrane potential, calcium mobilization, and the release of membrane-bound calcium (Suzald et al., 1988). In humans, there has been a tentative association made between exposure to chlordane and blood dyscrasias (AMA Council on Drugs, 1962; Furie and Trubowitz, 1976; Infante etal., 1976;Stieglitzetal., 1967). In summary, the immune system was most sensitive to chlordane-induced immunotoxicity when exposure occurred in utero. Alterations also occurred following in vitro exposure of white blood cells to chlordane.
36.2.3 DIELDRIN The humoral immune response to bacterial antigen was decreased by oral administration of 50 ppm dieldrin to rabbits (Wassermann et al., 1969). In addition, administration of 2040 ppm dieldrin to ducks decreased the resistance to hepatitis virus (Friend and Trainer, 1974). Others have studied the effects of dieldrin on the murine immune system. In this model, dieldrin was administered in corn oil by intraperitoneal injection for up to 16 days. Antiviral resistance to the mouse hepatitis virus 3 (MHV3) was decreased after exposure to dieldrin (Krzystyniak et al., 1985, 1986). Dieldrin acted to increase host susceptibility to virus through suppression of the humoral immune responses and alterations in cell-mediated immunity (Bernier et aI., 1987; Hugo et aI., 1988a, b; Krzystyniak et aI., 1985). However, studies showed that the use of a pathogenic antigen, which could itself suppress the immune response, in the process of examining the immunotoxic effects of dieldrin did not contribute to the observed immune suppression (Fournier et aI., 1988). The authors suggest that the alterations in cell-mediated immunity may occur at the level of antigen recognition rather than proliferation of the lymphocytes in response to stimulation (Hugo et aI., 1988a, b).
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Alterations in macrophage function that may contribute to the observed reduction in host resistance to MHV3 were observed (Krzystyniak et aI., 1986, 1987). Specifically, antigen presentation and phagocytosis were altered by exposure to dieldrin (Bernier et aI., 1988; Kaminski et aI., 1982; Krzystyniak et aI., 1985, 1989). In these studies, it was shown that, after exposure to dieldrin, the quantity of antigen taken up, cell associated, and released by macrophages was suppressed. These alterations in macrophage function and the decrease in the ability to generate a humoral immune response may occur through alteration in membrane integrity and fluidity (Alberts et aI., 1965; Antunes-Madeira and Madiera, 1979).
36.2.4 HEPTACHLOR Administration of 1 ppm of heptachlor to chickens for 3-8 weeks decreased the weight of the bursa of Fabricus (lymphoid organ in the chicken) (Rodica and Stefania, 1973). In addition, administration of heptachlor to rats decreased the levels of serum gamma globulins (Klimova, 1970). In vitro exposure of guinea pig PMNs to heptachlor and heptachlor epoxide induced the generation of superoxide anion, altered the membrane potential, induced calcium mobilization, and induced the release of membrane-bound calcium (Suzald et aI., 1988).
36.2.5 LINDANE Very few studies of the effect of lindane on the immune response have been performed. Administration of 150 ppm lindane per day orally for 1 month to mice did not alter the levels of IgA, IgG 1, IgG2a, and IgM in the serum or the ability to generate a humoral immune response to SRBC (Andre et aI., 1983). However, the serum levels of IgG2b were elevated by this exposure. In the rabbit, administration of 3 mg/kg/day or more lindane for 5 weeks decreased the humoral immune response to bacterial antigen (Desi et aI., 1978; Kaliser, 1968). In this study, the no-observable-adverse-effect level (NOAEL) for the immunotoxic effects of lindane was shown to be 1.5 mg/kg/day. In vitro exposure of human PBMC to 0.1-0.3 mM lindane inhibited the mitogenic response to a T-cell mitogen (Roux et aI., 1979). Further studies were conducted on the immunotoxicity of ,B-hexachlorocyclohexane (HCH), an isomeric contaminant of lindane, in mice. Mice were given up to 300 ppm HCH in their diet for 30 days. At 300 ppm, but not 100 ppm, suppression of the ability of splenocytes to proliferate in response to mitogen to generate a CTL response and NK activity was suppressed (Cornacoff et aI., 1988). As with chlordane, there is circumstantial evidence for the association of lindane with aplastic anemia and agranulocytosis, as well as other blood dyscrasias, in humans. However, the incidence of this type of toxicity associated with lindane exposure is very low (Gewin, 1959; Jedlicka et al., 1958; Loge, 1965; Mastromattco, 1964; Samuels and Milby, 1971; Sianchez-Madel et aI., 1963; West, 1967).
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36.2.6 MIREX Administration of 100 ppm mirex orally to chickens for 40 days decreased the weight of the thymus and the spleen, but increased the weight of the bursa of Fabricus (Subba Rao and Glick, 1977). However, administration of at least 500 ppm mirex orally for 5 weeks was required to inhibit the generation of a humoral immune response to BSA (Roux et al., 1979). 36.2.7 TOXAPHENE One study was conducted to examine the effects of toxaphene on the immune system. In this study, administration of 100200 ppm either to adult animals or in utero decreased the humoral immune response to BSA and phagocytosis by peritoneal macrophages, but had no effect on the DTH response to tuberculin antigen (AlIen et aI., 1983). 36.2.8 ENVIRONMENTAL EXPOSURE Recent studies have evaluated the effect of living in a contaminated environment on the immune function of marine mammals. Recent mass mortality among several populations has lead to the hypothesis of increased susceptibility to viral and opportunistic infections. In these studies, harbor seals were fed fish from either contaminated or noncontaminated sources over 2.5 years. Seals fed contaminated fish were suppressed in both humoral and cell-mediated immune responses (De Swart et aI., 1996; Ross et aI., 1995. Further studies in free-ranging dolphins showed that an inhibition of proliferative responses was correlated with concentrations of organochlorines in the blood (Lahvis et aI., 1995). 36.2.9 SUMMARY In summary, the effects of a number of organochlorine pesticides on the immune system have been examined. For the most part, these compounds suppressed the generation of immune responses and reduced the resistance of the test animal to infections. Several laboratories have shown that these compounds alter membrane fluidity and modulate membranemediated events. In fact, those who suggest a mechanism of action for the immunotoxic effects of organochlorine pesticides refer to this as a possibility.
36.3 ANTICHOLINESTERASES 36.3.1 ORGANOPHOSPHATES ESTERS Organophosphate pesticides are widely used compounds in agriculture, by consumers, and in public health situations due to their relatively low toxicity, rapid removal from the environment, and lack of bioaccumulation. However, organophosphates have the dubious distinction of being the pesticides most
often implicated in poisonings due to pesticides (parathion contributing in large part to this). Organophosphates of low mammalian toxicity, such as malathion, are used in situations such as structural treatment for mosquito eradication, spraying of tobacco plants, and aerial spraying of urban populations for eradication of fruit flies. Therefore, there is potential for both high exposure to organophosphate pesticides during occupational exposure and low-level exposure of large segments of the general population. A great deal of information is available with regards to the effects of organophosphates on the immune system. 36.3.1.1 Parathion The effects of parathion on the immune system have been extensively studied. Wiltrout et al. (1978) showed that subacute administration of parathion (2.2-22.3 mg/kg/day) to mice blocked the generation of a humoral immune response (Wiltrout et aI, 1978). In addition, Dandliker et al. (1985) demonstrated that parathion was able to suppress both cellular and humoral immunity (Dandliker et aI., 1985). Alternatively, one study showed a decrease in the lymphoid organ weight with no change in the humoral immune response following repeated exposure to parathion. Others showed that parathion suppressed the humoral immune response following acute, subacute, and in vitro exposure (Bartholomew et aI., 1984; Casale et aI., 1983; Duggan et al., 1984). Peroral dosing of parathion to mice with a cytomegalovirus infection elevated mortality (Raise, 1983). The proliferative response of human lymphocytes to mitogens was suppressed following in vitro exposure to paraoxon (Waterhouse and Tourney, 1984). Paraoxon and two structurally related compounds inhibited the production of interleukin 2 by rat splenocytes (Pruett and Chambers, 1988). The in vitro exposure (1-125 J.-Lg/ml) of splenocytes to parathion and methyl parathion blocked the generation of a cell-mediated immune response (Rodgers et aI., 1986b). Methyl parathion (up to 3 mg/kg/day) increased the virulence of Salmonella typhimurium infection in rabbits (Fan, 1981; Fan et aI., 1984). In another study, methyl parathion administered over 4 weeks (up to 1/10 LDso) did not affect the generation of humoral or cellular immune responses in rabbits (De si et aI., 1978). Administration of a single high dose or repeated lower doses of methyl parathion to mice elevated the humoral immune response with no effect on DTH reaction (Institoris et al., 1992). Administration of 0.22-0.44 mglkg methyl parathion over three generations was studied. Alterations in immune function were detected at a dose of 0.29 mg/kg methyl parathion, but the parameters altered varied between generations (Institoris et aI., 1995). In vitro exposure of human PBMC to methyl parathion did not affect the proliferative response to mitogen, but decreased the chemotactic response (Lee et aI., 1979). Studies were also conducted as to the effects of parathion on hematopoiesis. Gallichio et al. Gallichio et al. showed that oral administration of parathion for 14 days, at a dose that did not affect the body weight or generate cholinergic symptoms, altered
36.3 Anticholinesterases the bone-marrow-derived stem cell colonies for up to 2 weeks after the last dose of parathion (Gallichio et aI., 1987a). In vitro exposure of human bone marrow cells to paraoxon or malaoxon significantly depressed the in vitro generation of colonies of burst-forming units-erythroid (BFU-E), colony-forming uniterythroid (CFU-E), and CFU-GM (Gallichio et aI., 1987b). These studies are difficult to correlate due to the differences in exposure route, immune parameters studied, and the species studied. In general, however, parathion was shown to be immunosuppressive, but the mechanism of these effects is unknown. 36.3.1.2 Malathion There have been several studies on the effects of malathion on the immune response. Repeated exposure to malathion results in allergic responses in man, guinea pigs, rabbits, rats, and mice (Centeno et aI., 1970; Cushman and Street, 1983; Hazelton, 1992; Magnusson and Kligman, 1987; Milby and Epstein, 1964; Vijay et aI., 1978). In contrast, a DTH response to malathion (up to 100 Il-g) was not observed in mice and guinea pigs (Cushman and Street, 1983; Kynoch and Smith, 1992). Administration of low doses of malathion for prolonged periods results in a decrease in the humoral immune response. For example, low doses of malathion (up to 50 mg/kg) given for 5-6 weeks to rabbits significantly lowered the humoral immune response to bacterial antigen (Desi et aI., 1986). In addition, a cholinergic dose of malathion suppressed the generation of a humoral immune response, whereas mUltiple low doses did not (Casale et aI., 1983). In contrast, exposure of mice to malathion dip (2 or 8%) or chickens to 400-1600 ppm malathion in their diet did not alter the generation of humoral immunity (Relford et aI., 1989; Varshneya et aI., 1988). Administration of high, noncholinergic doses of malathion (up to 715 mg/kg) to mice elevated the generation of a humoral immune response and proliferative responses to mitogen. Acute or subacute administration of malathion did not affect the generation of a CTL response to allogeneic tumor (Rodgers et al., 1986c). In vitro exposure of human PBMC or mouse splenocytes to malathion suppressed the proliferative response to mitogens and the generation of hydrogen peroxide (Rodgers et aI., 1986a; Rodgers and Ellefson, 1990a). In mouse splenocytes, the generation of CTL responses was also blocked by in vitro exposure to malathion (up to l25Il-g/kg). Further studies indicated that, when malathion was coincubated with liver enzymes to allow metabolism, the metabolites of malathion were no longer able to block the generation of a CTL response or the proliferative response to mitogens (Rodgers et aI., 1985a, 1986b; Rodgers and Ellefson, 1990a). In contrast, in vitro exposure of murine peritoneal cells or human PBMC to metabolized malathion elevated the respiratory burst activity of these cells (Rodgers and Ellefson, 1990a). Further studies were conducted to determine the mechanism of action of malathion on the immune system. Cell separation and reconstitution experiments after acute administration of high, noncholinergic doses of malathion (up to
773
715 mg/kg) showed that the macrophages were the cell type affected by malathion (Rodgers and Ellefson, 1990a). Further studies showed that acute administration of malathion elevated the respiratory burst of peritoneal leukocytes. The lowest observable adverse effect level (LOAEL) and NOAEL for the effect of acute administration of malathion on the respiratory burst of peritoneal leukocytes were shown to be 0.25 and 0.1 mg/kg malathion, respectively (Rodgers and Ellefson, 1992). Further studies in animals administered malathion for 14 or 90 days showed systemic degranulation of basophilic cells and macrophage activation (Rodgers and Xiong, 1997a, b). Microscopic examination of the peritoneal cells showed that peritoneal mast cells were degranulated within 4 hr after malathion administration. In addition, the percentage of peritoneal phagocytes ingesting mast cell granules and the number of granules ingested per cell were elevated. Further, the systemic release of mast cell mediators, ,B-hexosaminidase and histamine, was observed after oral administration of malathion (Rodgers and Ellefson, 1992; Rodgers and Xiong, 1998). This exposure to mast cell products may elevate macrophage function. Studies involving mast cell-deficient mice showed that the presence of mast cells was necessary for an elevation in macrophage and immune function after malathion was administered (Rodgers, 1997). More recently, it was shown that the release of mast cell mediators, both inflammatory mediators and histamines, contributes to alterations in macrophage function that occur after oral administration of malathion (Rodgers and Xiong, 1996, 1997c). In vitro exposure of rat basophilic leukemia (RBL-l) cells to paraoxon, but not parathion, and malathion caused degranulation of these cells (Rodgers and Ellefson, 1992). Further studies on purified metabolites of malathion showed that degranulation of both normal basophilic cells (rat and human) and the RBL-l tumor cell line occurred after in vitro exposure to not only the potent inhibitors of anticholinesterase, malaoxon and isomalathion, but also the nonneurotoxic metabolite, dicarboxylic acid, of malathion (Xiong and Rodgers, 1997). This may explain why immunologic effects were observed at doses two to three orders of magnitude lower than the noncholinergic dose of malathion. These data, together with the report in the literature that diisopropyl fluorophosphate (DFP) and soman cause mast cell degranulation in an IgE-independent manner, suggest that malathion or a metabolite of malathion may act to elevate the immune response through inhibition of a cell surface-serine esterase on mast cells, subsequent degranulation of mast cells, and exposure of macrophages to mast cell products (Kazimierczak et aI., 1984). Recent studies have shown that weekly administration of relatively high doses of malathion (33-100 mg/kg) will accelerate the onset of autoimmune disease in mice predisposed to SLE, but not in their littermates which do not have a gene for accelerated onset of autoimmune disease (Rodgers, 1997). 36.3.1.3 Effects of Impurities in Organophosphate Pesticides The effects of impurities in organophosphate pesticides (malathion, acephate, and fenitrothion) on the immune system have
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been studied. The maJonty of the studies were conducted with O,O,S-trimethyl phosphorothioate (OOS-TMP). Following acute, nontoxic doses of OOS-TMP (up to 10 mg/kg), the generation of both cell-mediated and humoral immunity was blocked following in vivo and in vitro exposure to antigen (Devens et aI., 1985; Rodgers et aI., 1985a, 1986b). However, OOS-TMP did not affect the proliferative response to mitogens, but did elevate the production of interleukin 2 (IL-2) (Rodgers et aI., 1986a). More recent studies have shown that 0,0,0trimethyl phosphorothioate (OOO-TMP) (up to 40 mg/kg) elevated humoral and cell-mediated immune responses and protected against OOS-TMP-induced immune suppression when coadministered with OOS-TMP (Rodgers et aI., 1989a). In addition, exposure to low levels of OOS-TMP protected against an immunosuppressive dose of OOS-TMP. Subchronic (14-day) exposure to OOS-TMP increased the humoral and cell-mediated immune responses, mitogenic responses, and IL-1 production (Rodgers et aI., 1985b). The suppression of the immune response following acute administration of OOS-TMP (up to 10 mg/kg) was dose and time dependent, and macrophages were shown to be the cell type most affected by OOS-TMP (Devens et aI., 1985; Rodgers et aI., 1985c). Macrophages from OOS-TMP-treated mice were shown to (1) be larger in size, (2) have increased nonspecific esterase activity, (3) be less effective at antigen presentation, (4) have increased phagocytic activity, (5) secrete increased levels oflL-l, (6) have decreased la and F4/80 expression, (7) release suppressive factors, (8) have increased respiratory burst activity, and (9) secrete increased levels of neutral proteases, plasminogen activator, elastase, and collagenase (Rodgers and Ellefson, 1988a, b, 1990b; Rodgers et aI., 1985c, d, 1987b). These effects were transient and macrophage function was comparable to controls within 7 days (at which time immune function was similar to controls) (Rodgers and Ellefson, 1988b, 1990b). OOS-TMP also caused thymic atrophy (Devens et al., 1985). More recent investigations have shown a reduction in the number of cells expressing T-cell markers in the thymus (Rodgers et aI., 1987a-c). Studies are ongoing to determine the identity of mouse thymus cells which are targeted following acute administration of OOS-TMP. In summary, acute in vivo administration of OOS-TMP was immunosuppressive while stimulating macrophage function; repeated exposures stimulated cellular and humoral immune responses. In vivo exposure to 0 ,S,S,-trimethyl phosphorodithioate enhanced or suppressed the generation of cell-mediated or humoral immune response at nontoxic or toxic (assessed by suppression of plasma cholinesterase) doses, respectively, following in vivo or in vitro stimulation with antigen (Rodgers et aI., 1987c, 1988b). In vivo exposure to OSS-TMP also elevated proliferative responses to mitogens, but suppressed IL-2 production (Rodgers et aI., 1988b). Fourteen-day exposure to OSS-TMP (20 or 40 mg/kg) elevated or suppressed (60 or 80 mg/kg), depending upon the dose, the generation of immune responses (Rodgers et aI., 1989b). Further studies showed that OSS-TMP altered both T- and B-lymphocyte function (Thomas
and Imamura, 1986). In vitro exposure to OSS-TMP enhanced or suppressed immune function, depending upon the OSS-TMP concentration and the in vitro metabolism system used (Rodgers et aI., 1988b; Thomas and Imamura, 1986). OSS-TMP also inhibited the cytolytic function of cloned murine and human CTL, but only if present during the time when the cell to be lysed was being recognized (as measured by conjugation) by the CTL (Rodgers et aI., 1988a). OSS-TMP was immunostimulatory at noncholinergic doses and immunosuppressive at cholinergic doses (similar to that described previously for malathion). In vitro exposure to OSS-TMP suppressed humoral and cellmediated immune function. Finally, one study showed that O,O-dimethyl, S-ethyl phosphorothioate, a synthetic analog of the impurities described previously, blocked cell-mediated and humoral immune response after in vitro exposure through impairment of lymphocyte function (Thomas et aI., 1986). These studies show that the impurities found in technical malathion can modulate immune function. The cell type affected, the duration of the effect, and the immune parameter modulated varied from compound to compound and were related to the duration of exposure. 36.3.1.4 Other Organophosphate Pesticides The effects of many organophosphates have been assessed on at least one facet of the immune system. Carbophenothion and crufomate suppressed the proliferation of human lymphocytes in response to mitogen (Park and Lee, 1978). Acute administration of dichlorvos slightly decreased splenic weight of mice, but did not affect the generation of a humoral immune response (Cas ale et aI., 1983). In addition, chronic, low-level exposure to dichlorvos (up to 1/10 LD50) suppressed the generation of serum antibodies following vaccination of rabbits with Salmonella typhimurium and the generation of cell-mediated immunity following a tuberculin vaccination (De si et aI., 1978, 1979). Acute administrations of cholinergic doses of dichlorvos (up to LD50) resulted in mobilization of bone marrow cells and suppression of cellular and humoral immune responses. The mediator (corticosterone or acetylcholine) associated with these alterations varied with the parameter measured (Zabrodski, 1993). In vitro treatment of rabbit PMN by diisopropylethyl phosphate or triisopropyl phosphate reduced locomotion of leukocytes (Woodin and Harris, 1973). Oral, acute administration of DFP to guinea pigs enhanced the serum complement and hemolysin activity and the generation of a humoral immune response, but suppressed lysozyme activity. Alternatively, repeated administration of DFP suppressed complement, hemolysin, and lysozyme activities, and the generation of a humoral immune response (Lis and Mierzejewski, 1980). Intraperitoneal injection of dimethoate reduced the thymic and splenic weight of treated mice and blocked the generation of a humoral immune response (Tiefenbach and Lange, 1980). Further administration of 7.04-14.1 mg/kg dimethoate over three generations to rats affected immune function. However, the parameter affected varied from generation to generation (lnstitoris
36.3 Anticholinesterases
et aI., 1995). Administration of fenchlorphos to chickens for 3-8 weeks increased the weight of the bursa of Fabricus (Rodica and Stefania, 1973). In vitro exposure of murine splenocytes to fenthion (up to 125 I-lg/ml) blocked their ability to generate a cell-mediated immune response (Rodgers et aI., 1986b). In contrast, topical administration of fenthion to newborn mice with encephalomyocarditis virus infection did not alter mortality (Crocker et aI., 1974). Leptophos (up to 500 pm) administered orally for 12 weeks did not affect the generation of the humoral immune response in the mouse (Koller et aI., 1976). Demeton-O-methyl decreased the generation of a humoral immune response in the rat when a single high dose was administered (Nikolayev et al., 1972). Monocrotophos (up to 8 mg/kg), given intraperitoneally one time per week for 6 weeks, modulated several hematological parameters, including an increase in clotting time, white blood cell count, splenic cellularity, and the percentage of large lymphocytes, neutrophils and basophils (Gupta et aI., 1982). In vitro exposure of human basophils to soman led to an IgEindependent release of histamine (Meier et aI., 1985). Esa et al. (1988) showed that in vitro exposure of human mononuclear cells to triphenyl phosphine oxide and tetra-o-cresyl piperazinyl diphosphoramidate caused suppression of antigen-specific proliferation. In addition, treatment of human monocytes with triphenyl phosphine oxide, tetra-o-cresyl piperazinyl diphosphoamidate, triphenyl phosphate, and triphenyl thiophosphate significantly inhibited their ability to present antigen to immune T cells. Exposure of mice to tris-(2,3-dichloropropyl) phosphate decreased the proliferative response of splenocytes to mitogens and increased the incidence of tumors after challenge, but did not alter splenic or thymic weight, hematological parameters, DTH, serum Ig levels, humoral immune responses to T-cell-dependent and -independent antigens, and mortality following Listeria monocytogenes infection (Luster et aI., 1981). Triphenyl phosphate caused an allergic reaction and suppressed the immune system by subchronic administration (Carlsen et aI., 1986; Hinton et aI., 1987).
36.3.1.5 Effects of Organophosphates Esters on Humans Based on Epidemiology Some epidemiology studies have indicated that organophosphates may have an effect on the human immune system. Exposure to organophosphates has been shown to cause allergic reactions (3- to 4-month exposure), a decrease in rosette-forming T cells and an increase in B cells, a decrease in leukocyte phagocytic activity, and an increased susceptibility to colds and subjective health complaints (Bellin and Chow, 1974; Hermanowicz and Kossman, 1984; Kanezaki et aI., 1973; Katsenovich et al., 1981; Malenkii, 1978; Sawinsky and Durst, 1973; Zaninovic, 1977). One study has shown a decrease in monocyte esterase activity in workers occupationally exposed to an organophosphate compound (Lee and Waters, 1977). Occupational exposure to organophosphate pesticide decreased PMN chemotaxis and adhesion, but increased NBT reduction. In these studies, there may be a suggestion of immune modula-
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tion, but the extent and mechanism of these effects are difficult to ascertain at this time. 36.3.1.6 Summary Most of these studies show that a variety of organophosphate pesticides reduce immune function in a variety of species. Although the site of action has not been identified at a molecular level, most investigators agree that organophosphate compounds probably act through inhibition of serine esterases. 36.3.2 CARBAMATES Carbamates, like organophosphate pesticides, act through inhibition of anticholinesterase and the symptoms of toxicity are similar to those observed with organophosphates. However, most of these compounds have low dermal toxicity, unlike most organophosphates. Carbamates are not broad spectrum pesticides, but, like organophosphates, they are relatively nonpersistent in the environment. Therefore, these compounds are widely used to eradicate the species for which they are indicated. 36.3.2.1 Carbaryl Carbaryl increased the serum level of IGGI and IgG2b without affecting the other Ig classes following oral exposure for 1 month (Andre et aI., 1983). In addition, administration of carbaryl (up to 10 I-lg/day) to quail for 5 days lowered their resistance to the protozoan parasite Histomonas meleagrides (Zeakes et aI., 1987). The humoral and cellular immune response of rabbits to antigen was unchanged following oral carbaryl (up to 150 ppm) for 4 weeks (Street and Sharma, 1975). Carbaryl suppressed a humoral immune response at very high doses (50% LDso) (Wiltrout et al., 1978). A study was conducted more recently in which the effect of carbaryl, administered by three different routes, was examined. The humoral immune response was suppressed after inhalation of carbaryl (up to 335 mg/kg), but not after oral or dermal exposure (Ladics et aI., 1994). In vitro exposure of splenocytes to carbaryl blocked their ability to generate a cellular immune response (Rodgers et aI., 1986b). Carbaryl suppressed the expression of complement activity in human serum when added to the assay (Casale et aI., 1989). Carbaryl and some of its metabolites inhibited the proliferation of interleukin-2-dependent T cells (Bavari et aI., 1989). In vitro exposure of large granular lymphocytes to carbaryl was also shown to inhibit the proliferation and the induction of NK activity in response to interleukin 2 (Casale et aI., 1990; Street and Sharma, 1975). These authors, as do others as discussed previously, suggest that carbaryl and other compounds that inhibit serine esterases act through this mechanism to modulate the generation and expression of immune responses. 36.3.2.2 Carbofuran Carbofuran did not affect the generation of a humoral immune response, but significantly suppressed cellular immunity when
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given over a 4-week period (Street and Sharma, 1975). Carbofuran (up to 0.6 mg/kg) also decreased the humoral immune response to neutral and pathogenic antigens and increased the cytolysis of macrophages by virus (Fournier et aI., 1988). However, carbofuran did not affect the generation of a DTH response in vivo or a cell-mediated immune response in vitro. On the other hand, carbofuran suppressed the expression of complement activity when added to human serum during the assay (Casale et aI., 1989). Carbofuran (up to 0.25 mg/kg), given intraperitoneally over 6 weeks, modulated several hematological parameters (Gupta et aI., 1982). 36.3.2.3 Aminocarb
Aminocarb decreased humoral immune response to neutral and pathogenic antigens and increased the cytolysis of macrophages by virus (Fournier et aI., 1988). Aminocarb did not decrease the resistance of mice to Salmonella typhimurium and MHV3 (Krzystyniak et aI., 1989). Aminocarb suppressed the generation of serum antibody titer to MHV3, but when given orally, enhanced the humoral immune response to a nonpathogenic antigen (Fournier et aI., 1986). Aminocarb did not affect the generation of a cell-mediated immune response or the ability of macrophages to process antigen (Fournier et aI., 1988). A more recent study compared the immunotoxic potential of aminocarb given by four different routes. Oral and dermal administration of relatively low doses of aminocarb elevated a humoral immune response. Intraperitoneal administration suppressed a humoral immune response whereas inhalation had no effect (Bernier et aI., 1995). 36.3.2.4 Aldicarb
Studies by Fiore et al. (1986) showed that women drinking well water containing detectable levels of aldicarb exhibited increased percentages and absolute numbers of CD8-positive peripheral blood lymphocytes. As a result of this increase in CD8+ cells, there was a decrease in the CD4/CD8 ratio in these women. In contrast, additional studies of exposed persons showed no effect on immune function or increase in clinical illness (Hong, 1991). Animal studies were conducted to further examine this possible alteration in immune function. Olson et al. (1987) and Shirazi et al. (1990) showed that low levels of aldicarb decrease the humoral immune response, but Thomas and co-workers (Thomas et aI., 1987; Thomas and Ratajczak, 1988) showed that low levels of aldicarb (up to 1000 ppb) did not affect the generation of cellular and humoral immune responses or the resistance of the host to infection or tumor challenge. Further studies were done on mice that received 1-100 ppb aldicarb in their drinking water for 34 days (Thomas et aI., 1990). In this study, no alterations were found in the percentage and absolute number of T cells, B cells, or T-cell subpopulations in the spleen (Thomas et aI., 1990). In addition, administration of aldicarb did not alter splenic NK activity or the ability of splenocytes to generate an allogeneic CTL response. A more recent study confirmed that chronic (90-day)
administration of 0.1-10 ppb aldicarb in drinking water did not affect any immune parameters measured (Hajoui et aI., 1992). Others have studied the effects of administration of aldicarb on macrophage and T-cell function. In these studies, aldicarb (up to 1000 ppb) in corn oil administered one time by intraperitoneal injection. Aldicarb administered intraperitoneally decreased the ability of macrophages to lyse tumors in an antibody-dependent cell-mediated cytotoxicity assay, but did not alter NK activity (Selvan et aI., 1989). In addition, aldicarb treatment suppressed the generation of a syngeneic mixed lymphocyte reaction by selectively decreasing the stimulatory activity of the macrophages without directly affecting autoreactive T cells (Dean et aI., 1990a, b). Further studies showed that the proliferation of splenocytes from treated mice to Con a and anti-CD3 antibodies was decreased after aldicarb exposure. Cell separation and reconstitution experiments showed that alteration in macrophage function, specifically decreases in IL-I production, may be responsible for this decrease in T-cell proliferation after stimulation (Dean et aI., 1990a, b). 36.3.2.5 Ethyl and Methyl Carbamate
Ethyl carbamate inhibited humoral immune response to T-celldependent and -independent antigens (Haran-Ghera and Peled, 1967; Luster et aI., 1982; Malmgren et aI., 1952; Parmiani, 1970; Parmiani et aI., 1969). However, ethyl carbamate did not affect or only slightly affected cell-mediated immunity (DiMarco et aI., 1972; Lappe and Steinmuller, 1970; Luster et aI., 1982; Parmiani, 1970). Ethyl carbamate did not affect the resistance of mice to encephalomyocarditis virus infection, but increased the incidence of induced leukemia (Chieco-Bianchi et aI., 1963). Administration of ethyl carbamate (200-400 mg/kg) for 14 days to mice reduced splenic and thymic weight and increased splenic myelopoiesis. Preinduction of P450 liver enzymes with phenobarbital resulted in an increase in the immunosuppression observed after administration of ethyl carbamate (Jeong et al., 1995). Macrophage phagocytic and bacteriocidal functions were unaffected, but the release of cytostatic factors from macrophages was elevated. Bone marrow myelopoietic function and splenic NK activity were suppressed after exposure to ethyl carbamate (up to 0.8 mg/kg) (Gupta et aI., 1982; Luster et aI., 1983). In contrast, 14-day exposure to methyl carbamate did not affect splenic or thymic weight, cell-mediated or humoral immunity, mitogenic responses, macrophage function, bone marrow function, or NK activity (Luster et aI., 1982). Perinatal exposure of mice to ethyl carbamate resulted in the induction of tumors in adults. A study of immune function of the pups after in utero exposure to ethyl carbamate (up to 1000 mg/kg) on days 7-16 of gestation or of neonates after exposure on postpartum days 5-14 (with a total of 1-2 mg/g ethyl carbamate given) was conducted. Postnatal exposure suppressed NK activity only. However, prenatal exposure increased leukocyte counts and suppressed the generation of a humoral immune response (Luebke et aI., 1986).
36.5 Herbicides 36.3.2.6 Summary The effect of carbamates on the immune system has been studied in a variety of systems. These compounds were shown either to not affect or to suppress the generation of immune responses depending upon the dose, route, and timing of exposure.
36.4 PYRETHROIDS Pyrethrum is a naturally occurring pesticide extracted from the chrysanthemum flower. Over the last several years, several pyrethroid pesticides, based upon the structure of pyrethrum, have been developed. This group of compounds is very active, has a high insect/mammal toxicity ratio, and does not persist in the environment. Therefore, these compounds are widely used. Very little information, however, is available regarding the effects of this class of compounds on the immune system. Initial studies showed that after acute and subchronic administration of cypermethrin (up to 1/2 LDso), a synthetic pyrethroid, there was an early dose-dependent suppression of the generation of a humoral immune response to Salmonella typhimurium in rabbits at doses that do not cause other toxicologic symptoms (De si et aI., 1986). In rats, the generation of a humoral immune response to SRBC and ovalbumin was suppressed by cypermethrin administration (up to 40 mg/kg) (Desi et aI., 1986). Further studies were conducted in mice and goats. Administration of cypermethrin (up to 50 mg/kg) intraperitoneally for 26 days (mice) or dermal exposure (up to 41.6 mg/kg) for 30 days (goats) resulted in a decrease in the DTH response to 2,4dinitrofluorobenzene. In addition, the generation of a humoral immune response was suppressed in this study (Tamang et aI., 1988). However, administration of up to 12 mg/kg cypermethrin for 28 days had no effect on immune function (Madsen et aI., 1996). Studies were conducted on the effects of exposure to deltamethrin for 10-30 days on the generation of immune responses in mice and rats (Kowalczyk-Bronisz et aI., 1990; Madsen et aI., 1996). In these studies, the authors observed little effect on the immune system. In contrast, oral administration of 6 mg/kg deltamethrin for 84 days or 15 mg/kg for 14 days resulted in suppression of humoral and cellular immune reponses (Lukowicz-Ratajczak and Krechniak, 1992). Exposure of mice to a newer pyrethroid pesticide, Supercypermethrin Forte, resulted in inhibition of a humoral immune response only after administrations of doses that resulted in mortality in some mice (Siroki et aI., 1994). Exposure of mouse splenocytes to allethrin, cypermethrin, fenpropathrin, and pennethrin in vitro decreased the proliferative response to mitogen (Stelzer and Gordon, 1984). These few studies suggest that cypermethrin may inhibit the immune system in a variety of species, but these effects are not observed for all compounds.
36.5 HERBICIDES Herbicides are chemicals used for the destruction of unwanted foliage. Other than the effects of dioxin, a contaminant in some
777
herbicides, on the immune system, very few studies have been done to examine the effects of herbicides on the immune system. Oral administration of 100 mg/kg/day atrazin, a triazine herbicide, for 3 days to rats decreased the number of white blood cells, but did not affect lymphoid organ weight or serum immunoglobulin levels. Oral administration of diuron, a substituted urea herbicide, for 3 weeks to rats increased the weight of lymphoid organs (Vos and Krajnc, 1983). One study conducted in rats showed that mecoprop, a phenoxy acid herbicide, altered the structure of the spleen and the thymus and altered the number of blood lymphocytes and granulocytes (Moeller and Solecki, 1989). The authors suggest that these alterations may be the result of chemically induced stress. Administration of 20-320 mg/kg/day Ordram for 12 days had no consistent effect on organ weights, natural killer activity, proliferative responses, DTH responses, and humoral immune response (Smialowicz et aI., 1985). Studies have also been conducted on the effects of propanil, a postemergence herbicide used in rice and wheat production. Propanil was given through intraperitoneal injection. Acute administration of this compound increased splenic weight and cellularity and suppressed the generation of humoral and cell-mediated immune responses at doses of 50 or 400 mg/kg, respectively (Barnett and Gandy, 1989). In addition, administration of 50-200 mg/kg propanil resulted in reduction in the number of myeloid and erythroid progenitor cells (Blyler et aI., 1994). Analysis of thymocyte subpopulations after propanil (100 mg/kg) administration showed a decrease in the number of single and double positive thymocytes (with no effect on splenic or lymph node populations) (Zhao et al., 1995). Further studies were conducted on the major metabolite of propanil, 3,4-dichloroaniline (DCA), on the immune system (Barnett et aI., 1992). Again, the compound was administered intraperitoneally. DCA, like propanil, increased the splenic weight and cellularity. In addition, DCA suppressed the generation of a humoral immune response to both T-celldependent and -independent antigens. Both propanil and DCA inhibited NK activity, but neither chemical affected the generation of a CTL response. These studies suggest that this herbicide suppressed selected immune responses, but not all immune cell types are affected. In one study, the effect of phenoxy herbicide exposure on immune function of exposed farmers was evaluated. This study showed a reduction in CD4 and CD8 cells, and proliferative responses were decreased to 12 days after exposure. The level of circulating cells, but not the proliferative responses, was normal by 70-90 days after exposure (Faustini et al., 1996). In studies of commercial herbicides, administration of 2,4dichlorophenoxyacetic acid (2,4-D) or Round Up for 26-28 days (twice weekly) did not affect immune function (Blakley, 1997; Blakley et aI., 1998). However, exposure to Tordon resulted in inhibition of humoral immunity at all exposure levels (Blakley, 1997).
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CHAPTER 36
Irnmunotoxicity of Pesticides
36.6 SUMMARY Many pesticides that are widely used and have great potential for occupational and public exposure have received only a cursory examination with regards to their immunotoxic potential. Many of the studies that are available were done during the early period of immunotoxicology and many reports do not state whether or not any other toxic signs were observed. In addition, very few compounds have received a thorough examination under well-defined treatment conditions. That is, many of the studies were done in multiple species, through various routes of administration, and using a variety of assays of immune function. For many of the compounds that have received extensive study, the site of action has been determined at the cellular level, and for some at the biochemical level, but the molecular site of action has not been determined for any of the pesticides discussed in this chapter.
36.7 FUTURE STUDIES Future studies should include a comprehensive examination of the immunotoxic potential for at least one compound in each class of pesticides. These studies should consider the route by which humans are exposed to a compound and an attempt should be made to mimic the human situation. In addition, the metabolic capacity for the compound of the test animal and humans should be considered when the test animal is being selected. Once the immunotoxic potential and parameters under which immune suppression or enhancement occur are established, the cellular, biochemical, and molecular site of action should be established. The establishment of a mechanism of action will allow (1) determination if this site is comparable between the test animal and the human immune system, (2) analysis of the potential for immunotoxicity for other compounds within the class by examination of structure-activity relationships at the site of action rather than the intact immune system, and (3) further dissection of the immune system through the use of toxic chemicals as biochemical tools. The biomarkers that should be used to evaluate toxity in human studies should be discerned from animal studies. For example, exposure of mice to malathion caused the release of histamine from basophilic cells resulting in a transient increase in the peripheral blood of exposed animals. The applicability of this biomarker to human populations could be assessed by measurement of workers acutely exposed to the agent. For example, the level of histamine in the blood of persons occupationally exposed could be assessed at baseline (e.g., after beginning of work after a long weekend) and then midday and at the end of the workday. By comparing baseline levels of histamine to those in blood taken during the workday, the utility of this parameter in human populations can be assessed. The difficulties with biomarkers of the immune system are the vast difference in immune function between people and within a person due to
changes other than environmental exposures. Until this variability can be controlled, accounted for, or understood, assessment of changes in immune function as a result of incidental exposure will be fraught with difficulty.
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CHAPTER 36
Immunotoxicity of Pesticides
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CHAPTER
37 Sensitive Population Groups Richard J. J ackson, Carol H. Rubin, and Michael McGeehin National Center for Environmental Health, Centers for Disease Control and Prevention
37.1 INTRODUCTION
37.1.1 EXPOSURE AMONG SENSITIVE POPULATION GROUPS
Exposure to pesticides does not affect all humans uniformly. Age, sex, genetic make-up, health status, and previous or concurrent exposures influence individual sensitivity. These parameters are interrelated and may combine to influence both qualitative and quantitative differences in sensitivity. Thus sensitivity and susceptibility are inextricably related. Defined population groups may be more susceptible to the toxic effects of pesticide exposure because of a greater inherent sensitivity and also because certain characteristics of the subpopulation may result in greater exposure. For example, young children are usually physiologically more sensitive than adults to a given pesticide exposure level (Mortensen et aI., 1996; Pope et aI., 1991). At the same time, the activities of young children (e.g., crawling on floor and engaging in hand-to-mouth behaviors) increase the likelihood of exposure in a household setting. This chapter will consider the question of sensitivity with regard to realistic exposure scenarios. Pesticides are often included in the list of chemicals that elicit adverse health outcomes in people identified as multiply chemically sensitive. Such people react to a wide variety of chemicals at exposure levels that are usually tolerated by the general population. This is a complex and controversial condition that is well-profiled in the literature (Cone and SuIt, 1992; Ziem and McTamney, 1997) and beyond the scope of this chapter. Sensitivity to pesticides is a multifaceted emerging issue. As new pesticide formulations become available, and as the number of mixtures of pesticides that are on the market increases, it is likely that parameters of sensitivity and susceptibility will modify. The following discussion attempts to summarize the observed effects of pesticides on humans and also the suspected effects, based upon animal models. Some of the categories of sensitivity addressed in this chapter (e.g., genetically imposed sensitivity) are presented elsewhere in this Handbook in greater depth and in a framework that goes beyond issues of sensitive populations. Handbook of Pesticide Toxicology Volume 1. Principles
Hill et al. (1995) measured 12 urinary metabolite pesticide residues, reflecting exposure to more than 30 different pesticides, in a sample of 1000 adults from the Third National Health and Nutrition Examination Survey, 1988-1994. Six of the pesticide residues were detectable in more than half of the population sampled (Hill et al., 1995). Para-nitrophenol (p-NP), the residue representing exposure to methyl parathion, was detected in 41 % of the samples. However, exposure rates are substantially higher among certain sensitive population groups than they are in the total popUlation. This was demonstrated in 1995 during an assessment of exposure to methyl parathion which had been illegally applied indoors. The urinary metabolite levels that define actual human exposure varied by age and sex (Fig. 37.1). People spending more time in the home (e.g., infants, the elderly, and the unemployed) had increased exposure potential and demonstrably elevated p-NP metabolite levels (Esteban et aI., 1996). At the same time, the reasons that kept many of these people at home (e.g., being of pre-school age, being pregnant, or having a chronic disease) also defined physiologically sensitive population groups.
37.2 QUALITATIVE AND QUANTITATIVE ASPECTS OF SENSITIVITY Pharmakokinetic differences in absorption, distribution, metabolism, and excretion are the basis for most subpopulation differences in pesticide sensitivity. This is particularly apparent among infants and children. 37.2.1 ABSORPTION Dermal absorption and gastrointestinal tract (GIT) absorption depend upon ratios of surface area to body weight, as well as on the characteristics of the absorptive surface. Children are more susceptible to dermal pesticide exposure because they have a greater surface area relative to their weight. The ratio of surface
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400 350 300 250 200 150 100 50
o
Age Group in Years Figure 37.1 Distribution of creatinine-adjusted para-nitrophenol (the urinary metabolite marker for methyl parathion) by sex and age among 306 persons exposed to the indoor application of methyl parathion in Lorain County, Ohio (1994).
area to weight of a newborn may be up to 2.5 times that of an adult. Thus similar exposure levels lead to significantly higher doses in infants than in adults (Rasmussen, 1979). Dermal characteristics may also increase sensitivity. A fullterm infant is born with a completely developed stratum corneum, the main barrier to percutaneous absorption (Lester, 1983). Human and animal studies of the antimicrobial hexachlorophene have shown no significant difference in percutaneous absorption between adults and full-term infants (Plueckhahn, 1973). Solomon and Fahrner (1977) found no difference in y-benzene hexachloride levels in the brain after topical application to newborn and adult guinea pigs. Similarly, Wester et al. (1977) reported no difference in percutaneous absorption of testosterone between newborn and adult Rhesus monkeys. In contrast, preterm infants appear to have increased percutaneous absorption for various chemicals (Shuman et aI., 1975). Serum levels of hexachlorophene have been higher in preterm than full-term infants bathed with the chemical (Greaves et aI., 1975; Tyra1a et aI., 1977). Dermal exposure of preterm infants to hexachlorophene has resulted in severe toxicity manifested by apnea, convulsions, and coma (Shuman et aI., 1975). Percutaneous absorptive changes associated with aging are not well-defined (Roskos et aI., 1986). In a study of percutaneous absorption of 14 pesticides in young and adult rats, 11 of the pesticides exhibited significant age- and dose-dependent differences in skin penetration. Among the pesticides that exhibited significant age-dependent differences in absorption, the highest young/adult penetration ratio was 1.53 (Hall et aI., 1988). However, this study did not include aged or senile rats. 37.2.1.1 GIT Absorption The extent to which pesticides are absorbed after oral exposure is determined by gastric acidity, emptying time, and intestinal motility (Morselli et aI. , 1980). Full-term neonates develop stomach acidity within 24 hours of birth, whereas their gastrointestinal motility is irregular and unpredictable (Morselli, 1976). Although there is little information on the rates of gastrointestinal absorption of pesticides in infants and children,
several studies have been done on the gastrointestinal absorption of pharmaceuticals in neonates. Morselli (1976) found that newborns absorb some pharmaceuticals (e.g., digoxin) the same as adults, some (e.g., ampicillin) to a greater extent, and some (e.g., phenobarbitol, chloramphenicol) to a lesser extent. The complexity of the gastrointestinal system makes it difficult to determine the effect of maturation on the absorption of pesticides (Warner, 1986). Hoffmann (1982) reviewed a number of studies on the gastrointestinal absorption of exogenous chemicals in experimental animals at various stages of development. He concluded that drugs which cross the intestinal epithelium by passive diffusion are absorbed from the gastrointestinal tract of immature animals at a higher rate than from the tract of adult animals. The higher absorption rate did not, however, result in higher blood levels because of a greater distribution volume in the neonate. 37.2.1.2 Transplacental Absorption Although adverse outcomes associated with fetal exposure to pesticides is not well defined, the passage of pesticides through the placenta is documented (Kreuzer et aI., 1997). In vitro placental perfusion by parathion resulted in significant transfer and 50% acetylcholinesterase depression (Benjaminov et aI., 1992). 37.2.2 DISTRffiUTION Once absorbed, pesticides are distributed within the plasma to various tissues and organs. Distribution varies according to the size of the individual, blood flow to the tissue, the pH of the body fluids, the distribution of body water in intracellular and extracellular compartments, and the extent of protein binding of the pesticide (Warner, 1986). These factors may compete with or complement each other. Increased sensitivity occurs when imbalances in distribution lead to a higher pesticide concentration in the target organ. An investigation of the passage of paraquat through the blood-brain barrier showed a high rate of entry into the brain of neonatal rats than into the brain of adult and elderly rats (Widdowson et aI. , 1996). A similar study reported higher paraquat concentrations among both very young
37.3 Age-Related Differences in Sensitivity and very old rats (Corasaniti et aI., 1991). Although there is little research directly relating to pesticide distribution, reviews of pharmaceutical distribution describe potential sensitivities among infants and children (Kearns and Reed, 1989).
785
resulting in high concentrations of the drug and its metabolites persisting in the blood. The decreased renal function of newborns, combined with their slower biometabolism, increases the likelihood that they will have pharmakokinetically based pesticide sensitivity.
37.2.3 METABOLISM
A major determinant of the toxicological effect of a pesticide is the manner in which it is handled in the body. Although the liver is the primary organ for xenobiotic metabolism, the kidney, intestines, lungs, and skin are also capable ofbiotransforming certain compounds, and alterations in these organs may lead to increased pesticide sensitivity (Reed and Besunder, 1989). Many such alterations are age-dependent. Neonates do not have full metabolic capacity; metabolic maturity is achieved at about 6 months of age in full-term infants and more slowly in premature infants (Warner, 1986). At birth, the metabolic deficiencies of the neonate include decreased hydroxylation and plasma enterase activity, a decreased amount of cytochrome P450, and a deficient glucorodination process (Done, 1964; Morselli, 1976; Morselli et aI., 1980; Warner, 1986). Metabolic activity may be further reduced by pathological conditions such as respiratory distress, cardiac insufficiency, hyperbilirubinemia, and low dietary intake (Morselli et aI., 1980). Once children's metabolic mechanisms are mature, their metabolic activity rises rapidly until they are about 3 years old. From 3 years to puberty, their metabolic activity slowly declines to adult levels and then further slows as they age (Warner, 1986). Neonates' metabolism of certain exogenous chemicals has also been found to be qualitatively different from that of adults. For example, in preterm and full-term infants, theophylline undergoes N-methylation to caffeine, whereas the opposite process occurs in adults (Reed and Besunder, 1989). Miller et al. (1976) described similar age-related differences in the metabolism of acetaminophen. Young rats store lindane more readily and for longer periods than do adult rats (Solomon and Fahrner, 1977). 37.2.4 EXCRETION
The kidney is the primary route of elimination for most pesticides. Although the kidneys of newborns have a full complement of glomeruli, their tubular size and mass are less than those of adults (Reed and Besunder, 1989). Adult kidney function is achieved at about one year of age (Morselli, 1976). Only a handful of studies have been done on the actual renal clearance of drugs and their metabolites from children and infants. The clearance of penicillin in premature infants was only 17% that in older children when corrected for surface area (Barnett et aI., 1949). Studies of chloramphenicol concentrations in the blood of neonates showed an inverse correlation between the age ofthe infant and the half-life of the drug (Reed and Besunder, 1989). The conjugation of chloramphenicol with glucoronic acid and renal excretion was lower among neonates,
37.3 AGE-RELATED DIFFERENCES IN SENSITIVITY As summarized previously and discussed at length elsewhere (Thomas, 1995), age-related sensitivity to pesticide exposure is a function of differences in size, in the maturity of biochemical and physiological functions in major body systems, and in body composition (proportions of water, fat, protein, and mineral content) (Forbes, 1987). In addition to having increased physiological sensitivity to pesticides, children are also at greater risk from pesticides because they have more opportunity for exposure (Go1dman, 1995; NRC, 1993; Thomas, 1995). 37.3.1 AGE-RELATED DIFFERENCES IN EXPOSURE
Sources and routes of exposure for children include ingestion (food and water), inhalation, and dermal contact inside and outside the home, and take-home-toxin exposure (Rogan, 1980). In Missouri, results of a telephone-interview survey of 238 families showed that nearly all families (97.8%) used pesticides at least one time per year in the home, garden, orchard, or yard; two-thirds used pesticides more than five times per year. More than 80% of families used pesticides during a pregnancy, and 70% used pesticides during the first 6 months of a child's life (Davis et aI., 1992). Other studies have quantitatively confirmed pesticide residue levels in home environments (Bradman et aI., 1997; Gurunathan et aI., 1998). The Gurunathan study showed that 2 weeks after certified applicators sprayed a single application of chlorpyrifos in apartment rooms, the pesticide continued to accumulate on children's toys and hard surfaces. Dietary exposure to pesticides varies both qualitatively and quantitatively with age. Infants and children consume more calories of food per unit of body weight than do adults, but they also consume far fewer types of foods. Consequently, infants and young children may consume much more of certain foods suspected of having elevated pesticide residues (e.g., apples), especially processed foods (e.g., apple juice, apple sauce). The younger the child, the less diverse the foods that he or she consumes (Thomas, 1995). 37.3.2 AGE-RELATED DIFFERENCES IN TOXICITY 37.3.2.1 Fetuses, Infants, and Young Children
Children are less able than adults to metabolize and excrete toxic substances (Drew et aI., 1983; Finhorn, 1982; Hudson
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et aI., 1972; Kalland, 1982; Widdowson and Dickerson, 1960). Their rapidly developing organ systems, especially the central nervous system, are highly susceptible to chemical interference. Exposures during brief but critical periods early in development can permanently alter the structure or function of an organ system (Swenberg and Fedtke, 1992; Vesell, 1982). Fetal sensitivity to pesticides may lead to unique transgenerational manifestations. In a study of cancer incidence among the progeny of male rats exposed to ethy lnitrosourea before mating, Tomatis et al. (1981) found that the fertility rate was lower and the preweaning mortality rate was higher in the experimental group than in the control group. Survival rates after weaning, as well as the total incidence of tumors, were similar in the progeny of treated males and of controls. However, neurogenic tumors were more frequent among progeny of rats in the experimental group (p = 0.08). These findings suggest that the toxic effects of mutagens can be transmitted through the male germ line. In an earlier study of the occurrence of tumors in first-, second-, and third-generation descendants of rats exposed to N-nitrosomethylurea during pregnancy, Tomatis et al. (1975) confirmed previous observations that exposure to a carcinogen during prenatal life may cause a genetic-chemical interaction leading to an increased cancer risk, which may persist for more than one generation. Other rat studies found that vinclozolin, a fungicide for fruits and vegetables, possessed antiandrogenic activity, causing feminization of male fetuses, sterility of adult males, and other developmental variations (Gray et aI., 1994; Ke1ce et aI., 1994). Pesticides can also differentially affect the fetus. In a review of the potential effects of chemical carcinogens during pregnancy and the perinatal period, Rice (1979) found that the fetal rat is up to several orders of magnitude more sensitive than the adult to certain carcinogens. Most such agents are direct-acting and independent of metabolism. Rice further reported that the fetus may be less vulnerable than the adult to those substances requiring enzyme-mediated metabolic conversion to chemically reactive derivatives to effect carcinogenesis. Newborns may exhibit the most extreme quantitative differences in sensitivity to chemicals because they are the group most anatomically and physiologically different from adults (Calabrese, 1986; Rodier, 1980). For example, infants may be at higher risk than adults of experiencing serious side effects from organophosphate poisoning because their nervous systems are not fully developed and because their ability to detoxify such agents is also generally lower than that of adults (Fuortes, 1993). Research suggests that neonatal and weanling animals have heightened sensitivity to organophosphates (Mendoza and Shields, 1977). Feeding methyl parathion to pregnant rats led to behavioral changes in offspring without visual signs of maternal toxicity (Gupta et aI., 1985). Lu et al. (1965) found that the LD 50 for malathion in rats increased from 124 mg/kg/day in the newborn to 386 mg/kg/day in the preweanling and to 925 mg/kg/day in the adult. On the other hand, Moretto et al. (1991) reported resistance to organophosphate-induced delayed polyneuropathy among young chickens. He attributed this lack
of sensitivity to a more efficient repair mechanism in developing chicks compared with that in hens. In addition, Pope et al. (1991) reported that cholinesterase activity recovers faster in neonates than in adults. Human Research Evidence of age-related sensitivity among human populations is generally based on ecological or caseseries reports (Garcfa-Rodrfguez et aI., 1996; Garry et aI., 1996; Kristensen et aI., 1997). An investigation of a cluster of cases of congenital abnormalities in a Hungarian village in 1989-1990 by Czeizel et al. (1993) showed that 11 of 15 (73 %) babies born live had congenital abnormalities; of these 11, four had Down's syndrome. The study identified a strong association between the occurrence of Down's syndrome and maternal consumption of fish containing elevated (100 mg/kg) levels of trichlorfon. In a statewide survey of 856 Iowa municipal drinking-water supplies in 1986-1987, Munger et al. (1997) concluded that newborn singletons in communities with herbicide-contaminated wells had elevated rates of intrauterine growth retardation. In a study exploring the role of chlorinated hydrocarbon pesticides in causing spontaneous abortions and premature labor, Saxena et al. (1980) found considerably higher amounts of organochlorine pesticide residues in the circulating blood and placental tissue of women undergoing spontaneous abortion or premature labor compared with the amount of these residues in women in full-term labor. Sharpe et al. (1995) reported a significantly increased risk for Wilms' tumor among Brazilian children whose parent(s) were occupationally exposed to pesticides during pregnancy. Currently, we lack sufficient evidence to define the adverse health effects of pesticide exposure on fetuses; more rigorous research is needed (Garcfa, 1998). 37.3.2.2 Adult and Aged Populations
Most occupational exposures to pesticide occur in healthy adults, and most safety regulations are standardized based upon this population. Scant research has been done that directly addresses pesticide sensitivity in geriatric populations. Although older people usually have higher levels of persistent chlorinated pesticides, these levels have not been associated with increased sensitivity (Sim et al., 1998). Similarly, aging of exposure pathways (e.g., changes in skin permeability) or metabolic pathways (e.g., reduction in hepatic biotransformation capabilities) has not been related to changes in pesticide sensitivity (Roskos et aI., 1986).
37.4 SEX-RELATED SENSITIVITY Sex-specific pesticide sensitivity, as well as sex-specific health effects, may occur after exposure to pesticides both during developmental and reproductive-age periods. Sex-based pharmacokinetic differences that may predispose a person to a sensitivity to pesticides have been described primarily in animal models. Sex-specific health effects of pesticides most often act
37.4 Sex-Related Sensitivity
through the mechanism of disrupting normal hormonal or endocrine relationships. Potential adverse outcomes include decreased sperm count, reduced sperm quality, infertility, premature menopause, or altered sexual behavior (Kavlock et aI., 1996). 37.4.1 ANIMAL STUDIES Research in animal models, primarily in rats, has suggested that males and females may be differentially sensitive to the toxicity of pesticides (Gaines, 1960, 1969). This difference is often attributed to a sex-specific variation in liver microsomal enzyme activity (Snawder and Chambers, 1991). The complexities and variations in pesticide metabolism are discussed in detail elsewhere in this text. Overall, sex does not appear to be a major modifier of the effects of pesticides. For example, chickens administered a single oral dose of an organophospate all experienced a delayed neurotoxicity but the ataxia was significantly more pronounced earlier in males than in females (Odom et al., 1992). However, the ultimate toxic effect of pesticide exposure was the same for both males and females despite variations in enzymatic activity that determined the temporal progression of the reaction. Animal studies of adverse birth outcomes have been used to evaluate the effect of pesticide exposure on paternally mediated birth defects (Anderson et aI., 1996). Potential mechanisms include direct germ-cell effects and indirect effects through transfer of chemicals to the mother via seminal fluid. Olshan and Faustman (1993) reviewed the experimental evidence for malemediated effects on offspring due to a variety of physical and chemical exposures and concluded that more basic animal and human research was necessary to determine the public health relevance of this exposure route. 37.4.2 HUMAN STUDIES: MALES Pesticides can adversely affect spermatogenesis, as documented by the potentially irreversible aspermia after occupational exposure to dibromochloropropane(DBCP) (Uihdetie, 1995; Whorton et aI., 1977). There has been concern that global increases in pesticide use may be causing a worldwide decline in the concentration and quality of sperm (Giwercman et aI., 1993). Empirical results have been conflicting, however. For example, two French studies that used different methods and different populations found varying results. Auger et al. (1995) analyzed sperm from 1351 healthy men and showed a decline in sperm quality among men in Paris from 1973 through 1992. However, Bujan et al. (1996) collected sperm samples from 302 healthy men in Toulouse and found no change in sperm concentrations over a similar period, 1977-1992. In a retrospective study of Danish farmers, Larsen et al. (1998) found no effect of pesticide exposure on male fecundity. Although there have been several meta analyses of more than 60 sperm health studies, the specific relationship between pesticides and male reproductive status remains undefined (Becker and Berhane, 1997).
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37.4.3 HUMAN STUDIES: FEMALES Sex-specific effects in women may manifest as infertility, endometriosis, or breast cancer. Smith et al. (1997) reported an increased risk for medically diagnosed infertility among women occupationally exposed to pesticides. Lebel et al. (1998) compared plasma concentrations of 11 chlorinated pesticides in 156 women in a case-control study. Neither crude geometric mean concentrations nor crude or adjusted means of the sum of chlordanes, nor the sum of dichlorodiphenyltrichloroethanes (DDEs) differed between case and control subjects. There was no significant linear trend in the adjusted odds ratios for endometriosis as organochlorine concentrations increased. Extensive reviews of the literature on organochlorines have concluded that there is not enough evidence to either support or reject the hypothesis that certain organochlorine compounds [such as DDT, DDE, polychlorinated biphenyls (PCBs), and tetrachloro-p-dioxin (TCDD)] increase the risk for breast, endometrial, or other human cancers (Adami et aI., 1995; Ahlborg et aI., 1995). Two other recent literature reviews concluded that DDE levels are not consistently elevated in women with breast cancer and that occupationally exposed women do not have an increased incidence of breast cancer (Safe and Zacharewski, 1997; Safe, 1997). However, individual epidemiological studies have not shown consistent results. Wolff et al. (1993) reported higher mean (P = 0.31) DDE levels for 58 female case subjects compared with 171 matched control subjects. After adjustment, researchers found that the DDE level was associated with a fourfold increase in relative risk for breast cancer. In a nested case-control study of 150 case and 150 control subjects, Krieger et al. (1994) found no differences (mean difference, 0.2 parts per billion; 95% confidence interval, -6.7, 7.2) between serum levels of DDE in case as compared with control subjects. However, DDE levels were higher among case subjects who were black than among control subjects who were black (mean difference, 5.7 parts per billion; 95% confidence interval, -3.3,14.8). Organochlorinelevels, in general, were significantly higher among black and Asian women than among white women. H0yer et al. (1998) reported that dieldrin level was associated with a significantly increased dose-related risk (adjusted odds ratio, 2.05; 95% confidence interval, 1.17,3.57) for breast cancer among 240 case subjects and 477 control subjects who were originally enrolled in the Copenhagen City Heart Study. B-Hexachlorocyclohexane increased the risk for breast cancer slightly but not significantly (adjusted odds ratio, 1.36; 95% confidence interval, 0.79, 3.57). There were no overall associations between the risk for breast cancer and DDT or DDE. Moysich et al. (1998) found no significant association between DDE exposure (odds ratio, 1.34; 95% confidence interval, 0.71, 2.55) and postmenopausal breast cancer among women enrolled in New York from 1986 to 1991. Hunter et al. (1997) reported that the median level of DDE was lower (P = 0.14) among 226 case subjects than among their matched pairs. Van't Veer et al. (1997) measured DDE in adipose tissue aspirated from the buttocks of women in several European countries and
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found 9.2% (P = 0.36) lower age-adjusted DDE concentrations among women with breast cancer compared with the concentrations found in control subjects. Pregnant or soon-to-be pregnant women represent a special category of people with a sensitivity to pesticides. Reports of pregnancy loss or of adverse birth outcomes associated with maternal pesticide exposure support the potential for risk (Sever et al., 1997). However, study results have been ambiguous and defined parameters of risk have not been identified (Nurminen, 1995; Savitz et al., 1989).
37.5 GENETIC PREDISPOSITION Genetic polymorphisms (i. e., genes with several variants, or alleles, the rarest of which occurs in at least 1% of the population) are important factors in determining an organism's sensitivity to environmental hazards (Barrett et al., 1997). Genetic sensitivity may influence the activation, detoxification, and cellular uptake of a pesticide (Table 37.1). Polymorphisms in genes that code for metabolic enzymes vary widely within a population and appear to play a primary role in altering sensitivity to environmental exposures (Garte, 1998; Takahashi et al., 1998; Wolff and Weston, 1997).
37.5.1 POLYMORPHISMS IN METABOLIC ENZYMES Pesticides are detoxified in the body through a series of reactions that ultimately biotransform them into water-soluble com-
pounds that are readily eliminated in the urine or feces. Several of the metaboloc enzymes that catalyze these reactions are known to be polymorphic. To date, the best known of these are the cytochrome P450 enzymes, the glutathione S-transferases (GSTs), and Paraoxonase (PONl). Although it is likely that polymorphisms in these enzymes influence human sensitivity to pesticide toxicity, research in this area is just beginning (Costa, 1996; Hirvonen, 1995; Hodgson and Levi, 1996).
37.5.1.1 Cytochrome P450 Enzymes The influence of cytochrome P450 enzyme polymorphisms has probably been best studied in patients with heart disease who demonstrate variable sensitivity to the anticoagulant effects of warfarin. Three allelic variants of the cytochrome P450 2C9 gene (CYP2C9) are known to affect the rate of oxidation of warfarin in humans (Yamazaki et al., 1998). One of the heterozygote forms was found to metabolize (S)-warfarin, but not (R)-warfarin, at a slower rate than wild-type CYP2C9. (S)Warfarin has a greater anticoagulant potency and, therefore, the CYP2C9 polymorphism may account for part of the variability observed between patients treated with therapeutic doses of warfarin (Takahashi et al., 1998). Pesticide exposure and family history have both been identified as risk factors for Parkinson's disease, and several studies have explored possible gene-toxin interactions in the etiology of this disease. In one study, subjects who reported pesticide exposure and who also had the polymorphism that metabolizes debrisoquine (CYP2D6 29B+) were three times more likely to have Parkinson's disease with dementia than Parkinson's with-
Table 37.1 Genetic Polymorphisms Potentially Involved in Modulating Human Susceptibility to Pesticide Toxicity Polymorphism
Effect of atypical phenotype
Pesticide class studied
Effect on toxicity
Reduced metabolism of xenobiotics
Coumarin rodenticides
Increased (Pope et aI., 1991)
Reduced metabolism of xenobiotics
Pesticides, in general
Increased (Hubble et aI., 1998)
Reduced metabolism of several
Organophosphates
Increased (Li et aI., 1993;
Fumigants
Increased (Ploemen et aI., 1995;
Metabolic enzymes Cytochrome P450 2C9 (CYP2C9) Cytochrome P450 2D6 (CYP2D6) Paraoxonase (PON1)
Mackness et aI., 1997)
organophosphates (OP) Glutathione S -transferase ()
Reduced metabolism of xenobiotics
Thier et aI., 1996)
(GSTTl) Glutathione S -transferase
7r
Reduced metabolism of xenobiotics
Pesticides, in general
Increased (Menegon et aI., 1998)
Reduced binding to OPs, which
Organophosphates
Increased (Fontoura-da-Silva and
(GSTP1) Target molecules Butyrylcholinesterase
increases availability of OP to bind
Chautard-Freire-Maia, 1996)
to acetylcholinesterase GABA-gated chloride ion channel Sodium ion channel
Reduced binding of pesticide to ion
Avermectins, cyclodienes
channels on neurons Reduced binding of pesticide to ion channels on neurons
Resistance to toxic effects seen in insects (Ffrench-Constant et aI., 1993)
DDT, pyrethroids
Resistance to toxic effects seen in insects (Knipple et aI., 1994)
37.5 Genetic Predisposition out dementia. Neither pesticide exposure nor the CYP2D6 polymorphism alone was found to be associated with increased risk for Parkinson's disease with dementia (Hubble et aI., 1998).
37.5.1.2 Glutathioue S-Transferases Several classes of GSTs are known to be polymorphic (e.g., GSTMl, GSTTl, GSTPl) and may influence sensitivity to pesticide toxicity. For instance, genomic loci coding for GSTs in the housefly have been implicated in mediating insecticide resistance (Zhou and Syvanen, 1997). Epidemiologic studies have investigated the potential health impact of interactions between GST polymorphisms and pesticide exposures. In a study of Parkinson's disease, patients with idiopathic Parkinson's disease were compared to healthy control subjects. Although pesticide exposure and family history were found to be risk factors for Parkinson's, none of the GST classes was independently associated with illness. However, when the analysis was restricted to participants who reported exposure to pesticides, the distribution of GSTPl genotypes varied significantly between patients with Parkinson's disease and healthy control subjects. This suggests that GSTPl polymorphisms might influence susceptibility to Parkinson's disease after pesticide exposure (Menegon et aI., 1998).
37.5.1.3 Paraoxonase Serum paraoxonase (PONl) is important in the detoxification of the organophosphate (OP) insecticides parathion, diazinon, and chlorpyrifos. Evidence in animals suggests that PONl protects them from poisoning by the OPs it metabolizes. Injection of purified PONl was shown to protect rodents against acute OP toxicity (Li et aI., 1993). In a recent study, PONl-deficient mice were much more sensitive than their wild-type littermates to the toxic effects of chlorpyrifos oxon, the activated form of chlorpyrifos, and to chlorpyrifos itself (Shih et aI., 1998). Studies in humans are just beginning. One recent study compared the PONl genotypes of patients with sporadic idiopathic Parkinson's disease to the genotypes of healthy control subjects. A specific allele of the PON 1 gene was found significantly more often in patients with Parkinson's disease than healthy control subjects. The authors suggested that PONl might influence susceptibility to Parkinson's disease by modulating the effect of the environmental neurotoxins, such as OP pesticides, that are metabolized by PONl (Kondo and Yamamoto, 1998).
37.5.2 POLYMORPHISMS IN TARGET AND TRANSPORT MOLECULES Genetic variability also occurs in molecules that bind to pesticides in the body. Polymorphisms in these genes influence sensitivity to pesticide toxicity by modulating the binding affinity between the pesticide and molecule. Target site (cholinesterases, ion-channel receptors) and transport molecules (serum albumin, P-glycoproteins) have been found
'/89
to exhibit genetic variation in various species and can affect individual sensitivity to pesticides. The cholinesterases (ChE) are the target sites for OP and carbamate pesticides. All vertebrates have two distinct ChEs, acetylcholinesterase (AcChE) and butyrylcholinesterase (BuChE), that hydrolyze the neurotransmitter acetylcholine. OPs bind irreversibly to ChEs, thus effectively inhibiting their ability to metabolize acetylcholine. Evidence suggests that BuChE acts as a scavenger, binding OPs and other toxins before they can bind to and inhibit AcChE. At least 20 polymorphisms in the BuChE gene have been found, most of which are functional and alter the activity of BuChE. The most common BuChE variant occurs in less than 5% of Europeans and Americans and in up to 11 % of people in other populations; therefore, genetic polymorphisms in BuChE may be important contributors to the observed variability in individual susceptibility to OPs and carbamates (Schwarz et aI., 1995). A recent epidemiologic study investigated the association between BuChE polymorphisms and AcChE activity in pesticide-exposed farmers. The farmers were classified as mildly poisoned if their RBC-AcChE activity levels less than 87.5%. The authors reported that the atypical BuChE phenotype was found significantly more often in farmers considered to be mildly poisoned than in the farmers with normal RBC-AcChE activity (::0::87.5%). These data suggest that different BuChE genetic variants offer differential protection against AcChE suppression by the OP and carbamate pesticides (Fontoura-da-Silva and and Chautard-Freire-Maia, 1996). Albumin also binds many compounds in the blood, therefore rendering them unavailable to bind to other molecules or to enter cells. Natural mutants of human serum albumin have different binding affinities for warfarin and may, therefore, affect sensitivity to the effects of warfarin (Vestberg et aI., 1992). Certain pesticides, such as 2,4-D, are bound almost exclusively to serum albumin in humans (Rosso et aI., 1998). Whether or not genetic variation in the binding affinity of serum albumin affects individual sensitivity to these pesticides has not been determined. P-glycoproteins are membrane-transport molecules that remove avermectins from cells in nematodes and mice, and are known to influence sensitivity to avermectin-induced neurotoxicity (Umbenhauer et aI., 1997; XU et aI., 1998). To investigate the role of P-glycoprotein in the transport of pesticides in humans, an experiment was conducted using murine melanoma cells transfected with the human MDRl gene, which codes for P-glycoprotein. The researchers found that a number of pesticides, including ivermectin and several organophosphate and organochlorine pesticides, are capable of binding to human P-glycoprotein. However, none of the pesticides, except endosulfan, was transported out of the cells (Bain and LeBlanc, 1996). Therefore, the role of P-glycoprotein, and consequently the MDRl gene, in altering susceptibility to pesticides in humans remains unclear (Umbenhauer et aI., 1997).
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37.6 HEALTH STATUS Compromised health status (e.g., malnutrition or immunosuppression) or preexisting disease (e.g., skin disease or seizure disorders) may increase sensitivity to pesticide exposure (Table 37.2). Often interrelationships between these factors are not precisely known. A case report by Solomon et al. (1995) described a neurotoxic reaction to a pesticide in an adult HIVseropositive patient. In this case, the routine application of 1% lindane for microscopically confirmed scabies precipitated seizures and encephalopathy. The authors attribute the adverse outcome to several independent factors, including diminished seizure threshold associated with HIV status, concurrent administration of chlorpromazine, and epidermal barrier dysfunction. Under normal circumstances, the skin acts as a barrier, preventing pesticides and other toxic ants from entering the circulation and causing systemic effects. However, Feldmann and Maibach (1974) applied 12 radiolabeled pesticides and herbicides to the forearm of human subjects and demonstrated that, even under normal circumstances, all of the chemicals tested were absorbed into the systemic circulation. Percutaneous absorption is a complicated process (Wester and Maibach, 1983), involving a series of steps that may be compromised by dermatitis, skin hydration status, or age. For example, increased trancutaneous absorption ascribed to congenital ichthyosiform erythroderma led to nausea and convulsions in a 1-year-old boy after a single application of 1% lindane cream (Friedman, 1987). Similarly, both Lange et al. (1981) and Tenenbein (1991) reported higher lindane blood levels among patients with generalized scabetic dermatitis than among treated patients whose skin was intact. Absorption is also facilitated by the hydration status of skin. Overly hydrated (e.g., recently bathed) skin facilitates systemic absorption. This observation is reported in the literature as a problem when bathing occurs before initiating topical treatment for mite or lice infestations. Overhydration can also potentially occur during the bathing or dipping of pets, especially if the pet owner or animal technician has skin excoriations and is treating
several pets within a short time. Cutaneous absorption is further enhanced if the pesticide is in a lipid vehicle (Solomon and Fahmer,1977). Disease processes that, like many pesticides, target the cholinergic system, may enhance pesticide sensitivity. Alzheimer's disease and several other neurodegenerative disorders (e.g., Parkinson's disease, Huntington's disease, and amyotrophic lateral sclerosis) involve pathological changes in the amount of available AcChE and BuChE (Rakonczay and Brimijoin, 1988). Increased senile plaque formation with decreased AcChE activity and increased circulating BuChE is seen in patients with Alzheimer's disease, and the severity of their dementia parallels the decline in available choline acetyltransferase in the brain (Perry et aI., 1978). These relationships between disease processes and pesticide sensitivity are complex and not well-defined. For example, anticholinergic drugs, such as carbamates and OPs, may cause symptoms similar to those observed in people with degenerative disease. At the same time, recent evidence suggests that use of selective anticholinergic agents may be therapeutic, especially for patients with Alzheimer's disease (Knapp et aI., 1994).
37.7 PREVIOUS OR CONCURRENT EXPOSURES THAT ALTER SENSITIVITY Individual sensitivity to pesticide exposure may be enhanced or diminished by simultaneous exposure to other chemicals. Such synergistic or antagonistic effects are best described in animal and plant models where research is conducted to maximize pesticide efficacy (Bemard and Philogene, 1993). Less is known about human exposure (Krishnan and Brodeur, 1994). In the workplace, simultaneous exposure to different classes of chemicals (e.g., pesticides and solvents) is often inevitable, and personal protective equipment is usually used to avoid interactive effects. Less understood, but increasingly likely, are
Table 37.2 Examples of Preexisting Conditions or Diseases that May Influence Pesticide Metabolism and Pesticide Health Effects Condition
Effect
Reference
Alzheimer's disease
Decreases available AcChE
Schwarz et al. (1995)
Parkinson's disease
Symptoms worsen with anti-ChE exposures
Ott and Lannon (1992)
Down's syndrome
Acetylcholinesterase enzyme deficiency
Percy et al. (1993)
Loiasis
Ivermectin may cause encephalopathy when patient has high
Gardon et al. (1997)
microfilaraemia Asthma
Potential for fatal asthma following pyrethrin inhalation
Wax and Hoffman (1994)
Carbamates enhance lung dysfunction in asthmatics
Senthilselvan et al. (1992)
Infectious hepatitis
Decreased plasma cholinesterase
Wagner (1995)
HIV
Decreased seizure threshold
Solomon et al. (1995)
Excoriated skin
Increases pesticide absorption
Ginsburg et al. (1977)
Increased absorption; seizures after lindane application
Friedman (1987)
Overly hydrated skin Ichthyosiform erythroderma
37.7 Previous or Concurrent Exposures that Alter Sensitivity
the pesticide interactions that may occur in the home due to complex product formulations and multiple product use. In a non-occupational setting the variety of pesticide exposures is usually not well-defined. For example, an analysis of moth- and mite-proofed household products, including vacuum cleaner bags, consistently found mixtures of pyrethroid, synergists, and (DEET) (Utunomiya et aI., 1997). In addition, low-level dietary exposure can occur through consumption of pesticide residues on fruits and vegetables (Melnyk et aI., 1997; Ratner et aI., 1983). Individuals with chronic multiple routes of exposure may be more sensitive to an acute pesticide exposure. This exposure may be further complicated if there is also a simultaneous exposure to nonpesticide cholinesteraseinhibiting chemicals. Such chemicals may potentiate the effect of organophosphate and carbamate exposure, thus potentially further enhancing sensitivity. 37.7.1 SYNERGISM Synergy is the interaction of two or more agents so that their combined effect is greater than the sum of their individual efforts. The result may be pesticide reactions or poisonings among population groups not previously identified as sensitive. A classic example of synergism is illustrated by the interaction of facilitatory toxins (e.g., snake venoms) that increase the release of acetylcholine, and anticholinesterase pesticides that inhibit the destruction of acetylcholine (Harvey and Karlsson, 1982). Similarly, simultaneous or concurrent use of organophospate pesticides may lead to intensified reactions if one pesticide interferes with the physiological detoxification of the other (Cohen and Murphy, 1974). Such an additive effect can also occur when, ostensibly, a single pesticide is being used. Baker et al. (1978) describe the enhanced toxicity of improperly stored malathion used by Pakistani applicators. The same synergistic toxic effect was observed in Belgium after people were exposed to pure malathion that had been stored for more than 5 years. The liquid pesticide had converted to a synergistic, and highly toxic, mixture of malathion and isomalathion (Dive et aI., 1994). Recent research suggests that pesticides frequently used indoors may synergistically interact with the materials that are in house dust (e.g., organic compounds and metals). Kang and Fang (1997) determined that several polycyclic aromatic hydrocarbons (PAHs) commonly found in house dust increased the potency of chlorpyrifos to inhibit AcChE by as much as 85% in vitro. However, there is a fine line between synergism and antagonism, and the combined effect of exposure to different chemicals is still not well-defined. The sequence of exposure may result in disparate outcomes. For example, exposure to the serine esterase inhibitor phenylmethylsulfonyl fluoride (PMSF) before OP exposure can protect a person against organophosphorusinduced delayed neurotoxicity (OPIDN). However, PMSF administration after OP exposure will exacerbate OPIDN (Pope and Padilla, 1990). Research in sheep has suggested that defined
791
periods must elapse between applications of different pesticides to avoid otherwise toxic pesticide interactions (Mohammad and St. Omer, 1985). It is likely that pesticide exposure occurs in combination with exposures to other agents, such as solvents (Petrelli et aI., 1993). Table 37.3 briefly lists results of animal and human studies which show that such combined exposures may alter sensitivity. People who are sensitive to solvents may be at increased risk for illness when they are exposed to low levels of pesticides. 37.7.1.1 Poverty and Pesticides Poverty may increase the potential for exposure to pesticides and also may be the underlying reason for nutritional deficits that heighten sensitivity. Poor people are more likely than those who are not poor to have higher residential exposures from heavy spraying for severe pest infestations in substandard housing (Moses et aI., 1993). Spraying by unlicensed applicators also increases the likelihood that inappropriate strengths or illegal formulations are being used (Esteban et aI., 1996). Often, itinerant and migrant workers receive additional exposures while working in fields recently sprayed with pesticides. The pesticides or pesticide residues are often carried into the home on work clothing. Research has also shown that less affluent people are more likely than affluent people to have higher levels of chlorinated pesticides stored in their bodies (Davies et aI., 1972). Other studies have shown that black people have higher chlorinated pesticide levels than white people (Finklea et aI., 1972; Rogan et aI., 1986). People living in poverty may be more sensitive to pesticide exposure because of nutritional factors such as low body fat, micronutrient imbalance, or protein deficiency. Although scant research has been done among human populations, multiple rat studies suggest that starvation and crowding increase the toxic effects of pesticides (Hayes and Laws, 1991). The combined effect of multiple environmental exposures that interact with pesticides (e.g., lead and tobacco smoke) is not known. 37.7.2 POTENTIATION Potentiation occurs when one exogenous chemical enhances or increases the effect of another. For instance, several overthe-counter and prescription medications potentiate the anticoagulant effect of warfarin. Examples of these potentiators include levamisole (Wehbe and Warth, 1996), ginseng (Janetzky and Morreale, 1997), fluoxetine (Dent and Orrock, 1997), fluoroquinolones (Jolson et aI., 1991), and Danshen (Yu et aI., 1997). People who are exposed to warfarin and who are concurrently exposed to any of the drugs just listed may experience adverse health effects because of potentiation. Although more commonly used therapeutically, warfarin is also an anticoagulant rodenticide. The 1997 Annual Report of the American Association of Poison Control Centers Toxic Exposure Surveillance System categorized 93.5% of the human poison exposures reported by poison control centers in the United States in
792
CHAPTER 37
Sensitive Population Groups
Table 37.3 Examples of Interaction between Environmental Exposures and Pesticide Exposure Exposure
Example of effect
Reference
Solvents
Lipid solvents increase dermal
Solomon and Fahrner (1977)
absorption (animal) Poverty
Low body fat or starvation increases susceptibility to acute intoxication from organochlorines and organophosphates (animal) Seizures following routine use of lindane in malnourished
Clarke and Clarke (1975), Iyaniwura (1990) Pramanik and Hansen (1979)
child; increases topical effects of y-benzene hexachloride Selenium deficiency Heavy metals
Potentiates paraquat-induced liquid peroxidation of lung
Glass et al. (1985)
tissue Dithiocarbamates increase movement of lead across
Oskarsson and Lind (1983)
blood-brain barrier (animal) Alcohol
Potentiates endosulfan hepatotoxicity (animal)
Singh and Pandey (1991)
Increases methyl parathion-induced chromosomal
Kumar et al. (1993)
aberrations (human) Antagonism of the acute toxicity of parathion (animal)
O'Shaughnesy and Sultatos (1995)
Smoking
Increases chromosomal abberations (human)
Rupa et al. (1989),
Chloroform and
Dithiocarbamates decrease bioactivation and decrease
Gopinath and Ford (1975)
Scarpato et al. (1996, 1997) carbon tetrachloride
toxic effects of chloroform and carbon tetrachloride (animal) Kepone increases hepatotoxicity of chloroform (animal)
1997 (Litovitz et aI., 1998). Of the 2,192,088 human exposures reported, 14,795 cases were associated with exposure to anticoagulants; 90% of the anticoagulant poisonings were among children under six years of age. Thus it appears that, when anticoagulant poisonings occur, they are often among young children. 37.7.3 NONPESTICIDE CHEMICALS THAT INHIBIT CHOLINESTERASE People can also exhibit sensitivity to pesticides when nonpesticide ChE-inhibiting chemicals alter the effect of OPs and carbamates and vice-versa. For example, Ware et al. (1990) describe a prolonged response to the neuromuscular blockade effects of succinylcholine in a patient whose exposure to an organophosphate pesticide had occurred within the week that the exposure occurred. OP and carbamate pesticides inhibit AcChE and interfere with nerve conduction. When a person is exposed to these pesticides, limited amounts of circulating BuChE (or pseudocholinesterase) may act as toxin scavengers. This action potentially prevents the pesticides from interacting with AcChE and disrupting nerve conduction (Neville et aI., 1990a, b; Schwarz et aI., 1995). However, such a potentially protective effect may be lost if the BuChE is reacting to synthetic or naturally occurring chemicals that can also act as ChE substrates. If sufficient quantities of these exogenous drugs are present, then a person may be more susceptible to lower levels of OP and carbamate insecticides than he or she would be otherwise. Nonpesticide
Hewitt et al. (1986)
chemicals that inhibit cholinesterase may also potentiate the effect of pesticide exposure. This sensitivity is similar to the sensitivity imposed by the genetic BuChE allele discussed in Section 40.4. Intuitively, if BuChE is an important first line of defense against low-level pesticide exposure, then people with decreased levels of BuChE will be at greater risk of poisoning from cholinesterase inhibitors than people whose levels are within normal limits (Anton, 1988; Cregler, 1989; Devenyi, 1989). Nonetheless, the opposite effect was observed when the corollary of this scenario was tested. Rats were pretreated with an OP (tetraisopropyl pyrophosphoramide) and then given a low toxic dose of cocaine (a BuChE substrate). Contrary to expectations, significantly more fatalities occurred among the rats that did not receive OP pretreatment than among those that did (Kambam et al., 1992a). Results of a similar experiment with pigs showed that inhibition of BuChE accelerated the metabolism of cocaine through an alternative route (Kambam et aI., 1992b). Although interaction clearly occurs when people are simultaneously exposed to different cholinesterase-binding chemicals, the direction of such an interaction and the manifestation of adverse health effects is not well-demonstrated; further research is necessary to understand these relationships. Rats that were pretreated with monoclonal antibodies to rat AcChE and then given a dose of an OP did demonstrate reduced AcChE activity but did not exhibit neurobehavioral changes (Padilla et aI., 1992). To further complicate matters, other drugs (e.g., tricyclic antidepressants) may act as anticholinergic blocking agents (Bal et aI., 1990). These drugs act like atropine on all
37.8 Implications for Public Health
793
Table 37.4 Naturally Occurring and Synthetic Cholinesterase Substrates Type of chemical
References
Examples
Analgesics
Aspirin, acetaminophen
Valentino et al. (1981)
Narcotics
Cocaine, procaine, heroin
Isenschmid et al. (1989),
Fungal antibiotics
Puromycin
Hersh (1981)
Gatley (1991) Onchidal
Mollusc secretion
Abrahamson et al. (1989)
Reptile polypeptides
FascicuIin
Karlsson et al. (1985)
Metals
Aluminum, scandium, yttrium
Marquis and Lerrick (1982)
Neuromuscular relaxants
Succinylcholine
Neville et al. (1990a)
Physostigmine derivatives
Neostigmine, demecarium, pyridostigmine
Shaw et al. (1985) Coleman et al. (1987)
muscarinic sites and may mask the effects of pesticide exposure (Marrs, 1993). Although this action essentially protects a person from all but nicotinic and central-nervous-system effects, it increases the likelihood of misdiagnosis of pesticide poisoning. Exogenous cholinesterase inhibitors may also have the effect of increasing the overall effective dose. Low-level exposure to OPs may result in a person exhibiting neurological symptoms when other cholinesterase inhibitors are also in the blood. Health care providers must be aware of the variety of environmental and recreational chemicals (Table 37.4) that are cholinesterase inhibitors because, even if they do not magnify or mask clinical signs of pesticide poisoning, their presence may affect treatment decisions and treatment efficacy.
37.8 IMPLICATIONS FOR PUBLIC HEALTH Historically, risk assessment models for pesticide toxicity have been formulated to reflect the pharmacokinetic patterns of the adult male. Pesticide regulations defining permissible levels of exposure have been based upon these risk assessments and have been primarily concerned with occupational exposures that are most relevant to a healthy working adult male. It is only recently that exposure levels in children, parous women, and other sensitive population groups have been acknowledged. In 1988, Congress requested the National Academy of Sciences (NAS) to evaluate the V.S. Environmental Protection Agency's (EPA) existing risk assessment methods to determine if they were adequately addressing the exposure and the potential risk that pesticides may pose to infants and children. That same year, the Children's Health Protection Advisory Committee (which had been formed to advise, consult with, and make recommendations to the EPA regarding reevaluation of existing EPA regulations to better protect children's health) reported that selected pesticides (triazines, OPs, and carbamates) were among the top five priority issues (Reigart, 1988). In 1993, the NAS report, Pesticides in the Diet of Infants and Children, recommended that future studies should compare metabolism and toxicity in adult and immature animals and that carcinogenicity
research should consider in utero exposure. EPA responded by revising its residue guidelines (Fenner-Crisp, 1995). In 1996, the Food Quality Protection Act (FQPA) was passed. It sought to provide added protection against pesticide risk, especially for infants and children, by setting a lO-year schedule for EPA to reevaluate 10,000 existing tolerances for pesticide residues on food (Goldman, 1998). In the absence of reliable data on children's toxicity, the FQPA directed EPA to use an extra lO-fold safety factor in its risk assessments to ensure that the greater sensitivity of infants and children would be addressed. 37.8.1 INFORMATION GAPS AND RESEARCH NEEDS Despite legislative progress in defining risk, most risk assessment models continue to rely upon animal data which may not accurately represent human exposure or human health outcomes. For example, based upon 2-year low-dose rat studies, proteinuria was identified as the most sensitive toxic end point for chlordecone exposure; however, even in high-dose occupational exposures, proteinuria has not been observed as a human health outcome (Guzelian, 1992). Similarly, animal models may not be able to adequately identify the effect of pesticide exposure on endocrine function during crucial periods of neurological development (Tilson, 1998). As discussed throughout this chapter, sensitive population groups experience numerous routes of exposure to an increasing variety of pesticide formulations. Risk assessment should reflect such exposure realities. Measuring pesticide metabolites in human biological specimens is the most accurate method of defining human exposure. Studies that actually quantify human exposure can more accurately associate adverse health effects with specific pesticides. Traditional risk assessment has used high-dose exposures (e.g., animal models or accidental high-dose occupational exposures) to predict low-dose health outcomes. This scenario may not be relevant to exposures that impact the endocrine system, especially during periods of rapid development (Sheehan and
794
CHAPTER 37
Sensitive Population Groups
vom Saal, 1997). Both the FQPA and the Safe Drinking Water Act require EPA to develop a screening program to test whether exogenous chemicals affect humans in the same way as a naturally occurring estrogen. Research activities must be focused on those populations (e.g., fetuses) most vulnerable to exogenous estrogens. Responsible use of pesticides and rational pesticide regulations must consider sensitive human population groups. Further research that includes assessment of all exposure routes, human biomonitoring of exposure, and the timing of exposure relative to development stages is necessary.
ACKNOWLEDGMENT The authors acknowledge Kim Blindauer, Amanda Niskar, Luke Naeher, and Anyana Banerjee for contributing both time and expertise to the preparation of this chapter.
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39,394-395. Thier, R., Pemble, S. E., Kramer, H., Taylor, J. B., Guengerich, P. F., and Ketterer, B. (1996). Human glutathioe S-transferase TI-I enhances mutagenicity of 1,2-dibromoethane, dibromomethane and 1,2,3,4-diepoxybutane in Salmonella typhimurium. Carcinogenesis 17, 163-166. Thomas, R. D. (1995). Age-specific carcinogenesis: Environmental exposure and susceptibility. Environ. Health Perspect. 103(Suppl. 6), 45-48. Tilson, H. A. (1998). Developmental neurotoxicology of endocrine disruptors and pesticides: Identification of information gaps and research needs. Environ. Health Perspect. 106,807-811. Tomatis, L., Cabral, 1. R. P., Likhachev, A. 1., and Ponomarkrov, V. (1981). Increased cancer incidence in the progeny of male rats exposed to ethylnitrosourea before mating. Int. J. Cancer 28, 475-478. Tomatis, L., Hilfrich, J., and Turusov, V. (1975). The occurrence of tumors in Fl, F2 and F3 descendants of BD rats exposed to N -nitrosomethylurea during pregnancy. Int. J. Cancer 15, 385-390. Tyrala, E. E., Hillman, L. S., Hillman, R. E., and Dodson, W. E. (1977). Clinical pharmacology of hexachlorophene in newborn infants. J. Pediatrics 91, 481-486. Umbenhauer, D. R., Lankas, G. R., Pippert, T. R., Wise, D. L., Cartwright, M. E., Hall, S. J., and Beare, C. M. (1997). Identification of a p-glycoproteindeficient subpopulation in the CF-l mouse strain using a restriction fragment length polymorphism. Toxicol. Appl. Pharmacol. 146, 88-94. Utunomiya, A., Hasegawa, K, and Mori, Y. (1997). Analysis of pyrethroid pesticides, synergists and repellent in moth/mite-proofed household products and their mutagenicity. Jpn. J. Toxicol. Environ. Health 43, 366-375. Valentino, R. J., Lockridge, 0., Eckerson, H. w., and La Du, B. N. (1981). Prediction of drug sensitivity in individuals with atypical serum cholinesterase based on in vitro biochemical studies. Biochem. Pharmacol. 30, 1643-1649. van't Veer, P., Lobbezoo, 1. E., Martfn-Moreno, Guallar, E., G6mez-Aracena, J., Kardinaal, A. F. M., Kohlmeier, L., Martin, B. c., Strain, J. J., Thamm, M., van Zoonen, P., Baumann, B. A., Huttunen, J. K, and Kock, F. J. (1997). DDT (dicophane) and postmenopausal breast cancer in Europe: Casecontrol study. Br. Med. J. 315, 81-85. Vesell, E. S. (1982). Dynamically interacting genetic and environmental factors that affect the reponse of developing individuals to toxicants. In "Banbury Report: Environmental Factors in Human Growth and Development" (V. R. Hunt, M. K. Smith, and D. Worth, eds.). Cold Spring Harbor Laboratory Press, Cold Spring Harbor, NY. Vestberg, K., Galliano, M., Minchiotti, L., and Kragh-Hansen, U. (1992). Highaffinity binding of warfarin, salicylate and diazepam to natural mutants of
human serum albumin modified in the c-terminal end. Biochem. Phamacol. 44, 151-1521. Wagner, S. L. (1995). Pitfalls in the laboratory diagnosis of pesticide intoxication. J. AOC Int. 78, 1-3. Ware, M. R., Frost, M. L., Berger, J. J., Stewart, R. B., and DeVane, L. C. (1990). Electroconvulsive therapy complicated by insecticide ingestion. J. Clin. Psychopharmacol. 10,72-73. Warner, A. (1986). Drug use in the neonate: Interrelationships of pharmacokinetics, toxicity, and biochemical maturity. Clin. Chem. 32, 721-727. Wax, P. M., and Hoffman, R. S. (1994). Fatality associated with inhalation of a pyrethrin shampoo. J. Toxicol.-Clinical Toxicol. 32, 457-460. Wehbe, T. w., and Warth, J. A. (1998). A case of bleeding requiring hospitalization that was likely caused by an interaction between warfarin and levamisole. Clinical Pharmacol. Ther. 59, 360--362. Wester, R. C., and Maibach, H. 1. (1983). Cutaneous pharmacokinetics: 10 steps to percutaneous absorption. Drug Metab. Rev. 14, 169-205. Wester, R. C., Noonan, P. K, Cole, M. P., and Maibach, H. 1. (1977). Percutaneous absorption of testosterone in the newborn rhesus monkey: Comparison to the adult. Pediatrics 11, 737-739. Whorton, D., Krauss, R. M., Marshall, S., and Milby, T. H. (1977). Infertility in male pesticide workers. Lancet (Dec. 17), 1259-1261. Widdowson, E. M., and Dickerson, J. W. T. (1960). The effect of growth and function on the chemical composition of soft tissues. Biochem. J. 77, 30-43. Widdowson, P. S., Farnworth, M. J., and Simpson, M. G. (1996). Influence of age on the passage of paraquat through the blood-brain barrier in rats: A distribution and pathological examination. Hum. Exp. Toxicol. 15,231236. Wolff, M. S., and Weston, A. (1997). Breast cancer risk and environmental exposures. Environ. Health Perspect. 105(Suppl. 4), 891-896. Wolff, M. S., Toniolo, P. G., Lee, E. w., Rivera, M., and Dubin, N. (1993). Blood levels of organochlorine residues and risk of breast cancer. J. Natl. Cancer Inst. 85, 648-652. Xu, M., Molento, M. B. w., Ribeiro, P., and Beech, R. P. R. (1998). Ivermectin resistance in nematodes may be caused by alteration of P-glycoprotein homolog. Mol. Biochem. Parasitol. 91, 327-335. Yamazaki, H. N. K C. K, Ozawa, N., Kawai, T., Suzuki, Y., Goldstein, J. A., Guengerich, P., and Shimada, T. (1998). Comparative studies on the catalytic roles of cytochrome P450 2C9 and its Cys- and Leu-variants in the oxidation of warfarin, flurbiprofen, and diclofenac by human liver microsomes. Biochem. Pharmacol. 56,243-251. Yu, C. M., Juliana, C. N., and Sanderson, J. E. (1997). Chinese herbs and warfarin potentiation by 'Danshen.' J. Internal Med. 241, 337-339. Zhou, Z. H., and Syvanen, M. (1997). A complex glutathione transferase gene family in the housefly Musca demestica. Mol. General Genetics 256, 187194. Ziem, G., and McTamney, J. (1997). Profile of patients with chemical injury and sensitivity. Environ. Health Perspect. 105(Suppl. 2),417-435.
CHAPTER
38 Pesticide Residues in Food and Cancer Risk: A Critical Analysis Lois Swirsky Gold, Thomas H. Slone, Bruce N. Ames, and Neela B. Manley University of California, Berkeley Ernest Orlando Lawrence Berkeley National Laboratory
38.1 INTRODUCTION Possible cancer hazards from pesticide residues in food have been much discussed and hotly debated in the scientific literature, the popular press, the political arena, and the courts. Consumer opinion surveys indicate that much of the U.S. public believes that pesticide residues in food are a serious cancer hazard (Opinion Research Corporation, 1990). In contrast, epidemiologic studies indicate that the major preventable risk factors for cancer are smoking, dietary imbalances, endogenous hormones, and inflammation (e.g., from chronic infections). Other important factors include intense sun exposure, lack of physical activity, and excess alcohol consumption (Ames et aI., 1995). The types of cancer deaths that have decreased since 1950 are primarily stomach, cervical, uterine, and colorectal. Overall cancer death rates in the United States (excluding lung cancer) have declined 19% since 1950 (Ries et aI., 2000). The types that have increased are primarily lung cancer [87% is due to smoking, as are 31 % of all cancer deaths in the United States (American Cancer Society, 2000)], melanoma (probably due to sunburns), and non-Hodgkin's lymphoma. If lung cancer is included, mortality rates have increased over time, but recently have declined (Ries et aI., 2000). Thus, epidemiological studies do not support the idea that synthetic pesticide residues are important for human cancer. Although some epidemiologic studies find an association between cancer and low levels of some industrial pollutants, the studies often have weak or inconsistent results, rely on ecological correlations or indirect exposure assessments, use small sample sizes, and do not control for confounding factors such as composition of the diet, which is a potentially important confounding factor. Outside the workplace, the levels of exposure to synthetic pollutants or pesticide residues are low and rarely Handbook of Pesticide Toxicology Volume 1. Principles
seem toxicologically plausible as a causal factor when compared to the wide variety of naturally occurring chemicals to which all people are exposed (Ames et aI., 1987, 1990a; Gold et aI., 1992). Whereas public perceptions tend to identify chemicals as being only synthetic and only synthetic chemicals as being toxic, every natural chemical is also toxic at some dose, and the vast proportion of chemicals to which humans are exposed are naturally occurring (see Section 38.2). There is, however, a paradox in the public concern about possible cancer hazards from pesticide residues in food and the lack of public understanding of the substantial evidence indicating that high consumption of the foods that contain pesticide residues-fruits and vegetables-has a protective effect against many types of cancer. A review of about 200 epidemiological studies reported a consistent association between low consumption of fruits and vegetables and cancer incidence at many target sites (Block et aI., 1992; Hill et aI., 1994; Steinmetz and Potter, 1991). The quarter of the population with the lowest dietary intake of fruits and vegetables has roughly twice the cancer rate for many types of cancer (lung, larynx, oral cavity, esophagus, stomach, colon and rectum, bladder, pancreas, cervix, and ovary) compared to the quarter with the highest consumption of those foods. The protective effect of consuming fruits and vegetables is weaker and less consistent for hormonally related cancers, such as breast and prostate. Studies suggest that inadequate intake of many micronutrients in these foods may be radiation mimics and are important in the carcinogenic effect (Ames, 2001). Despite the substantial evidence of the importance of fruits and vegetables in prevention, half the American public did not identify fruit and vegetable consumption as a protective factor against cancer (U.S. National Cancer Institute, 1996). Consumption surveys, moreover, indicate that 80% of children and adolescents in the United States (Krebs-Smith et
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aI., 1996) and 68% of adults (Krebs-Smith et aI., 1995) did not consume the intake of fruits and vegetables recommended by the National Cancer Institute (NCI) and the National Research Council: five servings per day. One important consequence of inadequate consumption of fruits and vegetables is low intake of some micronutrients. For example, folic acid is one of the most common vitamin deficiencies in people who consume few dietary fruits and vegetables; folate deficiency causes chromosome breaks in humans by a mechanism that mimics radiation (Ames, 2001; BIount et aI., 1997). Approximately 10% of the V.S. population (Senti and Pilch, 1985) had a lower folate level than that at which chromosome breaks occur (BIount et aI., 1997). Folate supplementation above the recommended daily allowance (RDA) minimized chromosome breakage (Fenech et al., 1998).
Given the lack of epidemiological evidence to link dietary synthetic pesticide residues to human cancer, and taking into account public concerns about pesticide residues as possible cancer hazards, public policy with respect to pesticides has relied on the results of high-dose, rodent cancer tests as the major source of information for assessing potential cancer risks to humans. This chapter examines critically the assumptions, methodology, results, and implications of cancer risk assessments of pesticide residues in the diet. Our analyses are based on results in our Carcinogenic Potency Database (CPDB) (Gold et aI., 1997b, 1999; http://potency.berkeley.edu), which provide the necessary data to examine the published literature of chronic animal cancer tests; the CPDB includes results of 5620 experiments on 1372 chemicals. Specifically, the following are addressed in the section indicated: Section 38.2. Human exposure to synthetic pesticide residues it the diet compared to the broader and greater exposure to natural chemicals in the diet Section 38.3. Cancer risk assessment methodology, including the use of animal data from high-dose bioassays in which half the chemicals tested are carcinogenic Section 38.4. Increased cell division as an important hypothesis for the high positivity rate in rodent bioassays and implications for risk assessment Section 38.5. Providing a broad perspective on possible cancer hazards from a variety of exposures to rodent carcinogens, including pesticide residues, by ranking on the HERP (human exposure/rodent potency) index Section 38.6. Analysis of possible reasons for the wide disparities in published risk estimates for pesticide residues in the diet Section 38.7. Identification and ranking of exposures in the V.S. diet to naturally occurring chemicals that have not been tested for carcinogenicity, using an index that takes into account the acutely toxic dose of a chemical (LDso) and average consumption in the V.S. diet Section 38.8. Summary of carcinogenicity results on 193 active ingredients in commercial pesticides.
38.2 HUMAN EXPOSURES TO NATURAL AND SYNTHETIC CHEMICALS Current regulatory policy to reduce human cancer risks is based on the idea that chemicals that induce tumors in rodent cancer bioassays are potential human carcinogens. The chemicals selected for testing in rodents, however, are primarily synthetic (Gold et aI., 1997a, b, c, 1998, 1999). The enormous background of human exposures to natural chemicals has not been systematically examined. This has led to an imbalance in both data and perception about possible carcinogenic hazards to humans from chemical exposures. The regulatory process does not take into account (1) that natural chemicals make up the vast bulk of chemicals to which humans are exposed; (2) that the toxicology of synthetic and natural toxins is not fundamentally different; (3) that about half of the chemicals tested, whether natural or synthetic, are carcinogens when tested using current experimental protocols; (4) that testing for carcinogenicity at near-toxic doses in rodents does not provide enough information to predict the excess number of human cancers that might occur at low-dose exposures; and (5) that testing at the maximum tolerated dose (MTD) frequently can cause chronic cell killing and consequent cell replacement (a risk factor for cancer that can be limited to high doses) and that ignoring this effect in risk assessment can greatly exaggerate risks. We estimate that about 99.9% of the chemicals that humans ingest are naturally occurring. The amounts of synthetic pesticide residues in plant foods are low in comparison to the amount of natural pesticides produced by plants themselves (Ames et aI., 1990a, b; Gold et aI., 1997a). Of all dietary pesticides that Americans eat, 99.99% are natural: They are the chemicals produced by plants to defend themselves against fungi, insects, and other animal predators. Each plant produces a different array of such chemicals (Ames et aI., 1990a, b). We estimate that the daily average V.S. exposure to natural pesticides in the diet is about 1500 mg and to burnt material from cooking is about 2000 mg (Ames et aI., 1990b). In comparison, the total daily exposure to all synthetic pesticide residues combined is about 0.09 mg based on the sum of residues reported by the V.S. Food and Drug Administration (FDA) in its study of the 200 synthetic pesticide residues thought to be of greatest concern (Gunderson, 1988; V.S. Food and Drug Administration, 1993a). Humans ingest roughly 5000-10,000 different natural pesticides and their breakdown products (Ames et aI., 1990a). Despite this enormously greater exposure to natural chemicals, among the chemicals tested in long-term bioassays in the CPDB, 77% (1050/1372) are synthetic (i.e., do not occur naturally) (Gold and Zeiger, 1997; Gold et aI., 1999). Concentrations of natural pesticides in plants are usually found at parts per thousand or million rather than parts per billion, which is the usual concentration of synthetic pesticide residues. Therefore, because humans are exposed to so many more natural than synthetic chemicals (by weight and by number), human exposure to natural rodent carcinogens, as defined
38.2 Human Exposures to Natural and Synthetic Chemicals
801
Table 38.1 Carcinogenicity Status of Natural Pesticides Tested in Rodentsa Carcinogensb : N =37
Acetaldehyde methylformylhydrazone, allyl isothiocyanate, arecoline·HCI, benzaldehyde, benzyl acetate, caffeic acid, capsaicin, catechol, clivorine, coumarin, crotonaldehyde, 3,4-dihydrocoumarin, estragole, ethyl acrylate, N2-y-glutamyl-p-hydrazinobenzoic acid, hexanal methylformylhydrazine, p-hydrazinobenzoic acid·HCI, hydroquinone, I-hydroxyanthraquinone, lasiocarpine, d-limonene, 3-methoxycatechol, 8-methoxypsoralen, N -methyl-N -formylhydrazine, a-methylbenzyl alcohol, 3-methylbutanal methylformylhydrazone, 4-methyicatechol, methylhydrazine, monocrotaline, pentanal methylformylhydrazone, petasitenine, quercetin, reserpine, safrole, senkirkine, sesamol, symphytine
Noncarcinogens: N =34
Atropine, benzyl aicohol, benzyl isothiocyanate, benzyl thiocyanate, biphenyl, d-carvone, codeine, deserpidine, disodium glycyrrhizinate, ephedrine sulfate, epigallocatechin, eucalyptol, eugenol, gaIlic acid, geranyl acetate, J'i-N-[y-I(+)-glutamyI1-4hydroxymethylphenylhydrazine, glycyrrhetinic acid, p-hydrazinobenzoic acid, isosafrole, kaempferol, dl-menthol, nicotine, norharman, phenethyl isothiocyanate, pilocarpine, piperidine, protocatechuic acid, rotenone, rutin sulfate, sodium benzoate, tannic acid, I-trans-8 9-tetrahydrocannabinol, turmeric oleoresin, vinblastine
aPungal toxins are not included. bThese rodent carcinogens occur in absinthe, allspice, anise, apple, apricot, banana, basil, beet, black pepper, broccoli, Brussels sprouts, cabbage, cantaloupe, caraway, cardamom, carrot, cauliflower, celery, cherries, chili pepper, chocolate, cinnamon, cloves, coffee, coliard greens, comfrey herb tea, coriander, corn, currants, dill, eggplant, endive, fennel, garlic, grapefruit, grapes, guava, honey, honeydew melon, horseradish, kale, lemon, lentils, lettuce, licorice, lime, mace, mango, marjoram, mint, mushrooms, mustard, nutmeg, onion, orange, paprika, parsley, parsnip, peach, pear, peas, pineapple, plum, potato, radish, raspberries, rhubarb, rosemary, rutabaga, sage, savory, sesame seeds, soybean, star anise, tarragon, tea, thyme, tomato, turmeric, and turnip.
by high-dose rodent tests, is ubiquitous (Ames et al., 1990b). It is probable that almost every fruit and vegetable in the supermarket contains natural pesticides that are rodent carcinogens. Even though only a tiny proportion of natural pesticides have been tested for carcinogenicity, 37 of 71 that have been tested are rodent carcinogens that are present in the common foods listed in Table 38.1. Humans also ingest numerous natural chemicals that are produced as by-products of cooking food. For example, more than 1000 chemicals have been identified in roasted coffee, many of which are produced by roasting (Clarke and Macrae, 1988; Nijssen et al., 1996). Only 30 have been tested for carcinogenicity according to the most recent results in our CPDB, and 21 of these are positive in at least one test (Table 38.2), totaling at least 10 mg of rodent carcinogens per cup of coffee (Clarke and Macrae, 1988; Fujita et aI., 1985; Kikugawa et aI., 1989; Nijssen et aI., 1996). Among the rodent carcinogens in coffee are the plant pesticides caffeic acid (present at 1800 ppm; C1arke and Macrae, 1988) and catechol (present at 100 ppm; Rahn and K6nig, 1978; Tressl et aI., 1978). Two other plant pesticides in coffee, chlorogenic acid and neochlorogenic acid (present at 21,600 and 11,600 ppm, respectively; Clarke and Macrae,
1988) are metabolized to caffeic acid and catechol but have not been tested for carcinogenicity. Chlorogenic acid and caffeic acid are mutagenic (Ariza et aI., 1988; Fung et aI., 1988; Hanham et aI., 1983) and clastogenic (Ishidate et aI., 1988; Stich et aI., 1981). Another plant pesticide in coffee, d-limonene, is carcinogenic but the only tumors induced were in male rat kidney, by a mechanism involving accumulation of c¥2u-globulin and increased cell division in the kidney, which would not be predictive of a carcinogenic hazard to humans (Dietrich and Swenberg, 1991; Rice et al., 1999). Some other rodent carcinogens in coffee are products of cooking, for example, furfural and benzo(a)pyrene. The point here is not to indicate that rodent data necessarily implicate coffee as a risk factor for human cancer, but rather to illustrate that there is an enormous background of chemicals in the diet that are natural and that have not been a focus of carcinogenicity testing. A diet free of naturally occurring chemicals that are carcinogens in high-dose rodent tests is impossible. It is often assumed that because natural chemicals are part of human evolutionary history, whereas synthetic chemicals are recent, the mechanisms that have evolved in animals to cope with the toxicity of natural chemicals will fail to protect against
Table 38.2 Carcinogenicity Status of Natural Chemicals in Roasted Coffee Positive: N = 21
Acetaldehyde, benzaldehyde, benzene, benzofuran, benzo(a)pyrene, caffeic acid, catechol, 1,2,5,6-dibenzanthracene, ethanol, ethylbenzene, formaldehyde, furan, furfural, hydrogen peroxide, hydroquinone, isoprene, limonene, 4-methyicatechol, styrene, toluene, xylene
Not positive:
Acrolein, biphenyl, chOline, eugenol, nicotinamide, nicotinic acid, phenol, piperidine
N=8 Uncertain:
Caffeine
Yet to test:
~ 1000
chemicals
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Pesticide Residues in Food and Cancer Risk: A Critical Analysis
synthetic chemicals, including synthetic pesticides (Ames et aI., 1987). This assumption is flawed for several reasons (Ames et aI., 1990b, 1996; Gold et aI., 1997a, b, c):
1. Humans have many natural defenses that buffer against normal exposures to toxins (Ames et aI., 1990b) and these are usually general, rather than tailored for each specific chemical. Thus, they work against both natural and synthetic chemicals. Examples of general defenses include the continuous shedding of cells exposed to toxins-the surface layers of the mouth, esophagus, stomach, intestine, colon, skin, and lungs are discarded every few days; deoxyribonucleic acid (DNA) repair enzymes, which repair DNA that was damaged from many different sources; and detoxification enzymes of the liver and other organs, which generally target classes of chemicals rather than individual chemicals. That human defenses are usually general, rather than specific for each chemical, makes good evolutionary sense. The reason that predators of plants evolved general defenses is presumably to be prepared to counter a diverse and ever-changing array of plant toxins in an evolving world; if a herbivore had defenses against only a specific set of toxins, it would be at great disadvantage in obtaining new food when favored foods became scarce or evolved new chemical defenses. 2. Various natural toxins, which have been present throughout vertebrate evolutionary history, nevertheless cause cancer in vertebrates (Ames et aI., 1990b; Gold et aI., 1997b, 1999; Vainio et aI., 1995). Mold toxins, such as aflatoxin, have been shown to cause cancer in rodents, monkeys, humans, and other species. Many of the common elements, despite their presence throughout evolution, are carcinogenic to humans at high doses (e.g., the salts of cadmium, beryllium, nickel, chromium, and arsenic). Furthermore, epidemiological studies from various parts of the world indicate that certain natural chemicals in food may be carcinogenic risks to humans; for example, the chewing of betel nut with tobacco is associated with oral cancer. Among the agents identified as human carcinogens by the International Agency for Research in Cancer (IARC) 62% (37/60) occur naturally: 16 are natural chemicals, 11 are mixtures of natural chemicals, and 10 are infectious agents (IARC, 1971-1999; Vainio et aI., 1995). Thus, the idea that a chemical is "safe" because it is natural, is not correct. 3. Humans have not had time to evolve a "toxic harmony" with all of their dietary plants. The human diet has changed markedly in the last few thousand years. Indeed, very few of the plants that humans eat today (e.g., coffee, cocoa, tea, potatoes, tomatoes, corn, avocados, mangos, olives and kiwi fruit) would have been present in a hunter-gatherer's diet. Natural selection works far too slowly for humans to have evolved specific resistance to the food toxins in these newly introduced plants. 4. Some early synthetic pesticides were lipophilic organochlorines that persist in nature and bioaccumulate in adipose tissue, for example, dichlorophenyltrichloroethane (DDT), aldrin, and dieldrin (DDT is discussed in
Section 38.5). This ability to bioaccumulate is often seen as a hazardous property of synthetic pesticides; however, such bioconcentration and persistence are properties of relatively few synthetic pesticides. Moreover, many thousands of chlorinated chemicals are produced in nature (Gribble, 1996). Natural pesticides also can bioconcentrate if they are fat soluble. Potatoes, for example, were introduced into the worldwide food supply a few hundred years ago; potatoes contain solanine and chaconine, which are fat-soluble, neurotoxic, natural pesticides that can be detected in the blood of all potato-eaters. High levels of these potato glycoalkaloids have been shown to cause reproductive abnormalities in rodents (Ames et aI., 1990b; Morris and Lee, 1984). 5. Because no plot of land is free from attack by insects, plants need chemical defenses-either natural or synthetic-to survive pest attack. Thus, there is a trade-off between naturally-occurring pesticides and synthetic pesticides. One consequence of efforts to reduce pesticide use is that some plant breeders develop plants to be more insect resistant by making them higher in natural pesticides. A recent case illustrates the potential hazards of this approach to pest control: When a major grower introduced a new variety of highly insect-resistant celery into commerce, people who handled the celery developed rashes when they were subsequently exposed to sunlight. Some detective work found that the pest-resistant celery contained 6200 parts per billion (ppb) of carcinogenic (and mutagenic) psoralens instead ofthe 800 ppb present in common celery (Beier and Nigg, 1994; Berkley et aI., 1986; Seligman et aI., 1987).
38.3 THE HIGH CARCINOGENICITY RATE AMONG CHEMICALS TESTED IN CHRONIC ANIMAL CANCER TESTS Because the toxicology of natural and synthetic chemicals is similar, one expects, and finds, a similar positivity rate for carcinogenicity among synthetic and natural chemicals that have been tested in rodent bioassays. Among chemicals tested in rats and mice in the CPDB, about half the natural chemicals are positive, and about half of all chemicals tested are positive. This high positivity rate in rodent carcinogenesis bioassays is consistent for many data sets (Table 38.3): Among chemicals tested in rats and mice, 59% (350/590) are positive in at least one experiment, 60% of synthetic chemicals (2711451), and 57% of naturally occurring chemicals (79/139). Among chemicals tested in at least one species, 52% of natural pesticides (37171) are positive, 61 % of fungal toxins (14123), and 70% of the naturally occurring chemicals in roasted coffee (21130) (Table 38.2). Among commercial pesticides reviewed by the EPA (U.S. Environmental Protection Agency, 1998), the positivity rate is 41 % (79/193); this rate is similar among commercial pesticides that also occur naturally and those that are only synthetic, as well as between commercial pesticides that have been canceled and those still in use. (See Section 38.8 for detailed summary results
38.3 The High Carcinogenicity Rate Among Chemicals Tested in Chronic Animal Cancer Tests Table 38.3 Proportion of Chemicals Evaluated as Carcinogenic Chemicals tested in both rats and micea Chemicals in the CPDB Naturally occurring chemicals in the CPDB Synthetic chemicals in the CPDB
350/590 (59%)
79/139 (57%) 2711451 (60%)
Chemicals tested in rats and/or micea Chemicals in the CPDB Natural pesticides in the CPDB
70211348 (52%) 37171 (52%)
Mold toxins in the CPDB
14123 (61%)
Chemicals in roasted coffee in the CPDB
2l/30 (70%)
Commercial pesticides in the CPDB
79/193 (41 %)
Physicians' Desk Reference (PDR): Drugs with reported cancer tests b FDA database of drug submissionsc
117/241 (49%) 125/282 (44%)
aFrom the Carcinogenic Potency Database (Gold et aI., 1997c, 1999). bDavies and Monro (1995). CContrera et al. (1997). 140 drugs are in both the FDA and the PDR databases.
of carcinogenicity tests on the 193 commercial pesticides in the CPDB, including results on the positivity of each chemical, its carcinogenic potency, and target organs of carcinogenesis.) Because the results of high-dose rodent tests are routinely used to identify a chemical as a possible cancer hazard to humans, it is important to try to understand how representative the 50% positivity rate might be of all untested chemicals. If half of all chemicals (both natural and synthetic) to which humans are exposed were positive if tested, then the utility of a test to identify a chemical as a "potential human carcinogen" because it increases tumor incidence in a rodent bioassay would be questionable. To determine the true proportion of rodent carcinogens among chemicals would require a comparison of a random group of synthetic chemicals to a random group of natural chemicals. Such an analysis has not been done. It has been argued that the high positivity rate is due to selecting more suspicious chemicals to test for carcinogenicity. For example, chemicals may be selected that are structurally similar to known carcinogens or genotoxins. That is a likely bias because cancer testing is both expensive and time consuming, making it prudent to test suspicious compounds. On the other hand, chemicals are selected for testing for many reasons, including the extent of human exposure, level of production, and scientific questions about carcinogenesis. Among chemicals tested in both rats and mice, chemicals that are mutagenic in Salmonella are carcinogenic in rodent bioassays more frequently than nonmutagens: 80% of mutagens are positive (1761219) compared to 50% (1351271) of nonmutagens. Thus, if testing is based on suspicion of carcinogenicity, then more mutagens should be selected than nonmutagens; however, of the chemicals tested in both species, 55% (271/490) are not mutagenic. This suggests that prediction of positivity is often not the basis for selecting a chemical to test. Another argument against selection bias is the high positivity rate for drugs (Ta-
803
ble 38.3), because drug development tends to favor chemicals that are not mutagens or suspected carcinogens. In the Physicians' Desk Reference (PDR), however, 49% (1171241) of the drugs that report results of animal cancer tests are carcinogenic (Davies and Monro, 1995) (Table 38.3). Moreover, while some chemical classes are more often carcinogenic in rodent bioassays than others (e.g., nitroso compounds, aromatic amines, nitroaromatics, and chlorinated compounds), prediction is still imperfect. For example, a prospective prediction exercise was conducted by several experts in 1990 in advance of the 2-year National Toxicology Program bioassays. There was wide disagreement among the experts on which chemicals would be carcinogenic when tested, and the level of accuracy varied by expert, thus indicating that predictive knowledge is uncertain (Omenn et aI., 1995). One large series of mouse experiments by Innes et al. (1969) has frequently been cited (U.S. National Cancer Institute, 1984) as evidence that the true proportion of rodent carcinogens is actually low among tested substances (Table 38.4). In the Innes study, 119 synthetic pesticides and industrial chemicals were tested, and only 11 (9%) were evaluated as carcinogenic. Our analysis indicates that those early experiments lacked power to detect an effect because they were conducted only in mice (not in rats), they included only 18 animals in a group (compared with the standard protocol of 50), the animals were tested for only 18 months (compared with the standard 24 months), and the Innes dose was usually lower than the highest dose in subsequent mouse tests if the same chemical was tested again (Gold and Zeiger, 1997; Gold et aI., 1999; Innes et al., 1969). To assess whether the low positivity rate in the Innes study was due to the lack of power in the design of the experiments, we used results in our CPDB to examine subsequent bioassays on the Innes chemicals that had not been evaluated as positive (results and chemical names are reported in Table 38.4). Among the 34 chemicals that were not positive in the Innes study and were subsequently retested with more standard protocols, 17 had a subsequent positive evaluation of carcinogenicity (50%), which is similar to the proportion among all chemicals in the CPDB (Table 38.4). Of the 17 new positives, 7 were carcinogenic in mice and 14 in rats. Innes et al. had recommended further evaluation of some chemicals that had inconclusive results in their study. If those were the chemicals subsequently retested, then one might argue that they would be the most likely to be positive. Our analysis does not support that view, however. We found that the positivity rate among the chemicals that the Innes study said needed further evaluation was 7 of 16 (44%) when retested, compared to 10 of 18 (56%) among the chemicals that Innes evaluated as negative. Our analysis thus supports the idea that the low positivity rate in the Innes study resulted from lack of power. Because many of the chemicals tested by Innes et al. were synthetic pesticides, we reexamined the question of what proportion of synthetic pesticides are carcinogenic (as shown in Table 38.3) by excluding the pesticides tested only in the Innes series. The Innes studies had little effect on the positivity rate: Table 38.3 indicates that of all commercial pesticides in the
804
CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
Table 38.4 Results of Subsequent Tests on Chemicals (Primarily Pesticides) not Found Carcinogenic by rnnes et al. (1969) Percentage carcinogenic when retested Retested chemicals All retested
Mice
Rats
Either mice or rats
7/26 (27%)
14/34 (41 %)
17134 (50%)
rnnes: not carcinogenic
3/10 (30%)
9/18 (50%)
10/18 (56%)
rnnes: needs further evaluation
4/16 (25%)
5/16 (31 %)
7116 (44%)
Of 119 chemicals tested by rnnes et al., 11 (9%) were evaluated as positive by rnnes et al. Carcinogenic when retested: atrazine (R), azobenzene* (R), captan (M, R), carbaryl (R), 3-(p-chlorophenyl)-I,I-dimethylurea* (R), p,p'-DDD* (M), folpet (M), manganese ethylenebisthiocarbamate (R), 2-mercaptobenzothiazole (R), N -nitrosodiphenylamine* (R), 2,3,4,5,6-pentachlorophenol (M, R), o-phenylphenol (R), piperonyl butoxide* (M, R), piperonyl sulfoxide* (M),2,4,6-trichlorophenol* (M, R), zinc dimethyldithiocarbamate (R), zinc ethylenebisthiocarbamate (R). Not carcinogenic when retested: (2-chloroethyl)trimethylammonium chloride*, calcium cyanamide*, diphenyl-p-phenylenediamine, endosulfan, p, p'ethyl-DDD*, ethyl tellurac*, isopropyl-N -(3-chlorophenyl) carbamate, lead dimethyldithiocarbamate*, maleic hydrazide, mexacarbate*, monochloroacetic acid, phenyl-,B-naphthylamine*, rotenone, sodium diethyldithiocarbamate trihydrate*, tetraethylthiuram disulfide*, tetramethy Ithiuram disulfide, 2,4,5trichlorophenoxyacetic acid. (M), positive in mice when retested; (R), positive in rats when retested; *, rnnes et al. stated that further testing was needed.
CPDB, 41 % 791193 are rodent carcinogens; when the analysis is repeated by excluding those Innes tests, 47% (77/165) are carcinogens.
38.4 THE IMPORTANCE OF CELL DIVISION IN MUTAGENESIS AND CARCINOGENESIS What might explain the high proportion of chemicals that are carcinogenic when tested in rodent cancer bioassays (Table 38.3)? In standard cancer tests, rodents are given a chronic, near-toxic dose: the maximum tolerated dose (MTD). Evidence is accumulating that cell division caused by the high dose itself, rather than the chemical per se, contributes to cancer in such tests (Ames and Gold, 1990; Ames et aI., 1993a; Butterworth and Bogdanffy, 1999; Cohen, 1998; Cunningham, 1996; Cunningham and Matthews, 1991; Cunningham et aI., 1991; Heddle, 1998). High doses can cause chronic wounding of tissues, cell death, and consequent chronic cell division of neighboring cells, which is a risk factor for cancer (Ames and Gold, 1990; Gold et aI., 1998). Each time a cell divides, there is some probability that a mutation will occur, and thus increased cell division increases the risk of cancer. At the low levels of pesticide residues to which humans are usually exposed, such increased cell division does not occur. The process of mutagenesis and carcinogenesis is complicated because many factors are involved, for example, DNA lesions, DNA repair, cell division, clonal instability, apoptosis, and p53 (a cell cycle gene that is mutated in half of human tumors) (Christensen et aI., 1999; Hill et aI., 1999). The normal endogenous level of oxidative DNA lesions in somatic cells is appreciable (Helbock et aI., 1998). In addition, tissues injured by high doses of chemicals have an inflammatory immune response involving activation of white cells in response to cell death (Adachi et aI., 1995; Czaja et aI., 1994; Gunawardhana et aI., 1993; Laskin and Pendino, 1995; Roberts and Kimber, 1999). Activated white cells release mutagenic oxidants (including peroxynitrite, hypochlorite, and
H202). Therefore, the very low levels of synthetic pesticide residues to which humans are exposed may pose no or only minimal cancer risks. It seems likely that a high proportion of all chemicals, whether synthetic or natural, might be "carcinogens" if administered in the standard rodent bioassay at the MTD, primarily due to the effects of high doses on cell division and DNA damage (Ames and Gold, 1990; Ames et aI., 1993a; Butterworth et aI., 1995; Cunning ham, 1996; Cunningham and Matthews, 1991; Cunningham et aI., 1991). For nonmutagens, cell division at the MTD can increase carcinogenicity; for mutagens, there can be a synergistic effect between DNA damage and cell division at high doses. Ad libitum feeding in the standard bioassay can also contribute to the high positivity rate (Hart et aI., 1995). In calorie-restricted mice, cell division rates are markedly lower in several tissues than in ad libitum-fed mice (Lok et aI., 1990). In dosed animals, food restriction decreased tumor incidence at all three sites that were evaluated as target sites (pancreas and bladder in male rats, liver in male mice), and none of those sites was evaluated as target sites after 2 or 3 years (U.S. National Toxicology Program, 1997). In standard National Cancer Institute (NCI)INational Toxicology Program (NTP) bioassays, for both control and dosed animals, food restriction improves survival and at the same time decreases tumor incidence at many sites compared to ad libitum-feeding. Without additional data on how a chemical causes cancer, the interpretation of a positive result in a rodent bioassay is highly uncertain. Although cell division is not measured in routine cancer tests, many studies on rodent carcinogenicity show a correlation between cell division at the MTD and cancer (Cunningham et aI., 1995; Gold et aI., 1998; Hayward et aI., 1995). Extensive reviews of bioassay results document that chronic cell division can induce cancer (Ames and Gold, 1990; Ames et aI., 1993b; Cohen, 1995b; Cohen and Ellwein, 1991; Cohen and Lawson, 1995; Counts and Goodman, 1995; Gold et aI., 1997b). A large epidemiological literature reviewed by PrestonMartin et al. (1990, 1995) indicates that increased cell division by hormones and other agents can increase human cancer.
38.4 The Importance of Cell Division in Mutagenesis and Carcinogenesis
Several of our findings in large-scale analyses of the results of animal cancer tests (Gold et aI., 1993) are consistent with the idea that cell division increases the carcinogenic effect in high-dose bioassays, including the high proportion of chemicals that are positive; the high proportion of rodent carcinogens that are not mutagenic; and the fact that mutagens, which can both damage DNA and increase cell division at high doses, are more likely than nonmutagens to be positive, to induce tumors in both rats and mice, and to induce tumors at multiple sites (Gold et aI., 1993, 1998). Analyses of the limited data on dose response in bioassays are consistent with the idea that cell division from cell killing and cell replacement is important. Among rodent bioassays with two doses and a control group, about half the sites evaluated as target sites are statistically significant at the MTD but not at half the MTD (p < 0.05). The proportions are similar for mutagens (44%, 148/334) and nonmutagens (47%, 76/163) (Gold and Zeiger, 1997; Gold et aI., 1999), suggesting that cell division at the MTD may be important for the carcinogenic response of mutagens as well as nonmutagens that are rodent carcinogens. To the extent that increases in tumor incidence in rodent studies are due to the secondary effects of inducing cell division at the MTD, then any chemical is a likely rodent carcinogen, and carcinogenic effects can be limited to high doses. Linearity of the dose-response relationship also seems less likely than has been assumed because of the inducibility of numerous defense enzymes that deal with exogenous chemicals as groups (e.g., oxidants, electrophiles) and thus protect humans against natural and synthetic chemicals, including potentially mutagenic reactive chemicals (Ames et aI., 1990b; Luckey, 1999; Munday and Munday, 1999; Trosko, 1998). Thus, true risks at the low doses of most exposures to the general population are likely to be much lower than what would be predicted by the linear model that has been the default in U.S. regulatory risk assessment. The true risk might often be O. Agencies that evaluate potential cancer risks to humans are moving to take mechanism and nonlinearity into account. The U.S. Environmental Protection Agency (EPA) recently proposed new cancer risk assessment guidelines (U.S. Environmental Protection Agency, 1996a) that emphasize a more flexible approach to risk assessment and call for the use of more biological information in the weight-of-evidence evaluation of carcinogenicity for a given chemical and in the dose-response assessment. The proposed changes take into account the issues that were discussed previously. The new EPA guidelines recognize the dose dependence of many toxicokinetic and metabolic processes and the importance of understanding cancer mechanisms for a chemical. The guidelines use nonlinear approaches to low-dose extrapolation if warranted by mechanistic data and a possible threshold of dose below which effects will not occur (National Research Council, 1994; U.S. Environmental Protection Agency, 1996a). In addition, toxicological results for cancer and noncancer endpoints could be incorporated together in the risk assessment process. Also consistent with the results discussed previously, are the recent IARC consensus criteria for evaluations of carcino-
805
genicity in rodent studies, which take into account that an agent can cause cancer in laboratory animals through a mechanism that does not operate in humans (Rice et aI., 1999). The tumors in such cases involve persistent hyperplasia in cell types from which the tumors arise. These include urinary bladder carcinomas associated with certain urinary precipitates, thyroid follicular-cell tumors associated with altered thyroidstimulating hormone (TSH), and cortical tumors of the kidney that arise only in male rats in association with nephropathy that is due to CV2u urinary globulin. Historically, in U.S. regulatory policy, the "virtually safe dose," corresponding to a maximum, hypothetical risk of one cancer in a million, has routinely been estimated from results of carcinogenesis bioassays using a linear model, which assumes that there are no unique effects of high doses. To the extent that carcinogenicity in rodent bioassays is due to the effects of high doses for the nonmutagens, and a synergistic effect of cell division at high doses with DNA damage for the mutagens, this model overestimates risk (Butterworth and Bogdanffy, 1999; Gaylor and Gold, 1998). We have discussed validity problems associated with the use of the limited data from animal cancer tests for human risk assessment (Bemstein et aI., 1985; Gold et aI., 1998). Standard practice in regulatory risk assessment for a given rodent carcinogen has been to extrapolate from the high doses of rodent bioassays to the low doses of most human exposures by multiplying carcinogenic potency in rodents by human exposure. Strikingly, however, due to the relatively narrow range of doses in 2-year rodent bioassays and the limited range of statistically significant tumor incidence rates, the various measures of potency obtained from 2-year bioassays, such as the EPA value, the TDso, and the lower confidence limit on the TDlO (LTDlO), are constrained to a relatively narrow range of values about the MTD, in the absence of 100% tumor incidence at the target site, which rarely occurs (Bemstein et aI., 1985; Freedman et aI., 1993; Gaylor and Gold, 1995, 1998; Gold et aI., 1997b). For example, the dose usually estimated by regulatory agencies to give one cancer in a million can be approximated simply by using the MTD as a surrogate for carcinogenic potency. The "virtually safe dose" (VSD) can be approximated from the MTD. Gaylor and Gold (1995) used the ratio MTDITDso and and TDso found by Krewski et al. the relationship between (1993) to estimate the VSD. The VSD was approximated by the MTD1740,000 for rodent carcinogens tested in the bioassay program of the NCIINTP. The MTD1740,000 was within a factor of 10 of the VSD for 96% of carcinogens. This is similar to the finding that in near-replicate experiments of the same chemical, potency estimates vary by a factor of 4 around a median value (Gold et aI., 1987a; Gold et aI., 1989; Gaylor et aI., 1993). Using the benchmark dose approach proposed in the EPA carcinogen guidelines, risk estimation is similarly constrained by bioassay design. A simple, quick, and relatively precise determination of the LTDlO can be obtained by the MTD divided by 7 (Gaylor and Gold, 1998). Both linear extrapolation and the use of safety or uncertainty factors proportionately reduce
q;
q;
806
CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
a tumor dose in a similar manner. The difference in the regulatory "safe dose," if any, for the two approaches depends on the magnitude of uncertainty factors selected. Using the benchmark dose approach of the proposed carcinogen risk assessment guidelines, the dose estimated from the LTDlO divided, for example, by a WOO-fold uncertainty factor, is similar to the dose of an estimated risk of less than 10-4 using a linear model. This dose is 100 times higher than the VSD corresponding to an estimated risk of less than 10-6 . Thus, whether the procedure involves a benchmark dose or a linearized model, cancer risk estimation is constrained by the bioassay design.
38.5 THE HERP RANKING OF POSSIBLE CARCINOGENIC HAZARDS Given the lack of epidemiological data to link pesticide residues to human cancer, as well as the limitations of cancer bioassays for estimating risks to humans at low exposure levels, the high positivity rate in bioassays, and the ubiquitous human exposures to naturally occurring chemicals in the normal diet that are rodent carcinogens (Tables 38.1-38.3), how can bioassay data best be used if our goal is to evaluate potential carcinogenic hazards to humans from pesticide residues in the diet? In several papers, we have emphasized the importance of setting research and regulatory priorities by gaining a broad perspective about the vast number of chemicals to which humans are exposed. A comparison of potential hazards can be helpful in efforts to communicate to the public what might be important factors in cancer prevention and when selecting chemicals for chronic bioassay, mechanistic, or epidemiologic studies (Ames et aI., 1987, 1990b; Gold and Zeiger, 1997; Gold et aI., 1992). There is a need to identify what might be the important cancer hazards among the ubiquitous exposures to rodent carcinogens in everyday life. One reasonable strategy for setting priorities is to use a rough index to compare and rank possible carcinogenic hazards from a wide variety of chemical exposures to rodent carcinogens at levels that humans receive, and then to focus on those that rank highest in possible hazard (Ames et aI., 1987; Gold et aI., 1992, 1994a). Ranking is thus a critical first step. Although one cannot say whether the ranked chemical exposures are likely to be of major or minor importance in human cancer, it is not prudent to focus attention on the possible hazards at the bottom of a ranking if, using the same methodology to identify a hazard, there are numerous common human exposures with much greater possible hazards. Our analyses are based on the HERP (human exposure/rodent potency) index, which indicates what percentage of the rodent carcinogenic dose (TDso in mg/kg/day) a human receives from a given average daily exposure for a lifetime (mg/kg/day). TDso values in our CPDB span a 10 million-fold range across chemicals (Gold et aI., 1997c). Human exposures to rodent carcinogens range enormously as well, from historically high workplace exposures in some occupations or pharmaceutical dosages to very low exposures from residues of synthetic chemicals in food or water.
The rank order of possible hazards for the given exposure estimates will be similar for the HERP ranking, for a ranking of regulatory "risk estimates" based on a linear model, or for a ranking based on TDlO, since all 3 methods are proportional to the dose. Overall, our analyses have shown that synthetic pesticide residues rank low in possible carcinogenic hazards compared to many common exposures. HERP values for some historically high exposures in the workplace and some pharmaceuticals rank high, and there is an enormous background of naturally occurring rodent carcinogens in typical portions or average consumption of common foods. This result casts doubt on the relative importance of low-dose exposures to residues of synthetic chemicals such as pesticides (Ames et aI., 1987; Gold et aI., 1992, 1994a). A committee of the National Research Council recently reached similar conclusions about natural versus synthetic chemicals in the diet and called for further research on natural chemicals (National Research Council, 1996). (See Section 38.7 for further work on natural chemicals. ) The HERP ranking in Table 38.5 is for average U.S. exposures to all rodent carcinogens in the CPDB for which concentration data and average exposure or consumption data were both available, and for which known exposure could be chronic for a lifetime. For pharmaceuticals the doses are recommended doses; for the workplace, they are past industry or occupation averages. The 87 exposures in the ranking (Table 38.5) are ordered by possible carcinogenic hazard (HERP), and natural chemicals in the diet are reported in boldface. Our early HERP rankings were for typical dietary exposures (Ames et aI., 1987; Gold et aI., 1992), and results are similar. Several HERP values make convenient reference points for interpreting Table 38.5. The median HERP value is 0.0025%, and the background HERP for the average chloroform level in a liter of U.S. tap water is 0.0003%. A HERP of 0.00001 % is approximately equal to a regulatory VSD risk of 10-6 based on the linearized multi-stage model (Gold et aI., 1992). Using the benchmark dose approach recommended in the new EPA guidelines with the LTDlO as the point of departure (POD), linear extrapolation would produce a similar estimate of risk at 10-6 and hence a similar HERP value (Gaylor and Gold, 1998), if information on the carcinogenic mode of action for a chemical supports a nonlinear dose-response curve. The EPA guidelines call for a margin-of-exposure approach with the LTDlO as the POD. Based on that approach, the reference dose using a safety or uncertainty factor of 1000 (i.e., LDlO/1000) would be equivalent to a HERP value of 0.001 %, which is similar to a risk of 10-4 based on a linear model. If the dose-response relationship is judged to be nonlinear, then the cancer risk estimate will depend on the number and magnitude of safety factors used in the assessment. The HERP ranking maximizes possible hazards to synthetic chemicals because it includes historically high exposure values that are now much lower [e.g., DDT, saccharin, butylated hydroxyanisole (BHA), and some occupational exposures]. Additionally, the values for dietary pesticide residues are averages in the total diet, whereas for most natural chemicals the ex-
38.5 The HERP Ranking of Possible Carcinogenic Hazards
807
Table 38.5 Ranking Possible Carcinogenic Hazards from Average V.S. Exposures to Rodent Carcinogens Possible hazard:
Potency TDso
HERP
Human dose of
(mg/kg/day)a
(%)
Average daily V.S. exposure
rodent carcinogen
Rats
Mice
Exposure references
140
EDB: production workers (high
Ethylene dibromide, 150 mg
1.52
(7.45)
Ott et al. (1980), Ramsey et al. (1978)
exposure) (before 1977) 17
Clofibrate
Clofibrate, 2 g
169
14
Phenobarbital, 1 sleeping pill
Phenobarbital, 60 mg
(+)
6.09
Havel and Kane (1982) AMA(l983)
6.8
1,3-Butadiene: rubber industry workers
1,3-Butadiene, 66.0 mg
(261)
13.9
Matanoski et al. (1993)
6.2
Comfrey-pepsin tablets, 9 daily
Comfrey root, 2.7 g
626
Tetrachloroethylene, 433 mg
101
(126)
Andrasik and Cloutet (1990)
Formaldehyde, 6.1 mg
2.19
(43.9)
Siegal et al. (1983)
Acrylonitrile, 405 !-lg
16.9
Trichloroethylene, 1.02 g
668
(1580)
(1978-1986) Hirono et al. (1978), Culvenor et al. (1980)
(no longer recommended) 6.1
Tetrachloroethylene: dry cleaners with dry-to-dry units (1980-1990)
4.0
Formaldehyde: production workers (1979)
2.4
Acrylonitrile: production workers
Blair et al. (1998)
(1960-1986) 2.2
Trichloroethylene: vapor degreasing
Page and Arthur (1978)
(before 1977) 2.1
Beer, 257 g
Ethyl alcohol, 13.1 ml
9110
(-)
Stofberg and Grundschober (1987)
1.4
Mobile home air (14 h/day)
Formaldehyde, 2.2 mg
2.19
(43.9)
Connor et al. (1985)
1.3
Comfrey-pepsin tablets, 9 daily
Symphytine, 1.8 mg
1.91
Methylene chloride, 471 mg
724
(1100)
CONSAD (1990)
(-)
Stofberg and Grundschober (1987) McCann et al. (1987)
Hirono et al. (1978), Culvenor et al. (1980)
(no longer recommended) 0.9
Methylene chloride: workers, industry average (l940s-1980s)
0.5
Wine,28.0g
Ethyl alcohol, 3.36 ml
9110
0.5
Dehydroepiandrosterone (DHEA)
DHEA supplement, 25 mg
68.1
0.4
Conventional home air (14 h/day)
Formaldehyde, 598 !-lg
2.19
(43.9)
0.2
Omeprazole
Omeprazole, 20 mg
199
(-)
0.2
Fluvastatin
Fluvastatin, 20 mg
125
0.1
Coffee, 13.3 g
Caffeic acid, 23.9 mg
297
(4900)
Stofberg and Grundschober (1987),
0.1
d-Limonene in food
d-Limonene, 15.5 mg
204
(-)
Stofberg and Grundschober (1987)
0.04
Bread, 67.6 g
Ethyl Alcohol 243 mg
9110
(-)
Stofberg and Grundschober (1987),
0.04
Lettuce, 14.9 g
Caffeic acid, 7.90 mg
297
(4900)
TAS (1989), Herrmann (1978)
PDR (1998) PDR (1998) Clarke and Macrae (1988)
Wolm et al. (1974) 0.03
Safrole in spices
Safrole, 1.2 mg
(441)
51.3
Hall et al. (1989)
0.03
Orange juice, 138 g
d-Limonene, 4.28 mg
204
(-)
TAS (1989), Schreier et al. (1979)
0.03
Comfrey herb tea, 1 cup (1.5 g root)
Symphytine, 38 J.l g
1.91
Culvenor et al. (1980)
(no longer recommended) 0.03
Tomato, 88.7 g
Caffeic acid, 5.46 mg
297
(4900)
TAS (1989), Schmidtlein and Herrmann (l975a)
0.03
Pepper, black, 446 mg
d-Limonene, 3.57 mg
204
(-)
Stofberg and Grundschober (1987),
0.02
Coffee, 13.3 g
Catechol, 1.33 mg
88.8
(244)
Stofberg and Grundschober (1987),
0.02
Furfural in food
Furfural, 2.72 mg
(683)
197
Stofberg and Grundschober (1987)
0.02
Mushroom (Agaricus bisporus) 2.55 g
Mixture of hydrazines, etc.
20,300
Stofberg and Grundschober (1987),
Hasselstrom et al. (1957) Tressl et al. (1978), Rahn and Konig (1978)
(whole mushroom)
Toth and Erickson (1986), Matsumoto et al. (1991) (continues)
808
CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
Table 38.5 (continued) Possible hazard:
Potency TDSO
HERP
Human dose of
(%)
Average daily V.S. exposure
0.02
Apple, 32.0 g
0.02
Coffee, 13.3 g
0.01
BHA: daily U.S. avg (1975)
0.01
Beer (before 1979), 257 g
rodent carcinogen
(mg/kg/day)a Rats
Mice
Exposure references
Caffeic acid, 3.40 mg
297
(4900)
EPA (1989a), Mosel and Herrmann (1974)
Furfural, 2.09 mg
(683)
197
Stofberg and Grundschober (1987)
BHA,4.6mg
606
(5530)
FDA (1991b)
Dimethylnitrosamine, 726 ng
0.0959
(0.189)
Stofberg and Grundschober (1987), Fazio et al. (1980), Preussmann and Eisenbrand (1984)
0.008
Aflatoxin: daily U.S. avg (1984-1989)
Aflatoxin, 18 ng
0.007
Cinnamon, 21.9 mg
Coumarin, 65.0 I!g
13.9
(103)
Poole and Poole (1994)
0.006
Coffee, 13.3 g
Hydroquinone, 333 I!g
82.8
(225)
Stofberg and Grundschober (1987),
0.0032
(+)
FDA (1992b)
Tressl et al. (1978), Heinrich and Baltes (1987) 0.005
Saccharin: daily V.S. avg (1977)
Saccharin, 7 mg
2140
(-)
0.005
Carrot, 12.1 g
Aniline, 624 I!g
194b
(-)
TAS (1989), Neurath et al. (1977)
0.004
Potato, 54.9 g
Caffeic acid, 867 I!g
297
(4900)
TAS (1989), Schmidtlein and Herrmann
0.004
Celery, 7.95 g
Caffeic acid, 858 I!g
297
(4900)
ERS (1994), Stohr and Herrmann (1975)
NRC (1979)
(1975c) 0.004
White bread, 67.6 g
Furfural, 500 I!g
(683)
197
Stofberg and Grundschober (1987)
0.003
d-Limonene
Food additive, 475 !J.g
204
(-)
Clydesdale (1997)
0.003
Nutmeg, 27.4 mg
d-Limonene,466 !J.g
204
(-)
Stofberg and Grundschober (1987),
0.003
Conventional home air (14 h/day)
Benzene, 155 !J.g
(169)
77.5
McCann et al. (1987)
0.002
Coffee, 13.3 g
4-Methylcatechol, 433 I!g
248
Bejnarowicz and Kirch (1963) Stofberg and Grundschober (1987), Heinrich and Baltes (1987), IARC (1991) 0.002
Carrot, 12.1 g
Caffeic acid, 374 I!g
297
(4900)
TAS (1989), Stohr and Herrmann (1975)
0.002
Ethylene thiourea: daily V.S. avg (1990)
Ethylene thiourea, 9.51 !J.g
7.9
(23.5)
EPA (1991a)
0.002
BHA: daily V.S. avg (1987)
BHA,700 !J.g
606
(5530)
FDA (1991b)
0.002 0.001
DDT: daily V.S. avg (before 1972 ban)d
DDT, 13.8!J.g
(84.7)
Plum,2.00g
Caffeic acid, 276 I!g
297
12.8 (4900)
ERS (1995), Mosel and Herrmann (1974)
0.001
Pear, 3.29 g
Caffeic acid, 240 I!g
297
(4900)
Stofberg and Grundschober (1987),
0.001
[VDMH: daily V.S. avg (1988)]
[VDMH, 2.82 !J.g (from Alar)]
(-)
3.96
EPA (1989a)
0.0009
Brown mustard, 68.4 mg
Allyl isothiocyanate, 62.9 I!g
96
(-)
Stofberg and Grundschober (1987),
0.0008
DDE: daily V.S. avg (before 1972 ban)d
DDE,6.91 !J.g
(-)
12.5
Duggan and Comeliussen (1972)
0.0007
TCDD: daily V.S. avg (1994)
TCDD, 12.0 pg
0.0000235
(0.000156)
EPA (1994b)
0.0006
Bacon, 11.5 g
Diethylnitrosamine, 11.5 ng
0.0266
(+)
Stofberg and Grundschober (1987),
0.0006
Mushroom (Agaricus bisporus) 2.55 g
Glutamyl-p-hydrazinobenzoate,
277
Stofberg and Grundschober (1987),
0.0005
Bacon, 11.5 g
Dimethylnitrosamine, 34.5 ng
0.0959
(0.189)
Stofberg and Grundschober (1987),
0.0004
Bacon, 11.5 g
N -Nitrosopyrrolidine, 196 ng
(0.799)
0.679
Stofberg and Grundschober (1987),
0.0004
EDB: daily V.S. avg (before 1984 ban)d
EDB,420ng
1.52
(7.45)
EPA (1984b)
0.0004
Tap water, 11iter (1987-1992)
Bromodichloromethane, 13 !J.g
(72.5)
47.7
AWWA(1993)
0.0003
Mango, 1.22 g
d-Limonene, 48.8 I!g
204
(-)
ERS (1995), Engel and Tress1 (1983)
Duggan and Come1iussen (1972)
Mosel and Herrmann (1974)
Car1son et al. (1987)
Sen et al. (1979) Chauhan et al. (1985)
107 I!g
Sen et al. (1979) Tricker and Preussmann (1991)
(continues)
38.5 The HERP Ranking of Possible Carcinogenic Hazards
809
Table 38.5 (continued) Possible hazard: HERP
Potency TDso (mg/kg/day)Q
Human dose of Average daily V.S. exposure
rodent carcinogen
Rats
0.0003
Beer, 257 g
Furfural, 39.9 IJg
(683)
197
Stofberg and Grundschober (1987)
0.0003
Tap water, 1 liter (1987-1992)
Chloroform, 17 J,lg
(262)
90.3
AWWA(1993) Gloria et at. (1997)
(%)
Mice
Exposure references
0.0003
Beer (1994-1995), 257 g
Dimethylnitrosamine, 18 ng
0.0959
(0.189)
0.0003
Carbaryl: daily V.S. avg (1990)
Carbaryl, 2.6 J,lg
14.1
(-)
FDA (1991a)
0.0002
Celery, 7.95 g
8-Methoxypsoralen,4.86IJg
32.4
(-)
ERS (1994), Beier et at. (1983)
(-)
0.0002
Toxaphene: daily V.S. avg (1990)d
Toxaphene, 595 ng
0.00009
Mushroom (Agaricus bisporus),
p-Hydrazinobenzoate, 28 IJg
0.00008
PCBs: daily V.S. avg (1984-1986)
5.57
FDA (l991a)
454b
Stofberg and Grundschober (1987),
(9.58)
Gunderson (1995) FDA (199Ia)
Chauhan et at. (1985)
2.55 g PCBs, 98 ng
1.74
0.00008
DDEIDDT: daily V.S. avg (l990)d
DDE,659ng
(-)
12.5
0.00007
Parsnip, 54.0 mg
8-Methoxypsoralen, 1.57 IJg
32.4
(-)
VFFVA (1989), Ivie et at. (1981)
0.00007
Toast, 67.6 g
Vrethane, 811 ng
(41.3)
16.9
Stofberg and Grundschober (1987),
0.00006
Hamburger, pan fried, 85 g
PhIP, 176 ng
4.22b
(28.6 b )
TAS (1989), Knize et at. (1994)
0.00006
Furfural
Food additive, 7.77 J,lg
(683)
0.00005
Estragole in spices
Estragole, 1.99 IJg
0.00005
Parsley, fresh, 324 mg
8-Methoxypsoralen, 1.17 IJg
0.00005
Estragole
Food additive, 1.78 J,lg
0.00003
Hamburger, pan fried, 85 g
MeIQx, 38.1 ng
0.00002
Dicofol: daily V.S. avg (1990)
Dicofol, 544 ng
0.00001
Beer, 257 g
U rethane, 115 ng
0.000006
Hamburger, pan fried, 85 g
0.000005
Hexachlorobenzene: daily V.S. avg
Canas et at. (1989) 197
Clydesdale (1997)
51.8
Stofberg and Grundschober (1987)
(-)
VFFVA (1989), Chaudhary et at. (1986)
51.8
Clydesdale (1997)
1.66
(24.3)
TAS (1989), Knize et at. (1994)
(-)
32.9
FDA (l99Ia)
(41.3)
16.9
Stofberg and Grundschober (1987),
IQ,6.38ng
1.65 b
(19.6)
TAS (1989), Knize et at. (1994)
Hexachlorobenzene, 14 ng
3.86
(65.1)
FDA (l991a) FDA (l991a)
32.4
Canas et al. (1989)
(1990) 0.000001
Lindane: daily V.S. avg (1990)
Lindane, 32 ng
(-)
30.7
0.0000004
PCNB: daily V.S. avg (1990)
PCNB (Quintozene), 19.2 ng
(-)
71.1
FDA (l991a)
0.0000001
Chlorobenzilate: daily V.S. avg (l989)d
Chlorobenzilate, 6.4 ng
(-)
93.9
FDA (l99Ia)
0.00000008
Captan: daily V.S. avg (1990)
Captan, 115 ng
2080
(2110)
FDA(l991a)
0.00000001
Folpet: daily V.S. avg (1990)
Folpet, 12.8 ng
(-)
1550
FDA (l991a)
<0.00000001
Chlorothalonil: daily V.S. avg (1990)
Chlorothalonil, <6.4 ng
828 C
(-)
FDA (l99Ia), EPA (l987a)
Chemicals that occur naturally in foods are in bold face. Daily human exposure: Reasonable daily intakes are used to facilitate comparisons. The calculations assume a daily dose for a lifetime. Possible hazard: The human dose of rodent carcinogen is divided by 70 kg to give a mglkg/day of human exposure, and this dose is given as the percentage of the TDso in the rodent (mglkg/day) to calculate the human exposure/rodent potency (HERP) index. TDso values used in the HERP calculation are averages calculated by taking the harmonic mean (see Section 38.8) of the TDsos of the positive tests in that species from the Carcinogenic Potency Database. Average TDso values, have been calculated separately for rats and mice, and the more potent value is used for calculating possible hazard. a., no data in the CPDB; a number in parentheses indicates a TDso value not used in the HERP calculation because the TDso is less potent than in the other species; (-), negative in cancer tests; (+), positive cancer testes) not suitable for calculating a TDSO' bThe TDso harmonic mean was estimated for the base chemical from the hydrochloride salt. C Additional data from the EPA that were not in the CPDB were used to calculate this TDSO harmonic mean. dNo longer contained in any registered pesticide product (EPA, 1998).
posure amounts are for concentrations of a chemical in an individual food (i.e., foods for which data are available on concentration and average consumption). Table 38.5 indicates that many ordinary foods would not pass the regulatory criteria used for synthetic chemicals if the same methodology were used for both naturally occurring and synthetic chemicals. For many natural chemicals, the HERP
values are in the top half of the table, even though natural chemicals are markedly underrepresented because so few have been tested in rodent bioassays. We will discuss several categories of exposure and indicate that mechanistic data are available for some chemicals, which suggest that the possible hazard may not be relevant to humans or would be low if nonlinearity or a threshold were taken into account in risk assessment.
810
CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
Occupational Occupational and pharmaceutical exposures to some chemicals have been high, and many of the single chemical agents or industrial processes evaluated as human carcinogens have been identified by historically high exposures in the workplace (Tomatis and Bartsch, 1990; IARC, 19711999). HERP values rank at the top of Table 38.5 for past chemical exposures in some occupations to ethylene dibromide, 1,3-butadiene, tetrachloroethylene, formaldehyde, acrylonitrile, trichloroethylene, and methylene chloride. When exposures are high, the margin of exposure from the carcinogenic dose in rodents is low. The issue of how much human cancer can be attributed to occupational exposure has been controversial, but a few percent seems a reasonable estimate (Ames et aI., 1995). In another analysis, we have used permitted exposure limits (PELs), recommended in 1989 by the U.S. Occupational Safety and Health Administration (OSHA), as surrogates for actual exposures and compared the permitted daily dose rate for workers, with the TD50 in rodents [PERP (permitted exposure/rodent potency) index] (Gold et aI., 1987b, 1994a). We found that the PELs for 9 chemicals were greater than 10% of the rodent carcinogenic dose and for 27 they were between 1 and 10% of the rodent dose. The 1989 PELS were vacated by the Supreme Court because of a lack of risk assessment on each individual chemical. For the PELs that are currently the legal standard, PERP values for 14 chemicals are greater than 10%. For trichloroethylene, we recently conducted an analysis based on an assumed cytotoxic mechanism of action and PBPKeffective dose estimates defined as peak concentrations. Our estimates indicate that occupational respiratory exposures at the PEL for trichloroethylene would produce metabolite concentrations that exceed an acute no observed effect level (NOEL) for hepatotoxicity in mice. On this basis, the OSHA PEL is not expected to be protective. In comparison the EPA maximum concentration limit (MCL) in drinking water of 5 J..Lg/l, based on a linearized multi stage model, is more stringent than our estimate of an MCL based on a 1000-fold safety (uncertainty) factor, which is 210 J..Lg/l (Bogen and Gold, 1997). Pharmaceuticals Some pharmaceuticals that are used chronically are clustered near the top of the HERP ranking (e.g., phenobarbital, clofibrate, and fluvastatin). In Table 38.3, we reported that 49% of the drugs in the PDR with cancer test data are positive in rodent bioassays (Davies and Monro, 1995), as are 44% of drug submissions to the FDA (Contrera et aI., 1997). Most drugs, however, are used for only short periods, and the HERP values for the rodent carcinogens would not be comparable to the chronic, long-term administration used in HERP. Assuming a hypothetical lifetime exposure at therapeutic doses (i.e., not averaged over a lifetime), the HERP values would be high-for example, phenacetin (0.3%), metronidazole (5.6%), and isoniazid (14%). Herbal supplements have recently developed into a large market in the United States; they have not, however, been a focus of carcinogenicity testing. The FDA regulatory requirements for safety and efficacy that are applied to pharmaceutical
drugs do not pertain to herbal supplements under the 1994 Dietary Supplements and Health Education Act (DSHEA), and few have been tested for carcinogenicity. Those that are rodent carcinogens tend to rank high in HERP because, similar to some pharmaceutical drugs, the recommended dose is high relative to the rodent carcinogenic dose. Moreover, under DSHEA, the safety criteria that have been used for decades by the FDA for food additives that are "generally recognized as safe" (GRAS) are also not applicable to dietary supplements (Burdock, 2000) even though supplements are used at higher doses. The NTP is currently testing several herbs or chemicals in herbs. Comfrey is a medicinal herb whose roots and leaves have been shown to be carcinogenic in rats. The formerly recommended dose of 9 daily comfrey-pepsin tablets has a HERP value of 6.2%. Symphytine, a pyrrolizidine alkaloid plant pesticide that is present in comfrey-pepsin tablets and comfrey tea, is a rodent carcinogen; the HERP value for symphytine is 1.3% in the comfrey pills and 0.03% in comfrey herb tea. Comfrey pills are no longer widely sold, but are available on the World Wide Web. Comfrey roots and leaves can be bought at health food stores and on the Web and can thus be used for tea, although comfrey is recommended for topical use only in the PDR for Herbal Medicines (Gruenwald et aI., 1998). Poisoning epidemics by pyrrolizidine alkaloids have occurred in the developing world. In the United States, poisonings, including deaths, have been associated with use of herbal teas containing comfrey (Huxtable, 1995). Over 200 pyrrolizidine alkaloids are present in more than 300 plant species (Prakash et aI., 1999). Up to 3% of flowering plant species contain pyrrolizidine alkaloids (Prakash et aI., 1999). Several pyrrolizidine alkaloids have been tested chronically in rodent bioassays and are carcinogenic (Gold et aI., 1997b). Dehydroepiandrosterone (DHEA) and DHEA sulfate are the major secretion products of adrenal glands in humans and are precursors of androgenic and estrogenic hormones (Oelkers, 1999; van Vollenhoven, 2000). DHEA is manufactured and sold widely for a variety of purposes including the delay of aging. In rats, DHEA induces liver tumors (Rao et aI., 1992a; Hayashi et aI., 1994), and the HERP value for the recommended human dose of one daily capsule containing 25 mg DHEA is 0.5%. The mechanism of liver carcinogenesis in rats is peroxisome proliferation, similar to clofibrate (Ward et aI., 1998; Woodyatt et aI., 1999). DHEA also inhibited the development of tumors of the rat testis (Rao et al., 1992b) and rat and mouse mammary gland (Schwartz et aI., 1981; McCormick et aI., 1996). A recent review of clinical, experimental, and epidemiological studies concluded that late promotion of breast cancer in postmenopausal women may be stimulated by prolonged intake of DHEA (Stoll, 1999); however, the evidence for a positive association in postmenopausal women between serum DHEA levels and breast cancer risk is conflicting (Bemstein et al., 1990; Stoll, 1999). Natural Pesticides Natural pesticides, because few have been tested, are markedly underrepresented in our HERP analysis. More important, for each plant food listed, there are about 50 additional untested natural pesticides. Although about
38.5 The HERP Ranking of Possible Carcinogenic Hazards
10,000 natural pesticides and their breakdown products occur in the human diet (Ames et aI., 1990b), only 71 have been tested adequately in rodent bioassays (Table 38.1). Average exposures to many natural-pesticide rodent carcinogens in common foods rank above or close to the median in our HERP table (Table 38.5), ranging up to a HERP of 0.1%. These include caffeic acid (in coffee, lettuce, tomato, apple, potato, celery, carrot, plum, and pear); safrole (in spices and formerly in natural root beer before it was banned); allyl isothiocyanate (in mustard); d-limonene (in mango, orange juice, black pepper); coumarin (in cinnamon); and hydroquinone, catechol, and 4-methylcatechol (in coffee). Some natural pesticides in the commonly eaten mushroom (Agaricus bisporus) are rodent carcinogens (glutamyl-p-hydrazinobenzoate, p-hydrazinobenzoate), and the HERP based on feeding whole mushrooms to mice is 0.02%. For d-limonene, no human risk is anticipated because tumors are induced only in male rat kidney tubules with involvement of a2u-globulin nephrotoxicity, which does not appear to be relevant for humans, as discussed in Section 38.2 (Hard and Whysner, 1994; International Agency for Research on Cancer, 1993; Rice et aI., 1999; U.S. Environmental Protection Agency, 1991a).
Synthetic Pesticides Synthetic pesticides currently in use that are rodent carcinogens in the CPDB and that are quantitatively detected by the FDA Total Diet Study (TDS) as residues in food are all included in Table 38.5. Many are at the very bottom of the ranking; however, HERP values are about at the median for ethylene thiourea (ETU), UDMH (from Alar) before its discontinuance, and DDT before its ban in the United States in 1972. These three synthetic pesticides rank below the HERP values for many naturally occurring chemicals that are common in the diet. The HERP values in Table 38.5 are for residue intake by females 65 and older, because they consume higher amounts of fruits and vegetables than other adult groups, thus maximizing the exposure estimate to pesticide residues. We note that for pesticide residues in the TDS, average consumption estimates for children (mg/kg/day in 1986-1991) are within a factor of 3 of the adult consumption (mg/kg/day), greater in adults for some pesticides, and greater in children for others (U.S. Food and Drug Administration, 1993b). DDT and similar early pesticides have been a concern because of their unusuallipophilicity and persistence, even though there is no convincing epidemiological evidence of a carcinogenic hazard to humans (Key and Reeves, 1994) and although natural pesticides can also bioaccumulate. In a recently completed 24-year study in which DDT was fed to rhesus and cynomolgus monkeys for 11 years, DDT was not evaluated as carcinogenic (Takayama et aI., 1999; Thorgeirsson et aI., 1994) despite doses that were toxic to both liver and central nervous system. However, the protocol used few animals and dosing was discontinued after 11 years, which may have reduced the sensitivity ofthe study (Gold et al., 1999). The HERP value for DDT residues in food before the ban was 0.0008%. Current U.S. exposure to DDT and its metabolites is in foods of animal origin, and the HERP value is low, 0.00008%. DDT
811
is often viewed as the typically dangerous synthetic pesticide because it concentrates in adipose tissue and persists for years. DDT was the first synthetic pesticide; it eradicated malaria from many parts of the world, including the United States, and was effective against many vectors of disease such as mosquitoes, tsetse flies, lice, ticks, and fleas. DDT was also lethal to many crop pests and significantly increased the supply and lowered the cost of fresh, nutritious foods, thus making them accessible to more people. A 1970 National Academy of Sciences report concluded: "In little more than two decades DDT has prevented 500 million deaths due to malaria, that would otherwise have been inevitable" (National Academy of Sciences, 1970). DDT is unusual with respect to bioconcentration, and because of its chlorine substituents it takes longer to degrade in nature than most chemicals; however, these are properties of relatively few synthetic chemicals. In addition, many thousands of chlorinated chemicals are produced in nature (Gribble, 1996). Natural pesticides can also bioconcentrate ifthey are fat soluble. Potatoes, for example, naturally contain the fat-soluble neurotoxins solanine and chaconine (Ames et aI., 1990a; Gold et aI., 1997a), which can be detected in the bloodstream of all potato eaters. High levels of these potato neurotoxins have been shown to cause birth defects in rodents (Ames et al., 1990b). The HERP value for ethylene thiourea (ETU), a breakdown product of certain fungicides, is the highest among the synthetic pesticide residues (0.002%), which is at the median of the ranking. The HERP would be about 10 times lower if the potency value of the EPA were used instead of our TD50; the EPA combined rodent results from more than one experiment, including one in which ETU was administered in utero, and obtained a weaker potency value (U.S. Environmental Protection Agency, 1992). (The CPDB does not include in utero exposures.) Additionally, the EPA has recently discontinued some uses of fungicides for which ETU is a breakdown product; and therefore exposure levels and HERP values would be lower. In 1984, the EPA banned the agricultural use of ethylene dibromide (EDB), the main fumigant in the United States, because of the residue levels found in grain (HERP = 0.0004%). This HERP value ranks low, compared to the HERP of 140% for the high exposures to EDB that some workers received in the 1970s which is at the top of the ranking (Gold et aI., 1992). Two other pesticides in Table 38.5, toxaphene (HERP = 0.0002%) and chlorobenzilate (HERP = 0.0000001%), have been cancelled (Ames and Gold, 1991; U.S. Environmental Protection Agency, 1998). Most residues of synthetic pesticides have HERP values below the median. In descending order of HERP, these are carbaryl, toxaphene, dicofol, lindane, PCNB, chlorobenzilate, captan, folpet, and chlorothalonil. Some of the lowest HERP values in Table 38.5 are for the synthetic pesticides, captan, chlorothalonil, and folpet, which were also evaluated in 1987 by the National Research Council (NRC) and were considered by the NRC to have a human cancer risk above 10-6 (National Research Council, 1987). The contrast between the low HERP values for synthetic pesticide residues in our ranking and the higher NRC risk estimates is examined in Section 38.6.
812
CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
Cooking and Preparation of Food and Drink Cooking and preparation of food can also produce chemicals that are rodent carcinogens. Alcoholic beverages cause cancer in humans in the liver, esophagus, and oral cavity. The HERP values in Table 38.5 for alcohol in beer (2.1 %) and wine (0.5%) are high in the ranking. Ethyl alcohol is one of the least potent rodent carcinogens in the CPDB, but the HERP is high because of high concentrations in alcoholic beverages and high U.S. consumption. Another fermentation product, urethane (ethyl carbamate), has a HERP value of 0.00001 % for average beer consumption and 0.00007% for average bread consumption (as toast). Cooking food is plausible as a contributor to cancer. A wide variety of chemicals are formed during cooking. Rodent carcinogens formed include furfural and similar furans, nitrosamines, polycyclic hydrocarbons, and heterocyclic amines. Furfural, a chemical formed naturally when sugars are heated, is a widespread constituent of food ftavor. The HERP value for naturally occurring furfural in the average consumption of coffee is 0.02% and in white bread it is 0.004%. Furfural is also used as a commercial food additive, and the HERP for total average U.S. consumption as an additive is much lower (0.00006% ). Nitrosamines in food are formed by cooking from nitrite or nitrogen oxides (NO x ) and amines. Tobacco smoking and smokeless tobacco are a major source of nonoccupational exposure to nitrosamines that are rodent carcinogens [N -nitrosonornicotine and 4-(methylnitrosamino )-I-(3-pyridyl)-I-(butanone)] (Hecht and Hoffmann, 1998). Most exposure to nitrosamines in the diet is for chemicals that are not carcinogenic in rodents (Hecht and Hoffmann, 1998; Lijinsky, 1999). The nitrosamines that are carcinogenic are potent carcinogens (Table 38.5), and it has been estimated that in several countries humans are exposed to about 0.31 J1.g/day (National Academy of Sciences, 1981; Tricker and Preussmann, 1991), primarily N-nitrosodimethylamine (DMN), N -nitrosopyrrolidine, and N -nitrosopiperidine. The largest exposure is to DMN in beer: Concentrations declined more than 30-fold after 1979 (HERP = 0.01 %) when it was reported that DMN was formed by the direct-fired drying of malt, and the industry modified the process to indirect firing (Gloria et al., 1997). By the 1990s, the HERP was 0.0003% (Gloria et aI., 1997). The HERP values for the average consumption of bacon are lower: DMN = 0.0005%, DEN = 0.0006%, and NPYR = 0.0004%. DEN induced liver tumors in rhesus and cynomolgus monkeys and tumors of the nasal mucosa in bush babies (Thorgeirsson et aI., 1994). In a study of DMN in rhesus monkeys, no tumors were induced; however, the administered doses produced toxic hepatitis, and all animals died early. Thus, the test was not sensitive because the animals may not have lived long enough to develop tumors (Gold et al., 1999; Thorgeirsson et aI., 1994). A variety of mutagenic and carcinogenic heterocyclic amines (HAs) are formed when meat, chicken, and fish are cooked, particularly when charred. Compared to other rodent carcinogens, there is strong evidence of carcinogenicity for HAs in terms of positivity rates and multiplicity of target sites; however, con-
cordance in target sites between rats and mice for these HAs is generally restricted to the liver (Gold et aI., 1994b). Under usual cooking conditions, exposures to HAs are in the low ppb range, and the HERP values for pan-fried hamburger are low. The HERP value for PhIP is 0.00006%, for MeIQx it is 0.00003%, and for IQ it is 0.000006%. Carcinogenicity of the three HAs in the HERP table, IQ, MeIQx, and PhIP, has been investigated in studies in cynomolgus monkeys. IQ rapidly induced a high incidence of hepatocellular carcinoma (Adamson et aI., 1994). MeIQx, which induced tumors at multiple sites in rats and mice (Gold et aI., 1997c), did not induce tumors in monkeys (Ogawa et aI., 1999). The PhIP study is in progress. Metabolism studies indicate the importance of N -hydroxylation in the carcinogenic effect of HAs in monkeys (Snyderwine et aI., 1997). IQ is activated via N -hydroxylation and forms DNA adducts; the Nhydroxylation of IQ appears to be carried out largely by hepatic CYP3A4 and/or CYP2C9110, and not by CYPIA2; whereas the poor activation of MeIQx appears to be due to a lack of expression of CYPIA2 and an inability of other cytochromes P450, such as CYP3A4 and CYP2C9/1O, to N-hydroxylate the quinoxalines. PhIP is activated by N -hydroxylation in monkeys and forms DNA adducts, suggesting that it might turn out to have a carcinogenic effect (Ogawa et aI., 1999; Snyderwine et aI., 1997). Food Additives Food additives that are rodent carcinogens can be either naturally occurring (e.g., allyl isothiocyanate, furfural, and alcohol) or synthetic (e.g., BHA and saccharin; Table 38.5). The highest HERP values for average dietary exposures to synthetic rodent carcinogens in Table 38.5 are for exposures in the early 1970s to BHA (0.01 %) and saccharin in the 1970s (0.005%). Both are nongenotoxic rodent carcinogens for which data on the mechanism of carcinogenesis strongly suggest that there would be no risk to humans at the levels found in food. BHA is a phenolic antioxidant that is "generally regarded as safe" (GRAS) by the FDA. By 1987, after BHA was shown to be a rodent carcinogen, its use declined sixfold (HERP = 0.002%) (U.S. Food and Drug Administration, 1991b); this was due to voluntary replacement by other antioxidants and to the fact that the use of animal fats and oils, in which BHA is primarily used as an antioxidant, has consistently declined in the United States. The mechanistic and carcinogenicity results on BHA indicate that malignant tumors were induced only at a dose above the MTD at which cell division was increased in the forestomach, which is the only site of tumorigenesis; the proliferation is only at high doses and is dependent on continuous dosing until late in the experiment (Clayson et aI., 1990). Humans do not have a forestomach. We note that the doseresponse relationship for BHA curves sharply upward, but the potency value used in HERP is based on a linear model; if the California EPA potency value (which is based on a linearized multistage model) were used in HERP instead of the TDso, the HERP values for BHA would be 25 times lower (California Environmental Protection Agency, 1994). A recent epidemiological study in the Netherlands found no association between
38.5 The HERP Ranking of Possible Carcinogenic Hazards
BHA consumption and stomach cancer in humans (Botterweck et aI., 2000). Saccharin, which has largely been replaced by other sweeteners, has been shown to induce tumors in rodents by a mechanism that is not relevant to humans. Recently, both the NTP and the IARC reevaluated the potential carcinogenic risk of saccharin to humans. The NTP delisted saccharin in its Report on Carcinogens (U.S. National Toxicology Program, 2000a), and the IARC downgraded its evaluation to Group 3, "not classifiable as to carcinogenicity to humans" (International Agency for Research on Cancer, 1971-1999). There is convincing evidence that the induction of bladder tumors in rats by sodium saccharin requires a high dose and is related to the development of a calcium phosphate-containing precipitate in the urine (Cohen, 1995a), which is not relevant to human dietary exposures. In a recently completed 24-year study by the NCI, rhesus and cynomolgus monkeys were fed a dose of sodium saccharin that was equivalent to 5 cans of diet soda daily for 11 years (Thorgeirsson et aI., 1994). The average daily dose rate of sodium saccharin (mg/kg/day) was about 100 times lower than the dose that was carcinogenic to rats (Gold et aI., 1997c, 1999). There was no carcinogenic effect in monkeys. There was also no effect on the urine or urothelium, no evidence of increased urothelial cell proliferation or of formation of solid material in the urine (Takayama et aI., 1998). One would not expect to find a carcinogenic effect under the conditions of the monkey study because of the low dose administered. Additionally, however, there may be a true species difference because primate urine has a low concentration of protein and is less concentrated (lower osmolality) than rat urine (Takayama et aI., 1998). Human urine is similar to monkey urine in this respect (Cohen, 1995a). For three naturally occurring chemicals that are also produced commercially and used as food additives, average exposure data are available and they are included in Table 38.5. The HERP values are as follows: For furfural, the HERP value for the natural occurrence is 0.02% compared to 0.00006% for the additive; for d-limonene, the natural occurrence HERP is 0.1 % compared to 0.003% for the additive; and for estragole, the HERP is 0.00005% for both the natural occurrence and the additive. Safrole is the principal component (up to 90%) of oil of sassafras. It was formerly used as the main flavor ingredient in root beer. It is also present in the oils of basil, nutmeg, and mace (Nijssen et aI., 1996). The HERP value for average consumption of naturally occurring safrole in spices is 0.03%. In 1960, safrole and safrole-containing sassafras oils were banned from use as food additives in the United States (U.S. Food and Drug Administration, 1960). Before 1960, for a person consuming a glass of sassafras root beer per day for life, the HERP value would have been 0.2% (Ames et aI., 1987). Sassafras root can still be purchased in health food stores and can therefore be used to make tea (Heikes, 1994); the recipe is on the World Wide Web. Mycotoxins Of the 23 fungal toxins tested for carcinogenicity, 14 are positive (61%) (Table 38.3). The mutagenic mold
813
toxin, aflatoxin, which is found in moldy peanut and corn products, interacts with chronic hepatitis infection in human liver cancer development (Qian et aI., 1994). There is a synergistic effect in the human liver between aflatoxin (genotoxic effect) and the hepatitis B virus (cell division effect) in the induction of liver cancer (Wu-Williams et aI., 1992). The HERP value for aflatoxin of 0.008% is based on the rodent potency. If the lower human potency value calculated from epidemiological data by the FDA were used instead, the HERP would be about lO-fold lower (U.S. Food and Drug Administration, 1993b). Biomarker measurements of aflatoxin in populations in Africa and China, which have high rates of hepatitis Band C viruses and liver cancer, confirm that those populations are chronically exposed to high levels of aflatoxin (Groopman et aI., 1992; Pons, 1979). Liver cancer is unusual in the United States. Hepatitis viruses can account for half of liver cancer cases among non-Asians and even more among Asians in the United States (Yu et ai., 1991). Ochratoxin A, a potent rodent carcinogen (Gold and Zeiger, 1997), has been measured in Europe and Canada in agricultural and meat products. An estimated exposure of 1 ng/kg/day would have a HERP value close to the median of Table 38.5 (International Life Sciences Institute, 1996; Kuiper-Goodman and Scott, 1989). Synthetic Contaminants Polychlorinated biphenyls (PCBs) and tetrachlorodibenzo-p-dioxin (TCDD), which have been a concern because of their environmental persistence and carcinogenic potency in rodents, are primarily consumed in foods of animal origin. In the United States, PCBs are no longer used, but some exposure persists. Consumption in food in the United States declined about 20-fold between 1978 and 1986 (Gartrell et aI., 1986; Gunderson, 1995). The HERP value for the most recent reporting of the FDA Total Diet Study (19841986) is 0.00008%, toward the bottom of the ranking, and far below many values for naturally occurring chemicals in common foods. It has been reported that some countries may have higher intakes of PCBs than the United States (World Health Organization, 1993). TCDD, the most potent rodent carcinogen, is produced naturally by burning when chloride ion is present, for example, in forest fires or wood burning in homes. The EPA (U.S. Environmental Protection Agency, 2000) proposes that the source of TCDD is primarily from the atmosphere directly from emissions (e.g., incinerators) or indirectly by returning dioxin to the atmosphere (U.S. Environmental Protection Agency, 2000). TCDD bioaccumulates through the food chain because of its lipophilicity, and more than 95% of human intake is from animal fats in the diet (U.S. Environmental Protection Agency, 2000). Dioxin emissions decreased by 80% from 1987 to 1995, which the EPA attributes to reduced emissions from incineration of medical and municipal waste (U.S. Environmental Protection Agency, 2000). The HERP value of 0.0004% for average U.S. intake of TCDD (U.S. Environmental Protection Agency, 2000) is below the median of the values in Table 38.6. Recently, the EPA
814
CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
Table 38.6 Tumor Incidence Data Used in Recalculations of Carcinogenic Potency for 19 Chemicals in the NRC Report Weeks
Sex-
Pesticidea
on test
species b
Target organ
(mg/kglday)C
AcephateNA(Cnq)
105
FM
Liver
0,7.5,37.5, 150
1162, 3/61, 0/62, 15/61
AlachloJ>2(MOE)
110
FR
Nasal Turbinate
0,0.5,2.5, 15
0/44,0/47,0/44, 15/45
Dose groups
MR
Tumor incidence
108
MR
Thyroid gland
0,36,180,953
0/43,9143,7143, (2/4W
AzinphosmethylD(E)
114e
MR
Thyroid gland
0,3.9,7.8
119, 10/44, 12/43
(Guthion) BenomylCq
104
FM
Liver
0,75,225, 1130
1174,9170,20175, 15175
104
FR
Liver
0,2.8, 12.1,54.8
4/50, 2/49, 3/50, 17150
MR
Kidney
FM
Digestive tract
Captan B2
113
0,879,1480,2370
104
FM
FM
0,15,60,120,900
31.6
4,400 202
3/80, 26/80, 21180, 29/80
4,480
O!I 00, 111 00, 3/100, 411 00, 91100
01100,71100, 11100, 11100,71100
Hematopoietic
0, 0.3, 3, 30, 75
3/38, 1135, 11142, 31139, 34/41
21.7
3/47, 1146, 12/46, 32/47,40/47
MM ChlorothaloniINA (B2)(MOE) CypermethrinCq(Cnq)
129
PR
Kidney
0, 40, 80, 175
0/59,2/60,7/57, 19/58
101
FM
Lung
0,15,60,240
12/127, 6/64, 8/64, 14/61
FolpetB2
113
FM
Digestive tract
0,96,515,1280
0/104, 1180, 8/80,41180
0,93,502,1280
11104, 2/80, 8/80, 41180
Adrenal gland
0, 100,400, 1510
6/80,7/78, 16179, (18/80)d
MM Fosetyl AICq(Cnq)(Unclassified)
724
3/80, 19/80, 22/80, 39/80
MM Chlordimeform B2
36.8
1/50, 1150, 0/50, 7/50
MM 95
499
0/42,0/42, 1147, 14/48
AsulamNA(Cnq)
Captafol B2
TDso (mglkglday)
566 954 1,910
104
MR
GlyphosateCq(E)
104
MM
Kidney
0,150,750,4500
1149,0/49, 1150, 3/50
LinuronCq(Cnq)
104
MR
Testis
0, 2.5, 6.25, 31.3
4170,9/69,20170, 37/70
MetolachlorCq(Cnq)(MOE)
104
FR
Liver
0, 1.5, 15, 150
0/60, 1160, 2/60, 7/60
839
OryzalinCq
104
PR
Skin
0, 15,45, 135
1160, 2/60, 4/60, 9/60
394
1,860
(Aliette)
105
FM
Liver
0,15,45,150,300
112
PR
Adrenal gland
MR PermethrinCq
104
FM
4/56, 13/6 I, 18/64, 27/55, 32/57
213
20/64,40/67,52/69,44/65,28/35
MM ParathionCq(Cnq)
28.1
5/60, 6/60, 6/60, 24/59
MR Oxadiazon B2 (Cq)
62,000
Lung
0, 1.15, 2.25
1110,6/47, 13142
0, 1.6,3.15
0/9,7/49, 11146
0,3,375,750
15171,24/68, 35/68,44/69
7.95
717
aEPA weight-of-evidence evaluation reported as superscript. If more than one classification is reported, the first values are from the NRC report and values in parentheses are from the EPA's revised evaluations since 1987 (Burnam, 2000; Irene, 1995). B2: Sufficient evidence of carcinogenicity from animal studies with inadequate or no epidemiologic data-probable human carcinogen. Cq: Limited evidence of carcinogenicity from animal studies in the absence of human data-possible human carcinogen (quantifiable). Cnq: Limited evidence (not quantified by the EPA, i.e., no D: Human and animal data are either inadequate or absent-not classifiable as to human carcinogenicity. E: Evidence of noncarcinogenicity to humans. NA indicates that the chemical was not classified at the time of the NRC report. MOE: The Health Effects Division Carcinogenicity Peer Review Committee (HCPRC) recommended under the newly proposed EPA guidelines a margin-of-exposure approach for risk assessment for these three chemicals: alachlor, chlorothalonil, and metolachlor. For alachlor, the current Office of Pesticide Programs (OPP) classification is "Likely (high doses), Not Likely (low doses). " For chlorothalonil, the classification is "Likely" with recommendation for a nonlinear approach to risk assessment. Unclassified: For fosetyl AI, the HCPRC concluded that it "was not amenable to classification using current Agency cancer guidelines. The HCPRC concluded that pesticidal use of fosetyl-Al is unlikely to pose a carcinogenic hazard to humans" (Burnam, 2000). Captafol and chlordimeform uses have been canceled (U.S. Environmental Protection Agency, 1998). bFM, female mouse; MM, male mouse; FR, female rat; MR, male rat. If more than one group is reported, the potency calculation is a geometric mean of the TDsO for the experiments in this table only. cUnless mg/kg/day are given in the EPA memorandum, doses are converted from ppm to mg/kg body weight/day by standard EPA conversion factors: 0. 05 for rats and 0. 15 for mice. All chemicals were administered in the diet. dDoses in parentheses were not used in the calculation of either the TDso or the EPA qT. For fosetyl AI, the adrenal gland qT most closely replicated the NRC qj; in later EPA documents, urinary bladder was the target site and results were not considered appropriate for quantification (Quest et ai., 1991). eDosing was only for 80 weeks.
qn.
38.5 The HERP Ranking of Possible Carcinogenic Hazards
has reestimated the potency of TCDD based on a change in the dose-metric to body burden in humans (rather than intake) (U.S. Environmental Protection Agency, 2000) and a reevaluation of tumor data in rodents (which determined two-thirds fewer liver tumors) (Goodman and Sauer, 1992). Using this EPA potency for HERP would put TCDD at the median of HERP values in Table 38.6, 0.002%. TCDD exerts many of its harmful effects in experimental animals through binding to the Ah receptor (AhR) and does not have effects in the AhR knockout mouse (Birnbaum, 1994; Fernandez-Salguero et aI., 1996). A wide variety of natural substances also bind to the AhR (e.g., tryptophan oxidation products), and insofar as they have been examined, they have similar properties to TCDD (Ames et aI., 1990b), including inhibition of estrogen-induced effects in rodents (Safe et aI., 1998). For example, a variety of flavones and other plant substances in the diet and their metabolites also bind to the AhR [e.g., indole-3-carbinol (BC)]. BC is the main breakdown compound of glucobrassicin, a glucosinolate that is present in large amounts in vegetables of the Brassica genus, including broccoli, and gives rise to the potent Ah binder indole carbazole (Bradfield and Bjeldanes, 1987). The binding affinity (greater for TCDD) and the amounts consumed (much greater for dietary compounds) both need to be considered in comparing possible harmful effects. Some studies provide evidence of enhancement of carcinogenicity by BC (Dashwood, 1998). Additionally, both BC and TCDD, when administered to pregnant rats, resulted in reproductive abnormalities in male offspring (Wilker et aI., 1996). Currently, BC is in clinical trials for prevention of breast cancer (Kelloff et aI., 1996a, b; U.S. National Toxicology Program, 2000b) and is also being tested for carcinogenicity in rodents by NTP (U.S. National Toxicology Program, 2000b). BC is marketed as a dietary supplement at recommended doses about 30 times higher (Theranaturals, 2000) than present in the average Western diet (U.S. National Toxicology Program, 2000b). TCDD has received enormous scientific and regulatory attention, most recently in an ongoing assessment by the EPA (U.S. Environmental Protection Agency, 1994a, 1995a, 2000). Some epidemiologic studies suggest an association with cancer mortality. In 1997 the IARC evaluated the epidemiological evidence for carcinogenicity of TCDD in humans as limited (International Agency for Research on Cancer, 1997). The strongest epidemiological evidence was among highly exposed workers for overall cancer mortality. There is no sufficient evidence in humans for any particular target organ. Estimated blood levels of TCDD in studies of those highly exposed workers were similar to blood levels in rats in positive cancer bioassays (International Agency for Research on Cancer, 1997). In contrast, background levels of TCDD in humans are about 100- to WOO-fold lower than in the rat study. The similarities of worker and rodent blood levels and the mechanism of the AhR in both humans and rodents were considered by the IARC when it evaluated TCDD as a Group 1 carcinogen in spite of only limited epidemiological evidence. The IARC also concluded that "Evaluation of the relationship between the
815
magnitude of the exposure in experimental systems and the magnitude of the response, (i.e., dose-response relationships) do not permit conclusions to be drawn on the human health risks from background exposures to 2,3,7,8-TCDD." The NTP Report on Carcinogens recently evaluated TCDD in an addendum to the Ninth Report on Carcinogens as a known human carcinogen (U.S. National Toxicology Program, 2000a, 2001). The EPA draft final report (U.S. Environmental Protection Agency, 2000) characterized TCDD as a "human carcinogen," but concluded that "there is no clear indication of increased disease in the general population attributable to dioxin-like compounds" (U.S. Environmental Protection Agency, 2000). Possible limitations of data or scientific tools were given by the EPA as possible reasons for the lack of observed effects. In summary, the HERP ranking in Table 38.5 indicates that when synthetic pesticide residues in the diet are ranked on an index of possible carcinogenic hazard and compared to the ubiquitous exposures to rodent carcinogens, they rank low. Widespread exposures to naturally occurring rodent carcinogens cast doubt on the relevance to human cancer of low-level exposures to synthetic rodent carcinogens. In regulatory efforts to prevent human cancer, the evaluation of low-level exposures to synthetic chemicals has had a high priority. Our results indicate, however, that a high percentage of both natural and synthetic chemicals are rodent carcinogens at the MTD, that tumor incidence data from rodent bioassays are not adequate to assess low-dose risk, and that there is an imbalance in testing of synthetic chemicals compared to natural chemicals. There is an enormous background of natural chemicals in the diet that rank high in possible hazard, even though so few have been tested in rodent bioassays. In Table 38.5, 90% of the HERP values are above the level that would approximate a regulatory virtually safe dose of 10-6 if a qualitative risk assessment were performed. Caution is necessary in drawing conclusions from the occurrence in the diet of natural chemicals that are rodent carcinogens. It is not argued here that these dietary exposures are necessarily of much relevance to human cancer. In fact, epidemiological results indicate that adequate consumption of fruits and vegetables reduces cancer risk at many sites (Block et aI., 1992) and that protective factors like the intake of vitamins such as folic acid are important, rather than the intake of individual rodent carcinogens. The HERP ranking also indicates the importance of data on the mechanism of carcinogenesis for each chemical. For several chemicals, data have recently been generated that indicate that exposures would not be expected to be a cancer risk to humans at the levels consumed (e.g., saccharin, BHA, chloroform, d-limonene, discussed previously). Standard practice in regulatory risk assessment for chemicals that induce tumors in high-dose rodent bioassays has been to extrapolate risk to low dose in humans by multiplying potency by human exposure. Without data on the mechanism of carcinogenesis, however, the true human risk of cancer at low dose is highly uncertain and could be 0 (Ames and Gold, 1990; Clayson and Iverson, 1996; Gold et aI., 1992; Goodman, 1994). Adequate risk assess-
816
CHAPTER 38 Pesticide Residues in Food and Cancer Risk: A Critical Analysis
ment from animal cancer tests requires more information for a chemical about pharmacokinetics, mechanism of action, apoptosis, cell division, induction of defense and repair systems, and species differences.
38.6 PESTICIDE RESIDUES IN FOOD: INVESTIGATION OF DISPARITIES IN CANCER RISK ESTIMATES There are large disparities in the published cancer risk estimates for synthetic pesticide residues in the U.S. diet. In our HERP ranking in Table 38.5, the possible carcinogenic hazards of such residues rank low when viewed in the broadened perspective of exposures to naturally occurring chemicals that are rodent carcinogens. This section examines the extent to which disparities in risk estimates are due to differences in potency estimation from rodent bioassay data (qr vs. TDso) or to differences in estimation of human dietary exposure (Theoretical Maximum Residue Contribution vs. Total Diet Study). Our analysis is based on risk estimates for 29 pesticides, herbicides, and fungicides that were published by the National Research Council (NRC) in its 1987 report, Regulating Pesticides in Food: The Delaney Paradox (National Research Council, 1987). The NRC used potency and exposure estimates of the EPA and concluded that dietary risks for 23 pesticides were greater than one in a million and therefore were not negligible. The methodologies to estimate both potency and exposure differed between the NRC and our HERP index, and these differences are examined here to explain the difference in evaluation of possible cancer hazards from synthetic pesticide residues. For both the EPA and the HERP, risk estimation uses a linear extrapolation and is simply potency x dose. Our analysis below indicates that the disparities in risk estimates are due to widely different exposure estimates, rather than to different estimated values of carcinogenic potency. The NRC report used the standard regulatory default methodology of the EPA to estimate risk, that is, to evaluate the weight of evidence of carcinogenicity for a chemical from chronic rodent bioassays and extrapolate risk using an upper bound estimate of potency (qn and the 1inearized multi stage model (LMS) (Crump, 1984). Our HERP ranking used the TDso (the tumorigenic dose rate for 50% of test animals) as a measure of potency, and the HERP index is a simple proportion: exposure/potency (Section 38.4). To compare potency estimates, we first attempted to reproduce the tumor site and incidence data and the EPA qr values reported by the NRC so that we could use the correct data to estimate the TDso and then compare the two estimates. The NRC report did not present the tumor incidence data, and for most experiments the results did not appear in the general published literature. We obtained the results from EPA memoranda and personal communications (Table 38.6). The NRC report and the HERP ranking used two different estimates of human exposure to pesticide residues in the diet. The NRC used the EPA Theoretical Maximum Residue Contribution (TMRC), whereas the HERP ranking used the FDA
Total Diet Study (TDS). The TMRC is a theoretical maximum exposure, whereas exposure in the TDS is measured as dietary residues in table-ready food. We assess the magnitude of the differences between the two potency estimates q rand TDso when both use the same rodent results and then compare the differences between the two exposure estimates, TMRC and TDS, in order to determine the basis for disparate risk estimates. Since publication of the NRC report in 1987, the EPA has made several changes in the risk estimates of some pesticides. We discuss these changes, including: reevaluations of the weight of evidence of carcinogenicity using rodent bioassay results, changes in whether risks should be quantified, changes in exposure estimation, and proposed changes in risk assessment methodology. 38.6.1 REPRODUCIBILITY OF THE qi VALUES
The NRC, in Regulating Pesticides in Food: The Delaney Paradox (National Research Council, 1987), examined the potential human cancer risk for a group of synthetic herbicides, insecticides, and fungicides that the EPA had classified as to carcinogenicity based on rodent bioassay data. The NRC reported the following EPA data: (1) carcinogenic potency (qr>; (2) an upper bound estimate of hypothetical, lifetime daily human exposure, TMRC; and (3) an upper bound estimate of excess cancer risk over a lifetime, calculated as potency x exposure. We obtained data from the EPA for 19 of the 26 chemicals discussed by the NRC (Quest et aI., 1993; U.S. Environmental Protection Agency, 1984a, 1985-1988, 1985a, 1985b, 1986b, 1987b, 1988b, 1989b, 1989c, 1999a). We were not able to identify the animal data used in the NRC report for cryomazine, diclofop methyl, ethalfluralin, ethylene thiourea, o-phenylpheno1, pronamide, and terbutryn. To verify that we had identified the correct rodent results, we attempted to replivalue for each of the 19 pesticides to define cate the EPA the data set for our comparison of risk estimates. The Tox-Risk program (Crump & Assoc.) was used to calculate as the 95% upper confidence limit on the linear term in the LMS, which theoretically represents the slope of the dose-response curve in the low-dose region. If it was not clear which target site had been used by the EPA, we calculated more than one and used in our subsequent comparison of potency estimates whichever value. If the EPA memorandata best reproduced the EPA dum for a chemical stated that the was the geometric mean of two or more experiments, we used the same method. The bioassay data that most accurately reproduced the EPA for each chemical are given in Table 38.6. Superscripts indicate the EPA weight-of-evidence classification given in the NRC report, followed by subsequent reevaluations of the classification. Using the data in Table 38.6 with the Tox-Risk program, overall there was good reproducibility in potency estimation (Table 38.7). We were able to reproduce the EPA qr value for
q;
q;
q;
q;
q;
q;
38.6 Pesticide Residues in Food: Investigation of Disparities in Cancer Risk Estimates Table 38.7 Reproducibility of the EPA
qf Values Reported by the NRC
Pesticide
EPA qi reported by NRC (mg/kg/day)-I
Chlorothalonil
2.4 x 10- 2
Asulam
2.0 x 10- 2 3.4 x 10- 2
Oryzalin Permethrin Chlordimeform Fosetyl Al Captafol Oxadiazon Cypermethrin Folpet Linuron Captan Alachlor Acephate Benomyl Metolachlor Glyphosate Parathion Azinphosmethyl
817
3.0 x 10- 2 9.4 x 10- 1 4.3 x 10- 3
Recalculated qi
Recalculated q
(mg/kg/day)-I
EPAqi
f/
1.3 x 10- 2 1.4 x 10- 2
O.S 0.7
2.S x 10- 2 2.0 x 10- 2 7.2 x 10- 1 3.7 x 10- 3
0.7 0.7 0.8
2.S x 10- 2 1.3 x 10- 1
2.4 x 10- 2
0.9 1.0
1.3 x 10- 1
1.0
1.9 x 10- 2 3.S x 10- 3
2.1 x 10- 2 3.8 x 10- 3
1.1
3.3 x 10- 1 2.3 x 10- 3
3.7 x 10- 1 3.4 x 10- 3
1.1
6.0 x 10- 2 6.9 x 10- 3 2.1 x 10- 3 2.1 x 10- 3
9.S x 1.3 x 4.6 x 8.7 x
1.6
S.9 x 10- 5 1.8 x 10- 3 I.S x 10- 7
10- 2 10- 2 10- 3 10- 3 4.8 x 10-4 1.3 x 10°
7.3 x 10- 1
1.1
I.S 1.9 2.2 4.1 6.1 720 4,900,000
Recalculated qj uses the bioassay data in Table 38.6 and a linearized multistage model.
15 chemicals within a factor of 2.2, and for 17 within a factor of 6. The median ratio of the q{ reported by the NRC to the recalculated q{ is 1.1. We could not approximate the q{ for parathion or azinphosmethyl. The q{ published in the NRC report for azinphosmethyl appears to be an error (W. Bumam, Office of Pesticide Programs, EPA, personal communication). We concluded that the data set of 15 chemicals with a q{ reproducibility within a factor of 2.2 would be used in the comparison of risk estimates. The four-chemicals for which we could not reproduce the q{ within a factor of 2.2 have all been reevaluated by the EPA since the NRC report: Azinphosmethyl and glyphosate are considered to have evidence of noncarcinogenicity to humans (i.e., superscript E in Table 38.6) (Bumam, 2000); a margin-of-exposure approach is recommended for metolachlor (MOE in Table 38.6); and parathion is classified as having limited evidence without a q{ value (Cnq) in Table 38.6. 38.6.2 COMPARISON OF POTENCY ESTIMATES: q~ AND TDso
Using the incidence data identified as those used by the EPA (Table 38.6), we estimated the TDso, that is, the dose rate (in mg/kg body weight/day) that is estimated to reduce by 50% the proportion of tumor-free animals at the end of a standard lifespan (Peto et aI., 1984; Sawyer et aI., 1984). The TDso does not involve extrapolation to low dose. It is inversely related to the slope (Peto et aI., 1984; Sawyer et aI., 1984; see Section 38.8 for details), and a comparison with q{ can be made by
using In(2)jTDso. An adjustment for rodent-to-human extrapolation, such as a surface area or other allometric correction factor, is usually applied to the q{ for regulatory purposes. For comparison purposes, the TDso was adjusted by the same interspecies scaling factor that was used by the EPA for q{, that is, (body weight)2/3, a factor of approximately 5.5 for rats and 13.0 for mice. The two potency estimates were then compared by computing the ratio q{ j(ln(2)jTDso). The dose calculation and standardization methods used for the TDso calculation in this chapter follow the EPA methods, some of which differ from the standard methodology used to estimate TDso in the CPDB. 38.6.3 COMPARISON OF HUMAN EXPOSURE ESTIMATES
The risk estimates in the NRC report (National Research Council, 1987) differed from those in the HERP ranking for dietary residues of synthetic pesticides (Section 38.5). The NRC reported upper bound estimates of daily human exposure (i.e., the EPA TMRC). In contrast, the HERP values in Table 38.5 used the daily exposure estimates from the FDA Total Diet Study (TDS). Thirteen pesticides discussed in the NRC report were measured in the TDS, and we compared the exposure estimates from the two sources for these 13. We used results from the TDS for the years 1984-1986 (Gunderson, 1995; U.S. Food and Drug Administration, 1988), which are the closest to the time of the NRC report.
818
CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
The EPA TMRC is a theoretical maximum estimate for potential human dietary exposure to synthetic pesticides. Pesticides registered for food crop use in the United States must first be granted tolerances under the Federal Food, Drug and Cosmetic Act (FFDCA). Tolerances are the maximum, legally allowable residues of the pesticide, or its active ingredient, on raw agricultural commodities and in processed foods. A tolerance is typically set for each pesticide for each crop use (e.g., corn, barley, wheat) based on field trials. The manufacturer conducts these trials, using varying rates of application under diverse environmental conditions, to determine both the minimum application rate needed to be effective against pest targets and the duration of time before harvest when it has to be applied (these are the rates specified on the pesticide label). Residue measurements are made on various parts of the crop at several time intervals after application to determine the rate of decline in residues of the pesticide active ingredient, its metabolites, and/or degradation products. The maximum measured residue is then used to establish the tolerance. Each crop use of a pesticide can have a different tolerance. Thus, the tolerance value is an upper bound estimate of total pesticide residue on a crop in the field, rather than in the marketplace or in table-ready foods. To obtain the TMRC, the tolerance value is multiplied by the mean U.S. food consumption estimate for each food item on which the pesticide is legally permitted, and exposures are combined for all such foods. The EPA, in calculating the TMRC, generally assumed that (l) each pesticide is used on all (100%) acres for each crop that the pesticide is permitted to be used on and (2) residues are present at the tolerance level (the highest allowable level in the field) in every food for which the pesticide is permitted. The National Food Consumption Survey conducted by the U.S. Department of Agriculture (USDA) is used for average food consumption estimates. Thus, the TMRC represents the hypothetical maximum exposure for a given pesticide (in mg/kg body weight/day) using field trial residue data. In contrast, the FDA Total Diet Study (TDS) measures detectable levels of pesticide residues as they are consumed, using a market basket survey of eight age-gender groups (Gartrell et aI., 1986; Gunderson, 1988, 1995; U.S. Food and Drug Administration, 1988, 1990, 1991a, 1992a). Market baskets of foods are collected 4 times per year, once from each of four geographic regions of the United States. Each market basket consists of 234 identical foods purchased from local supermarkets in three cities in each geographic area. The foods are selected to represent the diet of the U.S. population, prepared table-ready, homogenized together and then analyzed for pesticide residues, including some metabolites and impurities (Gartrell et aI., 1986; Gunderson, 1988, 1995; U.S. Food and Drug Administration, 1988, 1990, 1991a, 1992a). The levels of pesticide residues that are found are used in conjunction with the same USDA food consumption data used by the EPA in the TMRC in order to estimate the average dietary intake of pesticide residues in (mg/kg body weight/day) (Yess et aI., 1993). The TDS has been conducted annually by the FDA since 1961 (U.S. Food and Drug Administration, 1990), initiated primarily
in response to public concern about radionuclides in foods that might result from atmospheric nuclear testing. It is important to note that the TDS is distinct from FDA regulatory monitoring programs whose primary purpose is to ascertain that residues on crops at the "farm gate" or in the marketplace do not exceed maximum allowable levels and do not result from illegal pesticide use on crops for which the pesticide is not registered. FDA regulatory monitoring is designed only to make certain that regulations for pesticide use and application are followed, whereas the TDS is designed to provide an estimate of average daily dietary intake of pesticide residues in foods. Analytical methods for the TDS have been modified over time to permit measurement at concentrations 5 to 10 times lower than those used in FDA regulatory or incidence level monitoring. Generally, these methods can detect residues at 1 ppb (Gartrell et aI., 1986; Gunderson, 1988, 1995; U.S. Food and Drug Administration, 1988, 1990, 1991a, 1992a). 38.6.4 COMPARISON OF RISK ESTIMATES
Of the chemicals for which we were able to reproduce the EPA
qt reported by the NRC, 10 were measured in the FDA Total Diet Study, and these were used to compare risk estimates based on different exposure assessments. Our analyses of the sources of variation in cancer risk estimates for dietary synthetic pesticides are presented in Tables 38.8-38.10. A comparison of the variation in potency estimates to the variation in exposure estimates is given in Table 38.8. Table 38.9 reports hypothetical dietary exposure estimates from the NRC report, i.e., the TMRC and measured residues in the FDA TDS. In Table 38.10, risk estimates based on the TMRC are compared to risk estimates based on the TDS, using in both cases the EPA as reported by the NRC. Because of missing data or NRC results that could not be reproduced, not all chemicals are included in every table; we have used all chemicals for which appropriate data were available. TDso values were calculated from the same dose and incidence data in Table 38.6 that were used to recalculate and these TDso values are reported in Table 38.6. Table 38.8 values for the 19 compares TDso values to recalculated chemicals, using the ratio qj /(In(2)/TDso). The qj and TDso values are within a factor of 2 of each other for 10 chemicals, and within a factor of 3 for 18 chemicals. These small differences in potency estimates are within the range of differences in potency estimates from near-replicate tests where the same chemical is tested in the same sex, strain and species of test animal (Gold et aI., 1987a, 1989, 1998; Gaylor et aI., 1993). Differences in potency values are larger only for azinphosmethyl, by a factor of 6.1; there was no statistically significant increase in tumor incidence for azinphosmethyl. In contrast to the similarity of potency estimation between In(2) /TDso and ,there is enormous variation in dietary exposure estimates for synthetic pesticides between the EPA TMRC values and the FDA average dietary residues in foods prepared as consumed (Tables 38.8 and 38.9). For 5 pesticides (alachlor,
qt
qt,
qt
qt
38.6 Pesticide Residues in Food: Investigation of Disparities in Cancer Risk Estimates
819
Table 38.8 Comparison of Variation in Measures of Potency and Exposure Pesticides included in
Ratio of potency:
the TDS (FDA)
recalculated
PermethrinCq AcephateNA(Cnq)
1.5
579
0.7
1,130
ParathionCq(Cnq)
2.6
6,300
Azinphosmethy lD(E) FolpetB2
6.1
7,530
q[ /(In(2)/TDso)
0.8
Ratio of exposure: EPAlFDA
9,650"
LinuronCq(Cnq) Captan B2
2.5 1.7
16,900
ChlorothalonilNA (B2)(MOE)
1.9
99,100
Alachlo~2(MOE)
0.9
Captafol B2 CypermethrinCq(Cnq) Oxadiazon B2 (Cq)
2.2
11,600
1.2 3.0
Pesticides not measured in the TDS (FDA) AsulamNA(Cnq)
2.5
NAc
BenomylCq
2.2
NAC
Chlordimeform B2 Fosetyl AlCq(Cnq)(Unclassified)
1.7
NAC
1.8
NAC
GlyphosateCq(E)
2.5
NAC
MetolachlorCq(Cnq)(MOE) Oryzalin Cq
1.8
NAC
2.5
NAC
aFolpet was not detected by the FDA in 1984-1986. This value is for 1987. bThe FDA did not detect any residues; therefore, no ratio could be calculated. CNot applicable because not measured by the FDA. Asulam had no food uses.
Table 38.9 Dietary Exposure Estimates in 1986 by the EPA and the FDA for Pesticides Measured in the FDA Total Diet Studya Daily intake (I-lglkg/day) Pesticide
EPA TMRC (1986)
FDA TDS (1984-1986)
PermethrinCq Captan B2 FolpetB2
14.0 206 92.6 5.41 11.3 8.19 4.65 9.91 0.408 23.8 0.197 0.0938 0.486c
0.0242 0.0122
AcephateNA(Cnq) AzinphosmethyID(E) ParathionCq(Cnq) LinuronCq(Cnq) ChlorothaloniINA (B2)(MOE) Alachlo~2(MOE)
Captafol B2 CypermethrinCq(Cnq) Oxadiazon B2 (Cq) PronamideCq (B2)
0.0096 0.0048 0.0015 0.0013 0.0004 0.0001
NDb NDb NDb NDb NDb
aFDA dietary estimates are for 60--65-year-old females for 1984-1986 (Gunderson, 1995). Because of the agricultural usage of these chemicals and the prominence of fruits and vegetables in the diet of older Americans, the residues are slightly higher than for other adult age groups. bNot detected at limit of quantification (~I ppb). cDid not appear in Tables 38.1 and 38.3 because no bioassay data were available.
captafol, cypermethrin, oxadiazon and pronamide), FDA found no residues at the 1 ppb limit of quantification (Gartrell et aI., 1986; Gunderson, 1988, 1995; D.S. Food and Drug Administration, 1988, 1990, 1991a, 1992a; Yess et aI., 1993). Among chemicals detected by FDA, the TDS estimates were lower than the TMRC estimates by a factor of 99,100 for chlorothalonil, 16,900 for captan, 11,600 for linuron, and 9,650 for folpet (Table 38.8). For 4 other chemicals, the TDS estimates ranged from 579 to 7,530 times lower than TMRC. For the pesticides that EPA classified as having the strongest evidence of carcinogenicity in animal studies (B2), the differences in exposure estimates for EPA vs. FDA are particularly large (Table 38.8). Examination of FDA pesticide residue data collected over a period of 14 years (Gartrell et aI., 1986; Gunderson, 1988, 1995; D.S. Food and Drug Administration, 1988, 1990, 1991a, 1992a) indicates that dietary exposure to pesticide residues has not changed markedly over time. Thus, the large differences in exposure estimates between EPA and FDA cannot be explained simply by changes in pesticide use patterns. In standard regulatory risk assessment, an estimate of the lifetime excess cancer risk is obtained by multiplying q~ by human exposure; the true risk, however, may be zero, as the 1986 EPA cancer risk assessment guidelines indicated (D.S. Environmental Protection Agency, 1986a). A comparison of the risk estimates obtained by multiplying the q~ in the NRC report by
820
CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
Table 38.10 Comparison of Cancer Risk Estimates Based on Different Exposure Measures: TMRC Versus TDsa Cancer risk reported by NRC
Cancer risk based on TDS
Pesticideb
based on TMRC (EPA)
(FDA)
LinuronCq(Cnq)
1.5 x 10- 3
1.3 x 10- 7
Captafol B2
5.9 x 10-4 4.7 x 10-4
0
CaptanB2 PermethrinCq FolpetB2 Chlorothalonil NA (B2)(MOE) AcephateNA(Cnq)
2.8 x 10- 8 7.3 x 10- 7
4.2 x 10-4 3.2 x 10-4
3.4 x 10- 8
2.4 x 10-4
2.4 x 10- 9 3.3 x 10- 8
OxadiazonB2 (Cq)
3.7 x 10- 5 2.4 x 10-5 1.2 x 10- 5
0 0
CypermethrinCq(Cnq)
3.7 x 10- 6
0
AlachlorB 2 (MOE)
Each risk> I x 10- 6
qt
Each risk < I x 10- 6
qt
aRisk estimates use values in the NRC report for pesticides with reproducible values (see Table 38.2, column 1). EPA risks are reported in the NRC book Regulating Pesticides in Food (1987). bThree chemicals measured in the Total Diet Study (Table 38.4) are excluded: For parathion and azinphosmethyl, the values could not be reproduced; for pronamide, we were unable to obtain bioassay results.
qt
TMRC vs. TDS exposure values is presented in Table 38.10. The risks based on TMRC are also reported by NRC, and range from 10-3 to 10-6 . In contrast, risk estimates using TDS are all lower than 10-6 . There are no risk estimates in Table 38.10 for the chemicals that FDA did not detect, i.e., if there is no exposure, there is no risk. Even if the undetected chemicals are considered to be present in minute quantities, below the limit of quantification, risk estimates for these undetected chemicals would be negligible, i.e., less than 10-6 . Thus, for synthetic pesticide residues in the diet, large discrepancies in cancer risk estimates are due to differences in exposure estimates rather than to differences in carcinogenic potency values estimated by different methods from rodent bioassay data. The high risk estimates reported by NRC in 1987 were overestimates based on EPA human exposure assessments which assumed that dietary residues were at tolerance levels. For example, the TDS did not detect any residues in table-ready foods for 4 pesticides that were evaluated in the NRC report as greater than 10-6 risks (Table 38.10). 38.6.5 USE OF EXPOSURE ASSESSMENTS IN RISK ASSESSMENT The results of our analysis emphasize the importance of exposure assessment in risk estimation for synthetic pesticide residues in the diet. Both the TDS of FDA and the TMRC of EPA link estimates of food consumption patterns for groups of individuals with an estimate of pesticide concentrations in food. Since FDA and EPA use the same USDA consumption surveys to estimate dietary patterns, food consumption is not a source of variation in their exposure estimates. However, the methods of estimating the concentrations of pesticide residues in food differed markedly. The FDA measured actual residues in food
items that are bought at the market and prepared as typically eaten; the EPA used a theoretical construct, based on worst case assumptions for the maximally exposed individual and maximally allowable levels, to estimate residues that could legally occur on a given food crop at the farm gate or in the marketplace. The EPA assumption that every pesticide registered for use on a food commodity is used on every crop is another source of overestimation of exposure (Winter, 1992). In California, for example, 54 insecticides were registered for use on tomatoes in 1986; however, the maximum number of insecticides used by any tomato grower was 5, 52% of tomato growers used 2 or fewer insecticides, and 31 % used none at all (Chaisson et al., 1989). Similar findings are reported for herbicides and fungicides. FDA monitoring programs have been criticized for not measuring enough pesticides or sampling enough food items, for aggregating foods under a single representative core food (e.g., apple pie to represent all types of fruit pies), and for statistical design and sampling. In several other independent studies, however, frequency of detection and residue concentrations have also been consistently low, for example, residue data from FOODCONTAM, a national database for state surveys on pesticide and other residues in foods (Minyard and Roberts, 1991). McCarthy (1991) collected residue data on 16 pesticides for 50 crops at the "farm gate"; although all crops had been treated with the label rates of pesticide application, 93% of 134 samples had concentrations below half the tolerance. Post-harvest treatment of crops, such as removing husks or outer leaves, shelling, peeling, and washing, all reduce residue levels still further (Yess et aI., 1993), as does processing. Eilrich (1991) measured residue levels on four produce crops "from the farm gate to the table" for a fungicide whose active ingredient is
38.6 Pesticide Residues in Food: Investigation of Disparities in Cancer Risk Estimates chlorothalonil and found that dietary residues were similar to those reported by the FDA. Analyses by Nigg et al. (1990) and Winter (1992) of residue data from the California Department of Food and Agriculture confirm the FDA regulatory monitoring findings. Most crops have no detectable residues; crop residues that are found are small fractions of tolerance values. Thus, tolerances are poor indicators of human exposure, a function for which they were not designed. Although it is possible that a small percentage of people who obtain food crops close to the farm gate may have higher incidental dietary exposures, these concentrations are very unlikely to persist over time and would still be substantially lower than the TMRC values. In the TDS, approximately 264 pesticides, metabolites, and impurities are analyzed; only 51 had detectable residues, and only 3 were present in more than 10% of the sample foods (V.S. Food and Drug Administration, 1991a). These findings are similar to those obtained from the TDS during the 10 previous years (Gartrell et aI., 1986; Gunderson, 1988; V.S. Food and Drug Administration, 1988, 1990, 1991a, 1992a) and to those from surveys on pesticides of special interest. Even if exposure estimates based on the TDS were underestimates by an order of magnitude, the potential risks estimated using the EPA qi would still be low. The use of the TMRC as an estimate of human dietary exposure in quantitative cancer risk assessment is not justified, from either a scientific or a public policy perspective, because this measure often grossly exaggerates actual consumer exposure. The TMRC uses tolerances as surrogates for concentration in foods and therefore, by definition, the TMRC is not representative of the level likely to reach the consumer (Chaisson et aI., 1989). It does not take into account percentage of crop treated, actual pesticide application practices, chemical degradation from farm gate to table, and cooking or other processing. Some subsequent EPA exposure estimates have used "anticipated residues" instead of TMRC, which are calculated using tolerances and processing factors, using tolerances and percentage of crop treated, using field trial data, or using monitoring data. The anticipated residue also tends to be an overestimate because it is based on the average residue observed from maximum allowable pesticide application of a pesticide during field trials. Actual pesticide use is not always at the maximum level; hence, actual residues tend to be lower than the anticipated level (Chaisson et aI., 1989). For example, the EPA subsequently used anticipated residues to evaluate linuron and reported that less than 1% of the crop of barley, oats, and rye was treated. Despite this finding, for risk assessment purposes the EPA assumed that 100% of the crop was treated. The linuron comparison indicates how anticipated residues can be an overestimate: The TMRC in the NRC report was 4.65 f..Lg/kg/day; the anticipated residue reported by EPA was 0.185 f..Lg/kg/day (U.S. Environmental Protection Agency, 1995b); the TDS value was 0.0004 f..Lg/kg/day (Gunderson, 1995). Recent developments by government agencies have responded to the need for better quality information on exposure assessment of dietary residues. In response to the need identi-
821
fied by the National Academy of Sciences (NAS) for a standardized exposure database while developing the report Pesticides in the Diets of Infants and Children, the EPA has begun a National Pesticide Residue Database (NPRD) that collects data from the FDA, the VSDA, and private and commercial sources (http://www.epa.gov/pesticides/nprd). A mUltiagency effort, the Pesticide Data Program (PDP), is providing more information on actual exposure to dietary residues, food consumption, and pesticide usage (V.S. Environmental Protection Agency, 1999a). The PDP was established by the VSDA in 1991 to monitor pesticide residues in fresh and processed fruits and vegetables at terminal markets or distribution centers. Sampling procedures are designed to measure residues close to the time of consumption. Since 1994, the PDP testing protocol has included several foods in addition to fresh produce, such as canned and frozen fruits and vegetables and milk. The PDP is a critical component of the Food Quality Protection Act of 1996, and hence focuses on commodities that are consumed by infants and children (http://www. ams.usda.gov/science/pdp/what.htm). In 1998, PDP produce samples originated from 40 states and 25 foreign countries (V.S. Department of Agriculture, 2000). The PDP is currently used by the EPA to support its dietary risk assessment process [e.g., Eiden (1999)] and by the FDA to refine sampling for enforcement of tolerances. Given that exposure assessments for pesticide residues are available from the FDA TDS for about 38 years, it might be reasonable to compare those assessments to the new PDP assessments. A more complete characterization of exposures has been undertaken for some chemicals using biomarkers of exposure or distributions of exposure factors. Monte Carlo methods and other variance propagation techniques have been used to characterize the interindividual variability in exposures within a population and the uncertainty in exposure estimates (McKone, 1997). 38.6.6 USE OF TOXICOLOGICAL DATA IN RISK ASSESSMENT
Throughout this chapter, we have presented data indicating the limitations of tumor incidence results from rodent cancer tests in efforts to estimate human risk at low exposures. Our analysis of differences in risk estimates for dietary pesticide residues indicated that carcinogenic potency values were similar for In(2)jTDso and qi and therefore did not contribute substantially to the disparities in risk estimation. Similarity in potency estimates is expected: Bernstein et al. (1985) showed that carcinogenic potency values from standard bioassays are restricted to an approximately 32-fold range surrounding the maximum dose tested, in the absence of 100% tumor incidence. Estimates of carcinogenic potency derived from statistical models are highly correlated with one another because they are all highly correlated with the MTD, regardless of whether the estimate is based on the one-stage, multi stage, or Weibull model (Krewski et aI., 1990). This constraint on potency estimation
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CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
contrasts with the enormous extrapolation that is required from the MTD in bioassays to the usual human exposure levels of pesticide residues, often hundreds of thousands of times lower than the MTD. One implication of the boundedness of potency estimation based on tumor incidence data is that for a given exposure estimate, the risk estimate can be approximated from the MTD in the bioassay without conducting an experiment. We have shown that the VSD at 10-6 (Gaylor and Gold, 1995) and the risk estimate based on the LTDlO, whether using safety factors for a nonlinear dose-response relationship or a linear model, can all be approximated from the MTD within a factor of 10 of the estimate that would be obtained from tumor incidence data in standard bioassays (Gaylor and Gold, 1998) (see Section 38.4). Adequate qualitative evaluation of the weight of evidence for carcinogenicity of a chemical and quantitative extrapolation from high to low dose requires more information for a chemical, about pharmacokinetics, mechanism of action, cell division, induction of defense and repair systems, and species differences. The new EPA guidelines and recent evaluations of several chemicals by the EPA, recognize the importance of such additional information (U.S. Environmental Protection Agency, 1996a). The proposed EPA guidelines permit the use of nonlinear approaches to low-dose extrapolation if warranted by mechanistic data. In recent years, the EPA has reevaluated several of the weight-of-evidence classifications for pesticides in the NRC report. This is consistent with the recommendation in the proposed EPA cancer risk assessment guidelines, which calls for use of available toxicological data in a characterization of the weight of evidence. Several pesticides are no longer considered appropriate for quantitative risk estimation (see the superscripts in parentheses in Table 38.6). Of the 19 pesticides from the NRC report for which we obtained bioassay data from EPA, only 11 are currently considered by the EPA as appropriate for quantitative risk estimation (Table 38.6 superscripts). This contrasts with the NRC report evaluation that the risks for 16 of the 19 were greater than 10-6 . For example, 1inuron had the highest risk estimate of all pesticides in the NRC analysis. It was subsequently reclassified by the EPA as inappropriate for quantitative risk assessment based on biological considerations: The testicular tumors in rats were late forming and benign and were a relatively common tumor type; the hepatocellular tumors in mice were benign and only in the highest dose group; and there was no evidence of mutagenic activity (U.S. Environmental Protection Agency, 1988a, 1999b). For evaluation of the mode of action of a given chemical using the new EPA risk assessment guidelines, information other than bioassay data can be developed and included in the assessment of weight of evidence and whether the dose-response relationship is likely to be nonlinear; for example, pharmacokinetic data on absorption, distribution, and metabolism can be used to predict target organ concentrations and then compared in different species. Other relevant results can be obtained from studies of cell division at and below the carcinogenic dose or from receptor-binding assays. New animal models with genetic alterations that are designed to make an animal resemble the
human more closely or to make the animal more sensitive to a given response can complement or take the place of long-term cancer tests, for example, transgenic mouse models that use unique phenotypic properties such as the p53 gene-deficient model or receptor-binding assays (Blaauboer et aI., 1998). Critical evaluation and validation of these new methodologies and increasing use of fundamental toxicological research will improve the regulatory evaluation of potential human risk. Although the proposed guidelines offer some incentive to generate mechanistic data on a chemical, for most chemicals no such data will be available, and the default procedure will continue to be used. If bioassay data are to be used in risk assessment, it is desirable to facilitate generation of mechanistic data on the chemicals of interest (Clayson and Iverson, 1996), including chemicals for which past risk assessments have resulted in regulation.
38.7 RANKING POSSIBLE TOXIC HAZARDS FROM NATURALLY OCCURRING CHEMICALS IN THE DIET Because naturally occurring chemicals in the diet have not been a focus of cancer research, it seems reasonable to investigate some of them further as possible hazards because they often occur at high concentrations in common foods. Only a small proportion of the many chemicals to which humans are exposed will ever be investigated, and there is at least some toxicological plausibility that high-dose exposures may be important. Moreover, the proportion positive in rodent cancer tests is similar for natural and synthetic chemicals, about 50% (see Section 38.3), and the proportion positive among natural plant pesticides is also similar (Table 38.3). Therefore, one would expect many of the untested natural chemicals to be rodent carcinogens. In order to identify and prioritize untested dietary chemicals that might be a hazard to humans if they were to be identified as rodent carcinogens, we have used an index, HERT, which is analogous to HERP (see Section 38.5). HERT is the ratio of human exposure/rodent toxicity (LDso) in mg/kg/day expressed as a percentage, whereas HERP is the ratio of human exposure/rodent carcinogenic potency (in mg/kg/day) expressed as a percentage. HERT uses readily available LDso values rather than the TDso values from animal cancer tests that are used in HERP. This approach to prioritizing untested chemicals makes assessment of human exposure levels critical at the outset. The validity of the HERT approach is supported by three analyses: First, we have found that for the exposures to rodent carcinogens for which we have calculated HERP values (Gold et aI., 1992), the rankings by HERP and HERT are highly correlated (Spearman rank order correlation = 0.89). Second, we have shown that without conducting a 2-year bioassay the regulatory VSD can be approximated by dividing the MTD by 740,000 (Gaylor and Gold, 1995; and Section 38.4). Because the MTD is not known for all chemicals and the MTD and LDso are both measures of toxicity, acute toxicity (LDso) can reasonably be used as a surrogate for chronic toxicity (MTD).
38.7 Ranking Possible Toxic Hazards from Naturally Occurring Chemicals in the Diet
Third, LDso and carcinogenic potency are correlated (Travis et aI., 1990; Zeise et aI., 1984); therefore, HERT is a reasonable surrogate index for HERP because it simply replaces TDso with LDso. We have calculated HERT values using LDso values as a measure of toxicity and human exposure estimates based on the available data on concentrations of untested natural chemicals in commonly consumed foods and average consumption of those foods in the U.S. diet. Literature searches identified the most commonly consumed foods (Stofberg and Grundschober, 1987; Technical Assessment Systems 1989; United Fresh Fruit and Vegetable Association, 1989) and concentrations of chemicals in those foods (Nijssen et aI., 1996; U.S. National Institute for Occupational Safety and Health, 1999). We considered any chemical with available data on rodent LDso that had a published concentration of ~ 10 ppm in a common food and for which estimates of average U.S. consumption of that food were available. The natural pesticides among the chemicals in the HERT table (Table 38.11) are marked with an asterisk. Among the set of 121 HERT values (Table 38.11), the HERT ranged across 6 orders of magnitude. The median HERT value for average dietary exposures is 0.007%. It might be reasonable to investigate further the chemicals in the diet that rank highest on the HERT index and that have not been adequately tested in chronic carcinogenicity bioassays in rats and mice. We have nominated to the NTP the chemicals with the highest HERT values as candidates for carcinogenicity testing. These include solanine and chaconine, the main alkaloids in potatoes, which are cholinesterase inhibitors that can be detected in the blood of almost all people (Ames, 1983, 1984; Harvey et aI., 1985); chlorogenic acid, a precursor of caffeic acid; and caffeine, for which no adequate standard lifetime study has been conducted in mice. In rats, cancer tests of caffeine have been negative, but one study that was inadequate because of early mortality, showed an increase in pituitary adenomas (Yamagami et aI., 1983). How would the synthetic pesticides that are rodent carcinogens and that are included in the HERP ranking (Table 38.5) compare to the natural chemicals that have not been tested for carcinogenicity (Table 38.11) if they too were ranked on HERT? We calculated HERT using LDso values for the synthetic pesticide residues that are rodent carcinogens in the HERP table and found that they rank low in HERT compared to the naturally occurring chemicals in Table 38.11; 88% (1071121) of the HERT values for the natural chemicals in Table 38.11 rank higher in possible toxic hazard HERT than any HERT value for any synthetic pesticide that is a rodent carcinogen in the HERP table (Table 38.5). The highest HERT for the synthetic pesticides would be for DDT in 1970 before the ban (0.00004%), which is more than lOO-fold lower than the median HERT for the natural chemicals in the HERT table. Many interesting natural toxicants are ranked in common foods in the HERT table. Oxalic acid, a plant pesticide, which is one of the most frequent chemicals in the table, occurs widely in nature. It is usually present as the potassium or calcium salt and also occurs as the free acid (Hodgkinson, 1977). Oxalic acid
823
is reported in many foods in Table 38.11; the highest contributors to the average U.S. diet are coffee (HERT = 0.09%), carrot (0.08%), tea (0.02%), chocolate (0.01 %), and tomato (0.01 %). Excessive consumption of oxalate has been associated with urinary tract calculi and reduced absorption of calcium in humans (Beier and Nigg, 1994; Hodgkinson, 1977). Because of the high concentrations of natural pesticides in spices, we have reported the HERT values for average intake in Table 38.11, even though spices are not among the foods consumed in the greatest amounts by weight. The highest concentrations of chemicals in Table 38.11 are found in spices, which tend to have higher concentrations of fewer chemicals (Nijssen et aI., 1996). (Concentrations can be derived from Table 38.11 by the ratio of the average consumption of the chemical and the average consumption of the food.) High concentrations of natural pesticides in spices include those for menthone in peppermint oil (243,000 ppm), y-terpinene in lemon oil (85,100 ppm), citral in lemon oil (75,000 ppm) piperine in black pepper (47,100 ppm), and geranial in lemon juice (14,400 ppm) and lemon oil (11,300 ppm). Natural pesticides in spices have antibacterial and antifungal activities (Billing and Sherman, 1998) whose potency varies by spice. A recent study of recipes in 36 countries examined the hypothesis that spices are used to inhibit or kill food spoilage microorganisms. Results indicate that as mean annual temperature increases in a geographical area (and therefore so does spoilage potential), there is an increase in number of spices used and use of the spices that have greatest antimicrobial effectiveness. The authors argue that spices are used to enhance food flavor, but, ultimately, are continued in use because they help to eliminate pathogens and therefore contribute to health, reproductive success, and longevity (Billing and Sherman, 1998). Cyanogenesis, the ability to release hydrogen cyanide, is widespread in plants, including several foods, of which the most widely eaten globally are cassava and lima bean (Poulton, 1983). Cassava is consumed widely throughout the tropics and is a dietary staple for over 300 million people (Bokanga et aI., 1994). There are few effective means of removing the cyanogenic glycosides that produce hydrogen cyanide (HCN), and cooking is generally not effective (Bokanga et aI., 1994; Poulton, 1983). For lima beans in Table 38.6, the HERT is 0.01 %. Ground flaxseed, a dietary supplement (http://www.heintzmanfarms.comJ; Gruenwald et aI., 1998), contains about 500 ppm hydrogen cyanide glycosides. The HCN in flaxseed appear to be inactivated in the digestive tract of primates (Mazza and Oomah, 1995). The increasing popUlarity of herbal supplements in the United States raises concerns about possible adverse effects from high doses or drug interactions (Saxe, 1987). Because the recommended doses of herbal supplements are close to the toxic dose and because about half of natural chemicals are rodent carcinogens in standard animal cancer tests, it is likely that many dietary supplements from plants will be rodent carcinogens that would rank high in possible carcinogenic hazard (HERP) if they were tested for carcinogenicity. Whereas pharmaceuticals are federally regulated for purity, identifica-
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CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
Table 38.11 Ranking Possible Toxic Hazards to Naturally Occurring Chemicals in Food on the HERT (Human ExposurelRodent Toxicity) Index Possible hazard: HERT
Average human
LDSO (mg/kg)
(%)
Average daily consumption of food
consumption of chemical
Rats
Mice
Exposure references
4.3
Coffee, 500 ml (13.3 g)
*Caffeine, 381 mg
(192)
127
Stofberg and Grundschober (1987), Macaulay et al. (1984), IARC (1991)
0.3
Tea, 60.2 ml (903 mg)
*Caffeine, 29.4 mg
(192)
127
Stofberg and Grundschober (1987), Martinek and Wolman (1955), Wolman (1955), Lee (1973), Groisser (1978), Bunker and McWilliams (1979), Galasko et al. (1989), IARC (1991)
0.3
Potato, 54.9 g
*a-Chaconine, 4.10 mg
(84P)
19P
TAS (1989), Bushway and Ponnampalam (1981), Takagi et al. (1990)
0.2
Cola, 174 ml
*Caffeine, 20.8 mg
(192)
127
EPA (1996b), Bunker and McWilliams (1979), Galasko et al. (1989)
0.1
Coffee, 500 ml
*Chlorogenic acid, 274 mg
4000P
Stofberg and Grundschober (1987), Baltes (1977), IARC (1991)
0.09
Coffee, 500 ml
*Oxalic acid, 25.2 mg
382
Stofberg and Grundschober (1987), Kasidas and Rose (1980), IARC (1991), Vernot et al. (1977)
0.09
Black pepper, 446 mg
*Piperine, 21.0 mg
(514)
0.08
Carrot, boiled, 12.1 g
*Oxalic acid, 22.7 mg
382
0.08
Chocolate (cocoa solids) 3.34 g
*Theobromine, 48.8 mg
(1265)
0.05
Lemon juice, 1.33 ml
*Geranial, 19.2 mg
500
EPA (1996b), Mussinan et al. (1981)
0.05
Coffee, 500 ml
*Trigonelline, 176 mg
5000
Stofberg and Grundschober (1987), Clinton (1986), IARC (1991)
0.03
Chocolate (cocoa solids) 3.34 g
*Caffeine, 2.30 mg
(192)
0.02
Tea, 60.2 ml
*Oxalic acid, 6.67 mg
382
0.G2
Isoamyl alcohol: U.S. avg (mostly
Isoamyl alcohol, 18.4 mg
1300
330
Stofberg and Grundschober (1987) TAS (1989), Zarembski and Hodgkinson (1962), Vernot et al. (1977)
837
127
Stofberg and Grundschober (1987), IARC (1991)
Stofberg and Grundschober (1987), Zoumas et al. (1980) Stofberg and Grundschober (1987), Zarembski and Hodgkinson (1962), Kasidas and Rose (1980), IARC (1991), Vernot et al. (1977) Stofberg and Grundschober (1987)
beer, alcoholic beverages) 0.01
Beer, 257 m1
Isoamyl alcohol, 13.6 mg
1300
Stofberg and Grundschober (1987), Arkima (1968)
0.01
Chocolate (cocoa solids) 3.34 g
*Oxalic acid, 3.91 mg
382
Stofberg and Grundschober (1987), Kasidas and Rose (1980), Vernot et al. (1977)
0.01
Tomato, 88.7 g
*Oxalic acid, 3.24 mg
382
Stofberg and Grundschober (1987), Zarembski and Hodgkinson (1962), Kasidas and Rose (1980), Vernot et al. (1977)
0.01
Coffee, 500 ml
2-Furancarboxylic acid, 821 J.!g
0.01
Lima beans, 559 mg
Hydrogen cyanide, 28.5 J.!g
0.01
Potato chips, 5.2 g
*a-Chaconine, 136 J.!g"
0.01
Sweet potato, 7.67 g
*Ipomeamarone, 336 J.!g
0.009
Potato, 54.9 g
*a-Solanine, 3.68 mg
590
0.008
Isobutyl alcohol: U.S. avg
Isobutyl alcohol, 14.1 mg
2460
0.008
Hexanoic acid: U.S. avg
Hexanoic acid, 15.8 mg
3000
0.007
Phenethyl alcohol: U.S. avg
Phenethyl alcohol, 8.28 mg
1790
0.007
Carrot, 12.1 g
*Carotatoxin, 460 J.!g
0.006
Ethyl acetate: U.S. avg
Ethyl acetate, 16.5 mg
lOOP
(84P)
Stofberg and Grundschober (1987), Tressl et al. (1978), IARC (1991), Kitamura et al. (1978)
3.7
EPA (1996b), Viehoever (1940), Montgomery (1964)
19P
Stofberg and Grundschober (1987), Friedman and Dao (1990)
50
Stofberg and Grundschober (1987), Coxon et al. (1975) TAS (1989), Bushway and Ponnampalam (1981), Takagi et al. (1990) Stofberg and Grundschober (1987)
(5000)
Stofberg and Grundschober (1987)
1001
Crosby and Aharonson (1967), Wulf et al. (1978)
4100
Stofberg and Grundschober (1987)
(beer, grapes, wine)
(5620)
Stofberg and Grundschober (1987)
(mostly alcoholic beverages)
(continues)
38.7 Ranking Possible Toxic Hazards from Naturally Occurring Chemicals in the Diet
825
Table 38.11 (continued)
Possible hazard:
HERT
Average human
(%)
Average daily consumption of food
consumption of chemical
LDsO (mglkg) Rats Mice
0.005
Celery, 7.95 g
*Oxalic acid, 1.39 mg
382
0.005
Coffee, 500 m1
*3·Methylcatechol, 203 flg
0.005
Potato, 54.9 g
*Oxalic acid, 1.26 mg
Exposure references ERS (1994), Zarembski and Hodgkinson (1962), Vernot et al. (1977)
56V 382
Stofberg and Grundschober (1987), Heinrich and Baltes (1987), IARC (1991) TAS (1989), Zarembski and Hodgkinson (1962), Vernot et al. (1977)
0.004
Beer, 257 ml
Phenethy1 alcohol, 5.46 mg
1790
Stofberg and Grundschober (1987), Arkima (1968)
0.004
Corn, 33.8 g
*Oxalic acid, 1.12 mg
382
Stofberg and Grundschober (1987), Kohman (1939), Vernot et al. (1977)
0.004
Corn, 33.8 g
Methylamine, 906 flg
0.004
Peppermint oil, 5.48 mg
*Menthone, 1.33 mg
0.004
White bread, 67.6 g
Propionaldehyde, 2.09 mg
0.004
Beer, 257 ml
Isobutyl alcohol, 6.40 mg
2460
0.003
Tomato, 88.7 g
Methyl alcohol, 13.4 mg
5628
0.003
Wine, 28.0 ml
Isoamyl alcohol, 3.00 mg
1300
0.003
Coffee, 500 ml
Pyrogallol, 555 flg
0.003
Apple, 32.0 g
*Oxalic acid, 704 flg
382
EPA (1989a), Zarembski and Hodgkinson (1962), Kasidas and Rose (1980), Vernot et al. (1977)
0.003
Butyl alcohol: V.S. avg
Butyl alcohol, 1.45 mg
790
Stofberg and Grundschober (1987)
500 (1410)
317
Stofberg and Grundschober (1987), Neurath et al. (1977)
800
Stofberg and Grundschober (1987), Lorenz and Maga (1972)
(7300)
TAS (1989), Nelson and Hoff (1969), Kazeniac and Hall (1970)
300
Stofberg and Grundschober (1987), Tressl et al. (1978), IARC (1991)
Stofberg and Grundschober (1987)
Stofberg and Grundschober (1987), Arkima (1968)
Stofberg and Grundschober (1987), Postel et al. (1972)
(mostly apple, beer) 317
TAS (1989), Neurath et at. (1977)
Lettuce, 14.9 g Beer, 257 ml
Methylamine, 567 flg
0.003
Propyl alcohol, 3.29 mg
1870
(6800)
Stofberg and Grundschober (1987), Arkima (1968)
0.002
Banana, 15.7 g
trans- 2-Hexenal, 1.19 mg
(780)
685
TAS (1989), Hultin and Proctor (1961)
0.002
Orange, 10.5 g
*Oxalic acid, 651 flg
382
0.002
Wine, 28.0 ml
Ethyl lactate, 4.16 mg
(>5000)
0.002 0.002
Tomato, 88.7 g White bread, 67.6 g
* p-Coumaric acid, Butanal, 3.44 mg
2490
0.002
Tea, 60.2 ml
*Theobromine, 1.11 mg
0.002 0.002
Apple, 32.0 g Tomato, 88.7 g
0.002 0.002
0.003
TAS (1989), Zarembski and Hodgkinson (1962), Vernot et al. (1977) 2500
Stofberg and Grundschober (1987), Postel et al. (1972), Shinohara et at. (1979)
657P
TAS (1989), Schmidtlein and Herrmann (1975a) Stofberg and Grundschober (1987), Lorenz and Maga (1972), Smyth et al. (1951)
(1265)
837
*Epicatechin, 1.28 mg
1000P
Stofberg and Grundschober (1987), Blauch and Tarka (1983), Nagata and Sakai (1985), IARC (1991) EPA (1989a), Risch and Herrmann (1988)
Beer, 257 ml
*Tomatine, 621 flg Ethyl acetate, 4.42 mg
(5620)
500 4100
Lettuce, 14.9 g
*Oxalic acid, 447 flg
382
0.001
Apple, 32.0 g
*p-Coumaric acid, 573 flg
0.001
Apple, 32.0 g
*Chlorogenic acid, 3.39 mg
4000P
0.001
Coffee, 500 ml
Maltol, 462 flg
(1410)
0.001
Coffee, 500 ml
Nonanoic acid, 188 flg
1.02 mg
TAS (1989), Eltayeb and Roddick (1984) Stofberg and Grundschober (1987), Rosculet and Rickard (1968) TAS (1989), Kasidas and Rose (1980), Vernot et al. (1977)
657P
EPA (1989a), Mosel and Herrmann (1974) EPA (1989a), Jurics (1967), Perez-Ilzarbe et al. (1991)
550
Stofberg and Grundschober (1987), Tressl et al. (1978), IARC (1991)
224V
Stofberg and Grundschober (1987), Kung et al. (1967), IARC (1991) (continues)
826
CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
Table 38.11 (continued) Possible hazard: HERT
Average human
LDSO (mg/kg)
(%)
Average daily consumption of food
consumption of chemical
Rats
0.001
5-Methylfurfural: U.S. avg
5-Methylfurfural, 1.71 mg
2200
Stotberg and Grundschober (1987)
*fi-Pinene, 3.28 mg
4700
Stotberg and Grundschober (1987)
Mice
Exposure references
(mostly coffee) 0.001
fi-Pinene: U.S. avg (mostly pepper, lemon oil, nutmeg)
0.001
Broccoli, 6.71 g
*Oxalic acid, 268 fig
382
ERS (1994), Kohman (1939), Vernot et al. (1977)
0.001
Strawberry, 4.38 g
*Oxalic acid, 261 fig
382
Stotberg and Grundschober (1987), Zarembski and Hodgkinson (1962), Kasidas and Rose (1980), Vernot et al. (1977)
0.0009
Orange juice, 138 ml
Methyl alcohol, 3.48 mg
5628
0.0009
a-Pinene: U.S. avg (mostly pepper,
*a-Pinene, 2.25 mg
3700
(7300)
TAS (1989), Kirchner and Miller (1957), Tanner and Limacher (1984), Nisperos-Carriedo and Shaw (1990) Stotberg and Grundschober (1987)
nutmeg, lemon oil) 0.0009
White bread, 67.6 g
2-Butanone, 1.65 mg
2737
(4050)
Stotberg and Grundschober (1987), Lorenz and Maga (1972)
0.0008
Coffee, 500 ml
Pyridine, 5 I 9 fig
891
(1500)
Stotberg and Grundschober (1987), Silwar et al. (1987), IARC (1991)
0.0008
Acetone: U.S. avg
Acetone, 1.74 mg
(5800)
3000
Stotberg and Grundschober (1987)
(698)
316
Stotberg and Grundschober (1987), Neurath et al. (1977)
317
Stotberg and Grundschober (1987), Neurath et al. (1977)
(mostly tomato, bread, beer) 0.0008
Cucumber, pickled, 11.8 g
Dimethylamine, 182 fig
0.0008
Cabbage, raw, 12.9 g
Methylamine, 169 fig
0.0007
Tomato, 88.7 g
*Chlorogenic acid, 2.06 mg
4000P
0.0007
Wine, 28.0 ml
Methyl alcohol, 2.84 ml
5628
0.0007
Coffee, 500 ml
2-Methylpyrazine, 894 fig
1800
Stotberg and Grundschober (1987), Silwar et al. (1987), IARC (1991)
0.0007
Coffee, 500 ml
2,6-Dimethylpyrazine, 432 fig
880
Stotberg and Grundschober (1987), Silwar et al. (1987), IARC (1991)
0.0007
Cabbage, raw, green, 12.9 g
*p-Coumaric acid, 303 fig
0.0006
Peach, 9.58 g
*Chlorogenic acid, 1.78 mg
4000P
Stotberg and Grundschober (1987), Jurics (1967), MolIer and Herrmann (1983), Senter et al. (1989)
0.0006
Black pepper, 446 mg
*3-Carene, 2.00 mg
4800
Stotberg and Grundschober (1987), Pino et al. (1990)
0.0006
Cabbage, boiled, 12.9 g
*OxaIic acid, 155 fig
382
Stotberg and Grundschober (1987), Zarembski and Hodgkinson (1962), Vernot et al. (1977)
0.0006
Coffee, 500 ml
Butyric acid, 785 fig
2000
Stotberg and Grundschober (1987), Kung et al. (1967), IARC (1991)
0.0006
Coffee, 500 ml
2,5-Dimethylpyrazine, 399 fig
1020
Stotberg and Grundschober (1987), Silwar et al. (1987), IARC (1991)
0.0005
Coffee, 500 ml
5-Methylfurfural, 798 fig
2200
Stotberg and Grundschober (1987), Silwar et al. (1987), IARC (1991)
0.0005
Grapes, I I g
*Oxalic acid, 138 fig
382
Stotberg and Grundschober (1987), Kohman (1939), Vernot et al. (1977)
0.0005
Grapes, I I g
*Chlorogenic acid, 1.38 mg
4000P
Stotberg and Grundschober (1987), Jurics (1967)
0.0005
Black pepper, 446 mg
*fi-Pinene, 1.50 mg
4700
Stotberg and Grundschober (1987), Pino et al. (1990)
TAS (1989), Winter and Herrmann (1986) (7300)
657P
Stotberg and Grundschober (1987), Postel et al. (1972)
Stotberg and Grundschober (1987), Schmidtlein and Herrmann (1975b)
0.0004
Cucumber (raw flesh), 11.8 g
*Oxalic acid, 118 fig
382
Stotberg and Grundschober (1987), Kasidas and Rose (1980), Vernot et al. (1977)
0.0004
Potato chips, 5.2 g
*a-Solanine, 179 fig
590
Stotberg and Grundschober (1987), Ahmed and MUlIer (1978)
0.0004
Coffee, 500 ml
Propanoic acid, 785 fig
2600
Stotberg and Grundschober (1987), Kung et al. (1967), IARC (1991)
(continues)
38.7 Ranking Possible Toxic Hazards from Naturally Occurring Chemicals in the Diet
827
Table 38.11 (continued) Possible hazard: Average human
HERT
LDSO (mglkg)
(0/0)
Average daily consumption of food
consumption of chemical
Rats
0.0004
Peach, canned, 9.58 g
*Oxalic acid, 115 fl,g
382
0.0004
Lettuce, 14.9 g
Benzylamine, 172 fl,g
0.0004
Lemon juice, 1.33 ml
Octanal, 1.60 mg
5630
EPA (1996b), Mussinan et al. (1981)
0.0004
a-Phellandrene: D.S. avg
*a-Phellandrene, 1.59 mg
5700
Stofberg and Grundschober (1987)
Hexanal, 1.35 mg
4890
Mice
Exposure references Stofberg and Grundschober (1987), Zarembski and Hodgkinson (1962), Vernot et al. (1977)
600P
TAS (1989), Neurath et al. (1977)
(mostly pepper) 0.0004
White bread, 67.6 g
0.0004
Black pepper, 446 mg
*a-Pinene, 1.02 mg
3700
0.0004
Banana, 15.7 g
2-Pentanone, 424 fl,g
1600
0.0003
Grapes, 11 g
*Epicatechin, 243 fl,g
0.0003
Onion, raw, 14.2 g
Dipropyl trisulfide, 189 fl,g
0.0003
Coffee. 500 ml
2-Ethyl-3-methylpyrazine,
(8292)
Stofberg and Grundschober (1987), Lorenz and Maga (1972)
1600
TAS (1989), Hultin and Proctor (1961)
l000P
Stofberg and Grundschober (1987), Jurics (1967), Lee and Jaworski (1987)
800
Stofberg and Grundschober (1987)
Stofberg and Grundschober (1987), Pino et al. (1990)
Stofberg and Grundschober (1987), IARC (1991)
880
186 fl,g 0.0003
Pear, 3.29 g
*Chlorogenic acid, 823 fl,g
4000P
Stofberg and Grundschober (1987), J urics (1967)
0.0003
Carrot, 12.1 g
*Chlorogenic acid, 780 fl,g
4000P
TAS (1989), Winter et al. (1987)
0.0003
Lemon oil, 8 mg
*y-Terpinene, 681 fl,g
3650
Stofberg and Grundschober (1987), Ikeda et al. (1962), Staroscik and Wilson (1982a, 1982b)
0.0003
Lemon oil, 8 mg
*Geranial, 90.4 fl,g
500
Stofberg and Grundschober (1987), Bernhard (1960), Staroscik and Wilson (1982a, 1982b)
0.0003
Lemon oil, 8 mg
*,B-Pinene, 832 fl,g
4700
Stofberg and Grundschober (1987), Ikeda et al. (1962), Staroscik and Wilson (1982a, 1982b)
0.0002
Broccoli (raw), 6.71 g
*p-Coumaric acid, 90.6 fl,g
0.0002
Lemon oil, 8 mg
*Citral, 600 fl,g
4960
0.0001
Isoamyl acetate: D.S. avg
Isoamyl acetate, 1.70 mg
16,600
Dimethyl sulfide, 324 fl,g
3300
657P
ERS (1994), Schmidtlein and Herrmann (l975b)
(6000)
Stofberg and Grundschober (1987), Giinther (1968) Stofberg and Grundschober (1987)
(mostly beer, banana) 0.0001
Corn, canned, 33.8 g
0.0001
Onions, green, cooked, 137 mg
*Oxalic acid, 31.5 fl,g
382
0.0001
Coffee, 500 ml
Hexanoic acid, 245 fl,g
3000
0.0001
Pear, 3.29 g
*Epicatechin, 80.9 fl,g
(3700)
Stofberg and Grundschober (1987), Williams et al. (1972), Buttery et al. (1994)
(5000)
Stofberg and Grundschober (1987), Kung et al. (1967), IARC (1991) Stofberg and Grundschober (1987), Mosel and Herrmann (1974), Risch and Herrmann (1988)
EPA (1996b), Kohman (1939), Vernot et al. (1977)
1000P
0.00007
Nutmeg, 27.4 mg
*Myristicin, 207 fl,g
4260
0.00006
Banana, 15.7 g
Methyl alcohol, 236 fl,g
5628
0.00005
Lemon oil, 8 mg
*a-Pinene, 139 fl,g
3700
0.00005
Banana, 15.7 g
Isoamyl acetate, 584 fl,g
16,600
TAS (1989), Tressl et al. (1970)
0.00005
Strawberry, 4.38 g
*Chlorogenic acid, 136 fl,g
4000P
Stofberg and Grundschober (1987), Jurics (1967)
0.00004
Black pepper, 446 mg
*a-Phellandrene, 162 fl,g
5700
0.00002
Grapefruit juice, 3.29 ml
Methyl alcohol, 95.4 fl,g
5628
0.00002
Lemon oil, 8 mg
*a-Terpinene, 23.2 fl,g
1680
0.00001
Lemon oil, 8 mg
*a-Terpineol, 29.6 fl,g
Ehlers et al. (1998) (7300)
TAS (1989), Hultin and Proctor (1961) Stofberg and Grundschober (1987), Ikeda et al. (1962), Staroscik and Wilson (1982a, 1982b)
Stofberg and Grundschober (1987), Pino et al. (1990) (7300)
Stofberg and Grundschober (1987), Kirchner et al. (1953), Lund et al. (1981), Tanner and Limacher (1984), Pino et al. (1986) Stofberg and Grundschober (1987), Staroscik and Wilson (1982a, 1982b)
2830
Stofberg and Grundschober (1987), Staroscik and Wilson (1982a, 1982b) (continues)
828
CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
Table 38.11 (continued) Possible hazard: HERT (%)
Average daily consumption of food
Average human
LDso (mg/kg)
consumption of chemical
Rats
Mice
Exposure references
2830
Stofberg and Grundschober (1987), Pino et al. (1990)
0.00001
Black pepper, 446 mg
*a-Terpineol, 25.0 J.4g
0.00001
Garlic, blanched, 53.3 mg
Diallyl disulfide, 2.05 J.4g
260
EPA (1996b), Yu et al. (1989)
0.00001
Lemon oil, 8 mg
*Terpinolene, 29.6 J.4g
4390
Stofberg and Grundschober (1987), Staroscik and Wilson (1982a, 1982b)
0.000008
Garlic, blanched, 53.3 mg
Diallyl trisulfide, 592 ng
0.000001
Garlic, blanched, 53.3 mg
Diallyl sulfide, 2.28 J.4g
100 2980
EPA (1996b), Yu et al. (1989) EPA (1996b), Yu et al. (1989)
LDSO: Values are from the Registry of Toxic Effects of Chemical Substances (RTECS). Parentheses indicate the species with the higher (weaker) LDSO, which is not used in the HERT calculation. Daily human exposure: The average amount of the food consumed daily per person in the United States; when a chemical is listed rather than a food item, the value is the per person average in the total diet. All other calculations assume a daily dose for a lifetime. Possible hazard: The amount of chemical reported under "Human dose of chemical" is divided by 70 kg to give a mg/kg of human exposure. The HERT is this human dose (mg/kglday) as a percentage of the rodent LDso (mg/kg). An * preceding a chemical name indicates that the chemical is a natural pesticide. Abbreviations for LDso values: LO, LDLO; P, intraperitoneal injection; V, intravenous injection; J, injection (route not specified).
tion, and manufacturing procedures and additionally require evidence of efficacy and safety, dietary supplements are not. We found that several dietary supplements would rank high in the HERT table if we had included them by using the recommended dose and the LDso value for the extract: ginger extract (HERT = 0.8%), ginkgo leaf extract (HERT = 0.7%), ginseng extract (HERT = 0.7%), garlic extract (HERT = 0.1%), and valerian extract (HERT = 0.01%). These results argue for greater toxicological testing requirements and regulatory scrutiny of dietary supplements on the grounds that they may be carcinogens in rodents and that, if so, they are likely to rank high in possible carcinogenic hazard. Because these products lack requirements for toxicological testing, the NTP has established a research program on medicinal herbs and ingredients.
38.8 SUMMARY OF CARCINOGENICITY RESULTS IN THE CPDB ON ACTIVE INGREDIENTS OF COMMERCIAL PESTICIDES THAT HAVE BEEN EVALUATED BY THE U.S. EPA This section presents summary results on each of 193 commercial pesticide ingredients that are listed by the EPA in "Status of Pesticides in Registration, Reregistration, and Special Review," its Rainbow Report (U.S. Environmental Protection Agency, 1998) and that are also included in the CPDB. Results for pesticides that are negative for carcinogenicity in the CPDB are included. Approximately 1900 pesticides are listed in the Rainbow Report, but only 193 have published results of carcinogenicity experiments that meet the inclusion criteria of the CPDB (Gold and Zeiger, 1997). Table 38.12 provides a quick overview of the CPDB results on each pesticide, including the following information: the sex-species groups that have
been tested, the strongest level of evidence of carcinogenicity based on the opinion of the published author, carcinogenic potency (TDso), target organs in each species, and mutagenicity in Salmonella typhimurium. Carcinogenicity results for rats and mice are reported in Table 38.12, and in Table 38.13 for hamsters, monkeys, and dogs. For each pesticide, the details on each experiment are reported in the CPDB, published in the CRC Handbook of Carcinogenic Potency and Genotoxicity Databases (Gold et aI., 1997c) and in Environmental Health Perspectives (Gold et aI., 1999) as well as reported in http://potency.berkeley.edu. The following describe the data reported in Tables 38.12 and 38.13. Pesticides A chemical is considered a pesticide if it appears in the EPA "Status of Pesticides in Registration, Reregistration, and Special Review" (U.S. Environmental Protection Agency, 1998), the Rainbow Report. Included in the Rainbow Report is the status of pesticides that are undergoing pesticide reregistration, that have completed pesticide reregistration, that are under special review, or that are "new" (i.e., that have been registered since 1984). For 79 of the 193 commercial pesticides in the table, the active ingredient is no longer contained in any registered pesticide product; for these cases of voluntary or regulated cancellation, we indicate this fact by an asterisk next to the chemical name. If a commercial pesticide is also a chemical that occurs naturally, the chemical name is in boldface. Mutagenicity in Salmonella A chemical is classified as mutagenic in the Salmonella assay "+" if it was evaluated as either "mutagenic" or "weakly mutagenic" by Zeiger (1997) or as "positive" by the Gene-Tox Program (Auletta, personal communication; Kier et aI., 1986). All other chemicals evaluated for mutagenicity by these two sources are reported as "-." The symbol "." indicates that these sources did not provide an
38.8 Summary of Carcinogenicity Results in the Carcinogenic Potency Database
829
Table 38.12 Summary of Carcinogenicity Results in Rats and Mice in the Carcinogenic Potency Database on 193 Active Ingredients Commercial Pesticides that Have Been Evaluated by the V.S. Environmental Protection Agency Harmonic mean of Sal-
TDSO (mglkg/day)
Pesticide
CAS
moneIIa
Rat
Acrolein
107-02-8
Acrylonitrile*
107-13-1
+ +
16.9ffi .v
Mouse
Rat target sites Male
Female
ezy nrv orc smi
ezy mgl nas nrv
sto Aldicarb
116-06-3
Aldrin*
309-00-2
Allantoin*
97-59-6
Allyl isothiocyanate
57-06-7
3-Aminotriazoles
61-82-5
Anethole*
104-46-1
Anilazine*
101-05-3
Antimony potassium tartrate*
28300-74-5
Arsenate, sodium s
7631-89-2
Arsenious oxide
1327-53-3
Arsenite, sodium*
7784-46-5
Aspirin
50-78-2
Atrazine
1912-24-9
Azinphosmethy1
86-50-0
Benzaldehyde*
100-52-7
Benzene*
96 9.94ffi
thy
pit thy
B-
B-
Benzoic acid*
65-85-0
Benzyl aIcohol*
100-51-6
o-Benzyl-p-chlorophenol
120-32-1
Biphenyl*
92-52-4
Bis(tri-n-butyItin)oxide,
56-35-9
liv(B)
liv
liv
B-
B-
B-
B-
B-
B31. 7m
mgl
hmoute
ezy nas orc ski
ezy nas orc sto
+ 169m
1490ffi 77.5 m ,v
sto vsc 532-32-1
liv ubi
25.3 ffi
B-
71-43-2
Benzoate, sodium*
Female
orc smi sto
l.27ffi
+
Mouse target sites Male
sto
sto
ezy hag hmo
ezy hmo lun
lun pre
mglova
vsc
1350
kid
technical grade Boric acid
10043-35-3
tert-Butyl alcohol*
75-65-0
Butyl p-hydroxybenzoate*
94-26-8
p-tert-Butylphenol*
98-54-4
Cadmium chlorides *
10108-64-2
Calcium chloride*
10043-52-4
Capsaicin
404-86-4
Captan
133-06-2
Carbaryl
63-25-2
Carbon tetrachloride
56-23-5
Chloramben*
133-90-4
Chloranil*
118-75-2
Chlordane, technical grade*
57-74-9
Chlorinated trisodium phosphate
56802-99-4
Chlorine
7782-50-5
3-Chloro-p-toluidine
95-74-9
Chlorobenzilate*
510-15-6
(2-Chloroethyl) trimethyl-
999-81-5
64.6
21900
0.0114 ffi ,v
kid
thy
hmo 1un pro tes
lun
kid
ute
tba(B)
tba(B)
liv
Iiv mgl
167 ffi ,n
+ + +
2080m 14.1 2.29m ,n
2110m 150m
19i
19i
smi
smi
adrliv
adrliv
5230
liv
1.37m ,v
liv
liv
liv
liv
+ B93.9m ,v
B-
ammonium chloride
(continues)
830
CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
Table 38.12 (continued) Harmonic mean of SalPesticide
CAS
Chloroforms 3-(p-Chlorophenyl)-I,I-di-
monella
TDSO (mg/kg/day)
Rat target sites
Mouse target sites
Rat
Mouse
Male
Female
Male
Female
67-66-3
262 ffi
90.3 ffi
kid
liv
kid liv
liv
150-68-5
131
kid liv
2270ffi
kid
kid lun vsc
lun vsc
liv lun
lun
hrno liv lun
hmo liv lun
methylurea* Chloropicrin
76-06-2
Chlorothalonil
1897-45-6
Citric acid
77-92-9
Clonitralid
1420-04-8
Copper-8-hydroxyquinoline
10380-28-6
Coumaphos
56-72-4
Cyanamide, calcium*
156-62-7
Cyclohexanone*
108-94-1
Daminozide
1596-84-5
p,p'-DDD*
72-54-8
DDTs*
50-29-3
Deltamethrin
52918-63-5
Diallate*
2303-16-4
Diazinon
333-41-5
1,2-Dibromo-3-chloropropane*
96-12-8
+
I
kid
+ 2500n
1030ffi
84.7ffi
30.7 ffi 12.8ffi ,v
liv
liv
26.7 ffi
+ +
ute
0.259 ffi
2.72ffi
liv nas orc sto
adrmgl nas
lun nas sto
lun nas sto
lun sto vsc
eso lun mgl nas
orc sto 1,2-Dibromoethane*
106-93-4
3,5-Dichloro(N -I, l-dimethyl-2-
+
1.52ffi
23950-58-5
7.45 ffi ,V
nas per pit
liv lun mgl nas
sto vsc
pit sto vsc
119
sto sub vsc liv
propynyl)benzamide 2,3-Dichloro-1A-naphtho
117-80-6
quinone* 2,6-Dichloro-4-nitroaniline
99-30-9
1,2-Dichlorobenzene*
95-50-1
1A-Dichlorobenzene Dichlorodifluoromethane*
106-46-7 75-71-8
1,2-Dichloroethane*
107-06-2
a-(2,4-Dichlorophenoxy)propi-
120-36-5
+ 644
398 ffi
kid
+
8.04ffi
IOlffi
sto sub vsc
+
4.16
70.4ffi
hmo pan
mgl
liv
liv
lun
lun mgl ute
sto
sto
onic acid
204- Dichlorophenoxyacetic
acid
2,4-Dichlorophenoxyacetic
94-75-7 94-80-4
acid, n-butyl ester* 2,4-Dichlorophenoxyacetic
94-11-1
acid, isopropyl ester 3-(3A-Dichlorophenyl)-I,I-di-
330-54-1
methylurea Dichlorvos
62-73-7
Dicofol
115-32-2
32.9
liv
Dieldrins *
60-57-1
0.912ffi
liv
O,O-Diethyl-o-(3,5,6-trichloro-
2921-88-2
liv
2-pyridyl)phosphorothioate Dimethoate
60-51-5
Dimethoxane
828-00-2
+ +
716
hrno kid liv ski sub
(continues)
38.8 Summary of Carcinogenicity Results in the Carcinogenic Potency Database
831
Table 38.12 (continued) Harmonic mean of SalPesticide
CAS
Dimethylarsinic acid
75-60-5
2,4-Dinitrophenol*
51-28-5
Dioxathion*
78-34-2
n-Dodecylguanidine acetate
2439-10-3
EDTA, trisodium salt
150-38-9
monella
TDSO (mg/kglday) Rat
Mouse
Rat target sites Male
Female
Mouse target sites Male
Female
B-
B-
hag lun
haghmo lun
+
trihydrate* Endosulfan
115-29-7
Endrin*
72-20-8
Ethoxyquin
91-53-2
Ethyl alcohol
64-17-5
p, p'-Ethyl-DDD*
72-56-0
Ethylene glycol*
107-21-1
Ethylene oxide
75-21-8
Ethylenebisdithiocarbamate,
142-59-6
9110
adr liv pan pit
+ +
21.3 ffi ,v
63.7ffi
hmonrv per
nrv sto
mglute disodium di(2-Ethylhexyl)phthalate*
117-81-7
Eugenol
97-53-0
Fenaminosulf, formulated*
140-56-7
Fenthion
55-38-9
Fenvalerate
51630-58-1
Ferric dimethy ldithiocarbamate
14484-64-1
Fluometuron
2164-17-2
Fluoride, sodium
7681-49-4
Formaldehyde'
50-00-0
Fosetyl AI
39148-24-8
Furfural'*
98-01-1
Gibberellic acid
77-06-5
Glycerol a-monochlorohydrin*
96-24-2
Heptachlor
76-44-8
fJ-l,2,3,4,5,6-HexachlorocycIo-
625 ffi
894ffi
liv
Iiv
Iiv
2.19 m.v
43.9
hmonas
hmo nas
nas
liv
+
+
3660
+
683
ubi 197ffi
liv
liv
liv
+ Iiv
liv
319-85-7
1.21 m 27.8 m
liv
Iiv
58-89-9
30.7 m
liv
liv lun
hexane y -1 ,2,3,4,5,6-HexachlorocycIo-
hexane Hexachlorophene*
70-30-4
3-(Hexahydro-4,7 -methanoin-
2163-79-3
dan-5-yl)- I, I-dimethylurea* Hydrochloric acid
7647-01-0
Hydrogen peroxide
7722-84-1
8-Hydroxyquinoline*
148-24-3
Isopropyl-N -(3-chlorophenyl)
101-21-3
+ +
7540
smi
carbamate' Isopropyl-N -phenyl
122-42-9
carbamate' * Kepone*
143-50-0
Malathion
121-75-5
Maleic hydrazide
123-33-1
2.96
0.982m
liv
Iiv
liv
(continues)
832
CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
Table 38.12 (continued) Harmonic mean of SalPesticide
CAS
Manganese ethylenebisthiocar-
12427-38-2
2-Mercaptobenzothiazole*
149-30-4
2-Mercaptobenzothiazole, zinc
155-04-4
monella
TDSO (mg/kg/day) Rat
Mouse
Rat target sites Male
Female
157
tba(B)
tba(B)
344m
adr hmo pan pre
adrpit
Mouse target sites Male
Female
bamate
Mercuric chloride*
7487-94-7
Methidathion
950-37-8
Methoxychlor
72-43-5
Methyl bromide
74-83-9
Methyl parathion
298-00-0
Methylene chloride*
75-09-2
Metronidazole*
443-48-1
Mexacarbate*
315-18-4
Mirex*
2385-85-5
Naphthalene
91-20-3
I-Naphthalene acetamide
86-86-2
I-Naphthalene acetic acid
86-87-3
Nickel (11) sulfate
10101-97-0
3.12
st~
6.04
+ + + +
724m.i 542ffi l.77ffi
liv
1100m,i
mgl
mgl
livlun
liv lun
506ffi
pit tes
livmgl
lun
hmolun
1.45 ffi
adr kid liv
hmoliv
liv
liv
163 i
lun
+
hexahydrate* Nicotine
54-11-5
Nitrate, sodium
7631-99-4
Nitrite, sodium'
7632-00-0
Nitrofen*
1836-75-5
Oleate, sodium*
143-19-1
Oxamyl
23135-22-0
Oxytetracycline.HCI
2058-46-0
Parathion
56-38-2
Pentachloronitrobenzene
82-68-8
2,3,4,5,6-Pentachlorophenol
87-86-5
+ +
167ffi 420
hmo(B) liv 115ffi
hmo(B) liv pan
liv vsc
71.1 24ffi
liv
liv adr liv
adr liv vsc
(Dowicide EC-7)
Phenol
108-95-2
Phenothiazine*
92-84-2
Phenylmercuric acetate*
62-38-4
o-Phenylphenate, sodium
132-27-4
o-Phenylphenol
90-43-7
Phosphamidon*
13171-21-6
PicIoram, technical grade
1918-02-1
Piperonyl butoxide in solvent
51-03-6
Piperonyl sulfoxide*
120-62-7
Polysorbate 80*
9005-65-6
Potassium bicarbonate
298-14-6
Propazine
139-40-2
Propyl N -ethyl-N -butylthiocar-
1114-71-2
+ +
545 ffi ,v
kid ubi
232
ubi
ubi
62.2 13,OOOffi
liv ubi
ubi
adr nas
mgl nas
bamate n-Propyl isome*
83-59-0
Propylene glycol*
57-55-6
1,2-Propylene oxide
75-56-9
FD & C red no. 3*
16423-68-0
Rotenone
83-79-4
+
74.4ffi ,V
912ffi
st~
nas
nas
(continues)
38.8 Summary of Carcinogenicity Results in the Carcinogenic Potency Database
833
Table 38.12 (continued) Harmonic mean of SalPesticide
CAS
Safrole*
94-59-7
Simazine
122-34-9
Sodium bicarbonate
144-55-8
Sodium chloride
7647-14-5
Sodium chlorite
7758-19-2
Sodium dichromate
10588-01-9
Sodium hypochlorite
7681-52-9
Strobane*
8001-50-1
Sulfallate*
95-06-7
Telone II
542-75-6
2,4,5,4' -Tetrachlorodiphenyl
116-29-0
monella
TDso (mg/kg/day)
Rat target sites
Mouse target sites
Rat
Mouse
Male
Female
Male
Female
441ffi
5l.3 ffi ,v
liv
liv(B)
liv
liv
lun
+
4.M
+ +
26,lffi 94ffi
42.2ffi
sto
mgl
49.6
liv sto
sto
lOl ffi
126ffi
hmo kid
hmo
lun 0.884ffi
hmoliv mgl lun sto ubI
sulfone* Tetrachloroethylene*
127-18-4
Tetrachlorvinphos
961-11-5
Tetrakis(hydroxymethyl)phos-
55566-30-8
liv
liv
228
liv
5.57 ffi 584ffi
liv
liv
liv
liv
I 550ffi
smi sto
smi
liv
liv
tba
tba
phonium sulfate Tetramethylthiuram disulfide
137-26-8
Thiabendazole
148-79-8
Toxaphene*
8001-35-2
Trichloroacetic acid*
76-03-9
I, 1,1-Trichloroethane, technical
71-55-6
+ + +
grade* Trichlorofluoromethane*
75-69-4
N -(Trichloromethy lthio )phthal-
133-07-3
+
imide 2,4,6-Trichlorophenol*
88-06-2
2,4,5-Trichlorophenoxyacetic
93-76-5
405
lO70ffi
hmo
acid* Triethanolamine
102-71-6
Triethylene glycol
112-27-6
Trifluralin, technical grade
1582-09-8
Triphenyltin hydroxide
76-87-9
Urea*
57-13-6
Xylene mixture (60%
1330-20-7
lOOffi
+
330
liv lun sto
m-xylene, 9% o-xylene, 14% p-xylene, 17% ethylbenzene) FD & C yellow no.5
1934-21-0
Zinc dimethyldithiocarbamate
137-30-4
Zinc ethylenebisthiocarbamate*
12122-67-7
+
40.7 ffi
tba(B) thy
tba(B)
255
tba(B)
tba(B)
Abbreviations: " not tested; (B), data reported only for both sexes combined. Tissue codes: adr, adrenal gland; eso, esophagus; ezy, ear/Zymbal's gland; hag, harderian gland; hmo, hematopoietic system; kid, kidney; 19i, large intestine; liv, liver; lun, lung; mgl, mammary gland; nas, nasal cavity (includes tissues of the nose, nasal turbinates, paranasal sinuses, and trachea); nrv, nervous system; orc, oral cavity (includes tissues of the mouth, oropharynx, pharynx, and larynx); ova, ovary; pan, pancreas; per, peritoneal cavity; pit, pituitary gland; pre, preputial gland; pro, prostate; ski, skin; smi, small intestine; sto, stomach; sub, subcutaneous tissue; tba, all tumor bearing animals; tes, testes; thy, thyroid gland; ubI, urinary bladder; ute, uterus; vsc, vascular system. In a series of footnotes, we provide additional information about TDso values and test results in the CPDB. These are as follows: i, carcinogenic in rodents only by the inhalation route of administration; m, more than one positive test in the species in the CPDB; n, no results that were evaluated as positive by the published author for this species in the CPDB have statistically significant TDSO values (two-tailed p < 0.1); s, species other than rats or mice are reported for this chemical in Table 38.13; v, variation is greater than lO-fold among statistically significant (p < 0.1) TDso values from different positive experiments. Note: The commercial pesticides in boldface also occur naturally. 'Voluntary or regulated cancellations. The Active Ingredient Is No Longer Contained in Any Registered Pesticide Product.
834
CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
Table 38.13 Summary of Carcinogenicity Results in the Carcinogenic Potency Database in Other Species on 11 Commercial Pesticides Ingredients Evaluated by the V.S. Environmental Protection Agency Harmonic mean of Pesticide
CAS
Salmonella
TD50 (mg/kg/day)
Target sites
Hamsters 3-Aminotriazole
61-82-5
Cadmium chloride*
10108-64-2
DDT*
50-29-3
Dieldrin*
60-57-1
Formaldehyde
50-00-0
Furfural*
98-01-1
Isopropyl- N -(3-chlorophenyl)carbamate
101-21-3
Isopropyl-N -phenyl carbamate*
122-42-9
Nitrite, sodium
7632-00-0
+ + +
Cynomolgus monkeys DDT*
50-29-3
Rhesus monkeys Arsenate, sodium
7631-89-2
DDT*
50-29-3
Dogs Chloroform
67-66-3
The Commercial Pesticides in Boldface also Occur Naturally. 'Voluntary or regulated cancellations. The Active Ingredient Is No Longer Contained in Any Registered Pesticide Product. Abbreviations: ., not tested.
evaluation. For a chemical of interest, results for other genotoxicity tests are reported for some chemicals in the Genotoxicity Database (Zeiger, 1997). Carcinogenicity For each positive chemical in the CPDB, results are included on carcinogenic potency (by species) and target organ (by sex-species); if there are no positive results, then the symbol "-" appears. The classification of positivity in this summary table is based on a positive result in at least one experiment. There may be additional experiments on the same chemical that are negative in the CPDB, but this is not reflected in the table. An experiment is classified as positive or negative on the basis of the published author's opinion. A target site is classified as positive for NCIINTP if the evaluation in technical report was "carcinogenic" or "clear" or "some" evidence of carcinogenic activity ["c" or "p" on the plot of the CPDB (Gold et aI., 1997c, 1999)]. In the general literature, a site is classified as a target if the author of the published paper considered tumors to be induced by compound administration ("+" on the plot). In some cases authors do not clearly state their evaluation (blank in author's opinion in plot), and in some NCIINTP technical reports the evidence for carcinogenicity is considered "associated" or "equivocal"; these are not classified as positive. We use the author's opinion to determine positivity because it often takes into account more information than statistical significance alone, such as historical control rates for particular sites, survival and latency, and/or dose response. Generally, this
designation by author's opinion corresponds well with the results of statistical tests for the significance of the dose-response effect (two-tailed p < 0.01). For some chemicals, the only experiments in the CPDB for a species or a sex-species group were NCIINTP bioassays that were evaluated as inadequate, and we indicate these with an "I," in the potency and target organ fields. Carcinogenic Potency In the CPDB, a standardized quantitative measure of carcinogenic potency, the TD50, is estimated for each set of tumor incidence data. In a simplified way, the TD50 may be defined as follows: For a given target site(s), if there are no tumors in control animals, then the TD50 is that chronic dose rate (in mg/kg body weight/day) that would induce tumors in half the test animals at the end of a standard life span for the species. Because the tumor(s) of interest often does occur in control animals, the TD50 is more precisely defined as that dose rate (in mg/kg body weight/day) that, if administered chronically for the standard life span of the species, will halve the probability of remaining tumorless throughout that period. The TD50 is analogous to the LD50, and a low TD50 value indicates a potent carcinogen, whereas a high value indicates a weak one. The TD50 and the statistical procedures adopted for estimating it from experimental data have been described elsewhere (Gold et aI., 1997c; Peto et aI., 1984; Sawyer et aI., 1984). The range of TD50 across chemicals in the CPDB is at least lO millionfold for carcinogens in each sex of rat or mouse.
References
In Table 38.12, a carcinogenic potency value is reported for a chemical in each species with a positive evaluation of carcinogenicity in at least one test. If there is only one positive test on the chemical in the species, then the most potent TDso value from that test is reported. When more than one experiment is positive, in order to use all the available data, the reported potency value is a harmonic mean of the most potent TDso values from each positive experiment. We have shown that the harmonic mean is similar to the most potent TDso value for chemicals with more than one positive test (Gold et aI., 1989, 1997b). The harmonic mean (TH) is defined as TH =
ACKNOWLEDGMENTS We thank the many people who have worked on the analyses discussed in this chapter. Several collaborators were authors on work that has been updated in this chapter, including, Leslie Bernstein, David Freedman, David Gaylor, Bonnie R. Stem, Joseph P. Brown, Georganne Backrnan Garfinkel, Lars Rohrbach, and Estie Hudes. This work was supported through the University of California, Berkeley by National Institute of Environmental Health Sciences Center Grant ESOl896 (BNA and LSG), and by support for research in disease prevention from the Dean's Office of the College of Letters and Science (LSG); and by U.S. Department of Energy Grant DE-AC-03-76SF00098 through the E.O. Lawrence Berkeley National Laboratory (LSG).
1 --c--
1 n
;; L
i=l
1
REFERENCES
T,. I
To obtain the harmonic mean from each positive experiment, we select the lowest TDso value from among positively evaluated target sites with a statistically significant dose response (two-tailed p < 0.1). If no positive sites have a significant dose response, then we select the most potent (lowest TDso) from among positively evaluated sites with p :::: 0.1. When some experiments have positive significant results and others have only positive nonsignificant results, we discard the nonsignificant experimental results for the calculation of the harmonic mean. In some experiments, no TDso could be estimated because all dosed animals had the tumor of interest, and only summary data were available for animals with the tumor. For these cases, we use the 99% upper confidence limit of TDso as a replacement for the TDso. In a series of superscripts following the TDso value, we provide additional information about the carcinogenic potency and other test results in the CPDB. These are as follows: i = carcinogenic in rodents only by the inhalation route of administration. m = more than one positive test in the species in the CPDB. n
835
= no
results that were evaluated as carcinogenic by the published author for this species in the CPDB have statistically significant TDso values (two-tailed p < 0.1). s = species other than rats or mice are reported for this chemical in Table 38.13. v = variation is greater than lO-fold among statistically significant (two-tailed p < 0.1) TDso values from different positive experiments.
Target Sites Target sites are reported for each sex-species group with a positive result in the CPDB. Target sites are identified on the basis of an author's positive opinion for the particular site, in any experiment in the sex-species, using all results from both the general literature and the NCIINTP bioassays. Hence, if a chemical has two target sites listed in a sex-species, the results may represent two different experiments. Occasionally, the CPDB results are only for both sexes combined and this has been indicated with (B) next to the target site.
Adachi, Y., Moore, L. E., Bradford, B. U., Gao, w., and Thurman, R. G. (1995). Antibiotics prevent liver injury in rats following long-term exposure to ethanol. Gastroenterology 108, 218-224. Adamson, R. H., Shozo, T., Sugimura, T., and Thorgeirsson, U. P. (1994). Induction of hepatocellular carcinoma in nonhuman primates by the food mutagen 2-amino-3-methylimidazo[4,S-flquinoline. Environ. Health Perspect. 102, 190-193. Ahmed, S. S., and Miiller, K. (1978). Effect of wound-damages on the glycoalkaloid content in potato tubers and chips. Lebensm.-Wiss. Technol. 11, 144-146. American Cancer Society (2000). "Cancer Facts and Figures-2000." American Cancer Society, Atlanta. American Medical Association (AMA) Division of Drugs (1983). "AMA Drug Evaluations," 5th ed., pp. 201-202. AMA, Chicago. American Water Works Association (AWWA) (1993). "DisinfectantlDisinfection By-Products Database for the Negotiated Regulation." AWWA, Washington, DC. Ames, B. N. (1983). Dietary carcinogens and anti-carcinogens: Oxygen radicals and degenerative diseases. Science 221, 1256-1264. Ames, B. N. (1984). Cancer and diet. Science 224, 668-670, 757-760. Ames, B. N. (2001). DNA damage from micronutrient deficiency is likely to be a major cause of cancer. Mutat. Res. 475, 7-20. Ames, B. N., and Gold, L. S. (1990). Chemical carcinogenesis: Too many rodent carcinogens. Proc. Natl. Acad. Sci. U.S.A. 87, 7772-7776. Available at http://socrates.berkeley.edu/mutagenlpnasl.html. Ames, B. N., and Gold, L. S. (1991). Risk assessment of pesticides. Chem. Eng. News 69, 28-32, 48-49; Forum: 27-55. Ames, B. N., Gold, L. S., and Shigenaga, M. K. (1996). Cancer prevention, rodent high-dose cancer tests, and risk assessment. Risk Anal. 16,613-617. Available at http://potency.berkeley.edu/textlriskanaleditorial.html. Ames, B. N., Gold, L. S., and Willett, W. C. (1995). The causes and prevention of cancer. Proc. Natl. Acad. Sci. U.S.A. 92, S2S8-526S. Available at http://socrates.berkeley.edulmutagenlames. pnas3.html. Ames, B. N., Magaw, R., and Gold, L. S. (1987). Ranking possible carcinogenic hazards. Science 236, 271-280. Letters: 237,235 (1987); 237, 1283-1284 (1987); 237, 1399-1400 (1987); 238, 1633-1634 (1987); 240, 1043-1047 (1988). Ames, B. N., Profet, M., and Gold, L. S. (I 990a). Dietary pesticides (99.99% all natural). Proc. Natl. Acad. Sci. U.s.A. 87, 7777-7781. Available at http://socrates.berkeley.edulmutagenlpnas2.html. Ames, B. N., Profet, M., and Gold, L. S. (l990b). Nature's chemicals and synthetic chemicals: Comparative toxicology. Proc. Natl. Acad. Sci. U.S.A. 87, 7782-7786. Available at http://socrates.berkeley.edu/mutagenlpnas3.html. Ames, B. N., Shigenaga, M. K., and Gold, L. S. (1993a). DNA lesions, inducible DNA repair, and cell division: Three key factors in mutagenesis and carcinogenesis. Environ. Health Perspect. 101, 35-44. Ames, B. N., Shigenaga, M. K., and Hagen, T. M. (1993b). Oxidants, antioxidants, and the degenerative diseases of aging. Proc. Natl. Acad. Sci. U.s.A. 90,7915-7922.
836
CHAPTER 38
Pesticide Residues in Food and Cancer Risk: A Critical Analysis
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Pesticide Residues in Food and Cancer Risk: A Critical Analysis
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CHAPTER
39 Perceptions of Pesticides as Risks to Human Health Paul Slovic Decision Research, Inc.
39.1 INTRODUCTION Public perceptions of risk have been studied systematically for more than 20 years, within the United States and in other countries. Throughout that time period, the use of pesticides has been perceived as one of the most risky activities pursued by human societies. It is difficult to pinpoint the origins of these perceptions but certainly Rachel Carson's book, Silent Spring, first published in 1962, has played an important role. In Carson's story, pesticides are singled out as among the most potent "elixirs of death." Referring to synthetic pesticides, Carson observed, "They have immense power not merely to poison but to enter the most vital processes of the body and change them in sinister and often deadly ways" (Carson, 1962, p. 25). Ironically, as will be shown, the heavy reliance on testing of chemicals with animals and the quantitative risk assessment that has developed during this same time period may have reinforced and maintained the public's fears of pesticides and other chemicals.
39.2 RISK-PERCEPTION STUDIES One of the first quantitative studies of risk perception took place in the United States in the late 1970s and early 1980s. This research showed that perceptions of risk can be described in terms of numerous characteristics or dimensions. Figure 39.1, for example, presents a spatial display of hazards within a perceptual space derived from more than 40,000 individual judgments. The factors in this space reflect the degree to which a hazard is perceived to be known or understood (vertical dimension) and the degree to which it evokes perceptions of dread, uncontrollability, and catastrophe (horizontal dimension). Research has also demonstrated that social response to risk is closely related to the position of a hazard within this space. The further to the right a hazard appears, the higher its perceived risk, the more people want to see its current risks reduced, and Handbook of Pesticide Toxicology Volume 1. Principles
the more they want to see strict regulation employed to achieve reduced risk. Media coverage appears to be most extensive and intense when something goes wrong in the upper right-hand quadrant of the space, in an activity whose risks are seen as being poorly understood, evoking dread, and potentially leading to a catastrophe. In this light, it is interesting to see that pesticides fall in the problematic upper right quadrant of the space, reflecting the fact that respondents in this study characterized them as poorly known or understood, delayed in effect, relatively new, uncontrollable, evoking dread, catastrophic, fatal, inequitable, and posing high risk to future generations. Their location in this space is not too distant from activities related to the use of nuclear power. Whereas public judgments of risk seem closely related to the characteristics that define the space in Fig. 39.1, expert's judgments of risk are not closely related to any of these various risk characteristics. Instead, experts appear to see riskiness as synonymous with expected annual mortality. As a result, many conflicts in society over "risk" may result from experts and laypeople having different definitions of the concept. In this light, it is not surprising that expert recitations of "risk statistics" often do little to change people's attitudes and perceptions. In addition to constructing spatial displays such as that in Fig. 39.1, research has compared perceptions of risk and benefit from a large number of activities and technologies. It is particularly instructive to compare perceptions of various radiation and chemical technologies. Nuclear power has a very high perceived risk and low perceived benefit, whereas diagnostic X-rays have the opposite pattern (low perceived risk, high perceived benefit). A parallel finding occurs with chemicals. Nonmedical sources of exposure to chemicals (e.g., pesticides, food additives, alcohol, cigarettes) are seen as very low benefit and high in risk; chemicals used in medicine (e.g., prescription drugs, antibiotics, vaccines) are generally seen as high in benefit and low in risk, despite the fact that they can be very toxic substances.
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Copyright © 200 1 by Academic Press. All rights of reproduction in any form reserved.
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CHAPTER 39
Perceptions of Pesticides as Risks to Human Health
Factor 2 Unknown risk
Laetrile. Microwave Ovens.
• DNA Technology
Water Fluoridation. Saccharin. .Nitrates Water Chlorination.·Hexachlorophene Coal Tar Hairdyes. Polyvinyl. · • Chloride Ora I Con tracep tIves. Diagnostic Valium • • IUD • Darvon
X-Rays Antibiotics. .Rubber Mfg.
Auto Lead. • Lead Paint
• Caffeine .Aspirin
.Electric Fields .DES .Nitrogen Fertilizers
.SST
• Radioactive Waste .Cadmium Usage .2,4,5-T Mirex • Trichloroethylene • Nuclear Reactor Accidents .Uranium Mining • Pesticides • Nuclear Weapons • Asbestos • PCBs Fallout Insulation • lOOT • Satellite Crashes Mercury FisSil Fuels Coal Burning (Pollution)
-r__~~~--~~--~~~~--~-------------------Factor1
______________._v_a_~~in_es____________ • Skateboards
Smoking (Disease). • Power Mowers • Snow mobiles Trampolines.
• Tractors • Alcohol • Chainsaws
. . • Elevators Home SWlmmln9 Po?ls • • Electric WIr & Appl ( Downhill SkIIng. . Recreational Boating • • Smoking Motorcycles Electric WIr & Appl (Shock). ·Bicycles
•
.Auto Exhaust (CO) • LNG Storage & .D-CON Transport I M" (D' ) oa Inlng Isease • Large Dams • SkyScraper Fires
.C
Dread risk
• Nerve Gas Accidents
Nuclear Weapons (War) • Underwater • Construction .Coal Mining Accidents • Sport Parachutes • General Aviation • High Construction • Railroad Collisions lcohol • Commercial Aviation ccidents • Auto Racing Auto A~idents • Handguns • Dynamite
1Factor 21
Controllable Not Dread Not Global Catastrophic Consequences Not Fatal Equitable Individual Low Risk to Future Generations Easily Reduced Risk Decreasing Voluntary
Not Observable Unknown to Those Exposed Effect Delayed New Risk Risk Unknown to Science
Observable Known to those Exposed Effect Immediate Old Risk Risks Known to Science
Uncontrollable Dread Global Catastrophic Consequences Fatal Not Equitable Catastrophic High Risk to Future Generations Not Easily Reduced Risk Increasing Involuntary
1
Factor 1 1
Figure 39.1 Location of 81 hazards in a two-factor space derived from the relationships among 15 risk characteristics. Each factor is made up of a combination of characteristics, as indicated by the lower diagram. [Slovic, P., Fischhoff, B., and Lichtenstein, S. (1985). Characterizing perceived risk. In "Perilous Progress: Technology as Hazard" (R. W. Kates, C. Hohenemser, and J. X. Kasperson, eds.), pp. 91-123. Westview, Boulder, CO.]
Research throughout the 1980s and 1990s has continued to show high levels of concern regarding the risks from the use of pesticides. Figures 39.2 and 39.3 present data from parallel national surveys in the United States and France conducted in 1992. The U.S. survey shows significant concerns regarding
pesticides in food, which were seen as close in risk to drinking alcohol, motor vehicle accidents, and nuclear power plants and higher in risk than two rather significant hazards, bacteria in food and storms and floods. The picture in France was similar (Figure 39.3). There was a somewhat greater frequency of
39.2 Risk-Perception Studies
847
Cigarette smoking Street drugs AIDS Nuclear waste Stress Chemical pollution Ozone depletion Suntanning Drinking alcohol Motor vehicle accidents Pesticides in food Nuclear power plants Blood transfusions Outdoor air quality Climate change Bacteria in food High-voltage power lines Food irradiation Coal/oil power plants Genet engr bacteria Radon in home Storms & floods VDTs Commercial air travel Medical X-rays 00%
•
Figure 39.2
20%
High risk
D
40%
60%
Moderate ~ Slight risk L::::::.:::J risk
100%
80%
' D ;
,Almost no risk
D
Don't know
Perceived health risks to American public: 1992 national survey.
high-risk responses to pesticides in food in France than in the United States. Similar results were obtained in a 1993 national survey of the Canadian public (Figure 39.4), with more than 30% of Canadians judging pesticides in food as high risk and more than 70% judging them as high or moderate risk. Pesticides in food were rated as higher in risk than drinking alcohol, nuclear power plants, asbestos, and bacteria in food. When the same survey was given to members of the Society of Toxicology of Canada (see Fig. 39.5), pesticides in food were judged far lower in risk (less than 10% judged them as high risk; only about 20% judged them high or moderate risk). Contrary to the public's opinions, the toxicologists judged bacteria in food to be riskier than pesticides in food. Members of the British Toxicology Society (BTS) surveyed by Slovic et al. (1997) also judged pesticides in food to pose rather small risks (see Fig. 39.6). The most recent data that my colleagues and I have collected come from a national survey in the United States in 1997. Pes-
ticides were seen as posing high risk to the American public by 26% of the respondents and posing high or moderate risk by 69%. Pesticides were rated almost as risky as stored nuclear waste, motor vehicles, nuclear power plants, and natural disasters (see Fig. 39.7). In another segment of that survey, respondents judged risks to individuals to be almost or as great from pesticides as from handguns and violent crime. 39.2.1 SOCIAL, CULTURAL, AND POLITICAL INFLUENCES ON RISK PERCEPTION
The data shown in Figs. 39.1-39.7 reflect only the tip of the iceberg. Below the surface are rumblings of quite complex social, cultural, and political forces shaping the observed ratings of risks from pesticides and other hazards. Recent studies have shown that such factors as gender, race, political worldview, affiliation, emotion, and trust are strongly correlated with risk
848
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Perceptions of Pesticides as Risks to Human Health
Nuclear waste AIDS Street drugs Cigarette smoking Chemical pollution Motor vehicle accidents Ozone depletion Drinking alcohol Stress Pesticides in food Genet engr bacteria Bacteria in food Suntanning Blood transfusions Nuclear power plants Food irradiation Climate change Outdoor air quality Storms & floods High-voltage power lines VDTs Medical X-rays Coal/oil-burning plants Commercial air travel Radon in home 00% High nsk
Figure 39.3
20%
40%
Moderate risk
60% Almost no nsk
80%
c
100% Don·t know
Perceived health risks to French public: 1992 national survey.
judgments. Equally important is that these factors influence the judgments of experts as well as the judgments of laypersons. For example, gender is strongly related to risk judgments and attitudes. Several dozen studies have documented the finding that men tend to judge risks as smaller and less problematic than do women (Brody, 1984; Carney, 1971; DeJoy, 1992; Gutteling and Wiegman, 1993; Gwartney-Gibbs and Lach, 1991; Pillisuk and Acredolo, 1988; Sj6berg and Drottz-Sj6berg, 1993; Slovic et aI., 1989, 1993; Spigner et aI. , 1993; Steger and Witt, 1989; Stem et aI. , 1993). A number of hypotheses have been put forward to explain sex differences in risk perception. One approach has been to focus on biological and social factors. Women have been characterized as more concerned about human health and safety because they are socialized to nurture and maintain life (Steger and Witt, 1989). They have been characterized as physically more vulnerable to violence, such as rape, and this may sensitize them to other risks (Baumer, 1978; Riger et al., 1978). The combination of biology and social experience
has been put forward as the source of a "different voice" that is distinct to women (Gilligan, 1982; Merchant, 1980). A lack of knowledge and familiarity with science and technology has also been suggested as a basis for these differences, particularly with regard to nuclear and chemical hazards. Women have been discouraged from studying science and there are relatively few women scientists and engineers (Alper, 1993). However, Barke, Jenkins-Smith, and Slovic have found that female physical scientists judge risks from nuclear technologies to be higher than do male physical scientists (Gilligan, 1982; Merchant, 1980). Similar results with scientists were obtained by Slovic et al. (1997), who found that female members of the BTS were far more likely than male toxicologists to judge societal risks, including pesticides, as moderate or high. Certainly female scientists in these studies cannot be accused of lacking knowledge and technological literacy. Some other factors must influence their decisions. Hints about the origin of these sex differences come from a study by Flynn et al., in which 1512 Americans were asked, for
39.2 Risk-Perception Studies
Cigarette smoking Ozone depletion Breast implants Street drugs Stress Chemical pollution Crime and violence Suntanning AIDS Motor vehicle accidents Nuclear waste Alcohol and pregnancy PCBs or dioxin Pesticides in food Food additives Drinking alcohol Nuclear power plants Climate change on prescription med. Asbestos Waste incinerators Malnutrition Hi-volt power lines Food irradiation Prescription drugs Genet engr bacteria Outdoor air quality Bacteria in food Molds in food Mercury in fillings Tap water Medical X-rays Indoor air quality VDTs Contraceptives Heart pacemakers Bottled water Contact lenses
-~
I
849
---• • •• • • I
I
~
••
I
I
I
I
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--•
00 %
I _ ~~h
-• ---I
• I
•
I
I
I
20%
60%
40%
~
Moderate risk
Slight nsk
I
80%
I Almost no risk
100%
Don't know -----l
Figure 39.4 Health risks to the Canadian public: 1992 Health and Welfare Canada survey (N = 1506). [Slovic, P., Malmfors, T., Krewski, D., Mertz, C. K., Neil, N., and Bartlett, S. (1995). Intuitive toxicology. n. Expert and lay judgments of chemical risks in Canada. RiskAnal. 15,661-675.]
each of 25 hazard items, to indicate whether the hazard posed (1) little or no risk, (2) slight risk, (3) moderate risk, or (4) high risk to society (Flynn et aI., 1994). Figure 39.8 shows the difference in the percentage of males and females who rated a hazard as a "high risk." All differences are to the right of the 0% mark, indicating that the percentage of high-risk responses was greater for women on every item (Flynn et aI., 1994). Perhaps the most striking result from this study is shown in Fig. 39.9, which presents the mean risk ratings separately for white males, white females, nonwhite males, and nonwhite
females. Across the 25 hazards, white males produced riskperception ratings that were consistently much lower than the means of the other three groups (Flynn et aI., 1994). When the data underlying Fig. 39.9 were examined more closely, Flynn et al. observed that not all white males perceived risks as low. The "white-male effect" appeared to be caused by about 30% of the white-male sample who judged risks to be extremely low (Flynn et aI., 1994). The remaining white males were not much different from the other subgroups with regard to perceived risk.
850
CHAPTER 39
Perceptions of Pesticides as Risks to Human Health
Cigarette smoking Motor vehicle accidents Stress Alcohol and pregnancy Crime and violence Suntanning Breast implants Street drugs AIDS Ozone depletion Drinking alcohol Asbestos Molds in food Nuclear waste PCBs and dioxins Bacteria in.food Chemical pollution Malnutrition Nuclear power reactors Medical X -rays Climate changes Pesticides in food Indoor air pollution Rx drugs Waste incinerators Contraceptives Non-Rx drugs Power lines Outdoor air quality Food additives Tap water Contact lenses Genet engr bacteria Pacemakers VDTs Mercury in fillings Food irradiation Bottled water
_.
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20%
I
40%
60%
, ~oderate- S~ght fISk
-
fisk
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80%
Almost no risk
-
100%
Don't know
Figure 39.5 Health risks to the Canadian public: 1993 survey of the Society of Toxicology of Canada (N = 150). [Slovic, P., Malmfors, T., Krewski, D., Mertz, C. K. , Neil, N., and Bartlett, S. (1995). Intuitive toxicology.
n. Expert and lay judgments of chemical risks in Canada. Risk Anal.
Further analyses showed that the subgroup of white males who perceive risks to be quite Iow can be characterized by very great trust in institutions and authorities and by anti-egalitarian attitudes, including a disinclination toward giving decisionmaking power to citizens in areas of risk management. The results of this study raise new questions. What does it mean for the explanations of gender differences when we see that the sizable differences between white males and white females do not exist for non white males and nonwhite females? Why do a substantial percentage of white
15,661-675.)
males see the world as so much less risky than everyone else sees it? Obviously, the salience of biology is reduced by these data on risk perception and race. Biological factors should apply to nonwhite men and women as well as to white men and women. The present data thus move us away from biology and toward sociopolitical explanations. Perhaps white males see less risk in the world because they create, manage, control, and benefit from many of the major technologies and activities. Perhaps women and non white men see the world as more dangerous
39.3 Intuitive Toxicology: Expert and Lay Judgments of Chemical Risks
851
Cigarette smoking Asbestos Dioxins Motor vehicle traffic Suntanning Nuclear waste Crime and violence
liii~~~~~~~i~~~;liiliiiiiii~~
Alcoholic beverages Environmental tobacco smoke
1:=:=======:;;;;;;;;
Chemicals in the workplace Depletion of the ozone layer Burning fossil fuels Nuclear power reactors Breast implants Radon in homes Outdoor air pollution Waste incinerators Indoor air pollution Prescription drugs Pesticides in food
i~~~iiii!iiiiii~!~~~~~~~~~~~1
Non-prescription drugs Contraceptive pills Medical X-rays Electric and magnetic fields Food irradiation Mercury in dental fillings Food additives Tap water 00%
20%
40%
~ ~oderate. S~ght ~ r~k r~k
60 %
D
80%
Almost no risk
.
100%
Don't know
Figure 39.6 Health risks to the average exposed citizen of your country: 1994 British Toxicology Society (N = 312). [Slovic, P. , Malmfors, T., Mertz, C. K., Neil , N. , and Purchase, 1. F. H. (1997). Evaluating chemical risks: Results of the British Toxicology Society. Hum. Exp. Toxicol. 16,289-304.]
because in many ways they are more vulnerable, because they benefit less from many of its technologies and institutions, and because they have less power and control over what happens in their communities and their lives. Although the survey conducted by Flynn et al. was not designed to test these alternative explanations, the race and gender differences in perceptions and attitudes point toward the role of power, status, alienation, trust, perceived government responsiveness, and other sociopolitical factors in determining perception and acceptance of risks. According to this view, the problem of risk conflict and contro-
versy goes beyond science. It is deeply rooted in the social and political fabric of our society.
39.3 INTUITIVE TOXICOLOGY: EXPERT AND LAY JUDGMENTS OF CHEMICAL RISKS The preceding sections have described general attitudes and perceptions regarding pesticides and other hazards. In parallel
852
CHAPTER 39
Perceptions of Pesticides as Risks to Human Health
Multiple sexual partners Street drugs 2nd-hand cigarette smoke Stored nuclear waste Motor vehicles Chemical manufacturing Nuclear power plants Natural disasters Pesticides Blood transfusions Lead in dust or paint Coal/oil-burning p. plants Tap water Airplane travel Radon in homes Electromagnetic fields Cellular phones Vaccines Asteroids 00%
• I
Figure 39.7 Survey.]
20%
High nsk
D
40%
Moderate fisk
[SJ Slight fisk
60%
U
80%
Almost no fisk
100%
~ Don·t ~
know
Perceived health risks to American public: U.S. population as a whole. [1997 National Risk
with this work, another stream of research has focused on chemical risks and attempted to go beneath the surface differences between experts and laypersons to document and understand the causes of these different views. This research has centered around a concept labeled intuitive toxicology (Kraus et aI., 1992; Slovic et aI., 1995, 1997). Research on intuitive toxicology was motivated by the premise that different assumptions, conceptions, and values underlie much of the discrepancy between expert and lay views of chemical risks. Research attempted to address this issue by exploring the cognitive models, assumptions, and inference methods that comprise laypeople's "intuitive toxicological theories" and comparing these theories with the cognitive models, assumptions, and inference methods of scientists working in the field of toxicology. The work began by identifying several fundamental principles and judgmental components within the science of risk
assessment. Questions were developed based on these fundamentals in order to determine the extent to which laypeople and experts share the same beliefs and conceptual framework. Questions addressed the following topics: (a) dose-response sensitivity; (b) trust in animal and bacterial studies; (c) attitudes toward chemicals; (d) attitudes toward reducing chemical risks; (e) conceptions of toxicity, including the toxicity of natural versus synthetic substances and the toxicity of prescription drugs versus chemicals in general; and (f) interpretation of evidence regarding cause-effect relationships between exposure to chemicals and human health. Questions on these topics were incorporated into a survey designed for both experts and the public. Each question was designed, whenever possible, according to a guiding hypothesis about how experts and "lay toxicologists" might respond. For example, a key principle in toxicology is the fact that "the dose makes the poison." Any substance can cause a toxic effect if the
39.3 Intuitive Toxicology: Expert and Lay Judgments of Chemical Risks
Stress Suntanning Nuclear Waste Nuclear Power Plants Ozone Depletion AIDS Drinking Alcohol Hi-Volt Power Lines Street Drugs Motor Vehicle Accidents Blood Transfusions Chemical Pollution Pesticides in Food Bacteria in Food Cigarette Smoking Storms & Floods Radon in Home Climate Change Food Irradiation Outdoor Air Quality Coal/Oil Burning Plants Genet Engr Bacteria Medical X-Rays Commercial Air Travel VDTs
853
45.3 34.3 52.4 25.9 38.8 54.7 34.7 15.8 55.6 32.9 25.1 41.6 32.0 18.7 57.9 11.5 12.2 22.9 18.0
-10%
-5%
0%
5%
10%
15%
20%
25%
Percent difference in high risk Figure 39.8 Perceived health risks to American public by gender: difference between males and females. Note: Base percentage equals male high-risk response. Percentage difference is female high-risk response minus male high-risk response. [F1ynn, J., Slovic, P., and Mertz, C. K. (1994). Gender, race, and perception of environmental health risks. Risk Anal. 14, 1101-1108.]
dose is great enough. Thus, experts were predicted to be quite sensitive to considerations of exposure and dose when responding to questions on this topic. In contrast, the often-observed concerns of the public regarding very small exposures or doses of chemicals led to the hypothesis that the public would have more of an "all-or-none" view of toxicity and would be rather insensitive to concentration, dose, and exposure (thus equating any exposure with harm). Because the science of toxicology and the discipline of risk assessment rely so heavily upon animal studies, experts were predicted to have a more favorable view than laypersons regarding the value of such studies. The prediction that laypersons lack sensitivity to dose-response considerations and thus fear even small exposures to toxic or carcinogenic substances led to the prediction that they would exhibit far more negative attitudes toward chemicals than experts. This last prediction was confirmed dramatically in the studies. The members of the public who responded to these surveys associated exposure to chemicals to a remarkable extent with danger, cancer, and death, consistent with the general opinions described in Figs. 39.1-39.7 for pesticides and other chemicals. Specifically, studies of intuitive toxicology on national populations in the United States, Canada, and France have found that about 70% of the public believe that "if a person is exposed to a chemical that can cause cancer, then that person will probably
get cancer some day" (Kraus et al., 1992; Krewski et al., 1995). About 75% of the respondents in these surveys agreed that "If even a tiny amount of a cancer-producing substance was found in my tap water, I wouldn't drink it." More than 50% agreed that "There is no safe level of exposure to a cancer-causing chemical." The concern that any exposure to a carcinogen, no matter how small, is likely to cause cancer is linked to a desire to avoid chemicals and reduce the risks of exposure to them at any cost. About 75% of the public surveyed agreed that "I try hard to avoid contact with chemicals and chemical products in my daily life." About 62% agreed that "It can never be too expensive to reduce the risks from chemicals." Responses of toxicologists were not at all in agreement with these views. Of particular importance in this research is the finding, as predicted, that the public is much less sensitive than the experts to considerations of dose and exposure. Although the public recognizes the importance of these factors in some domains (e.g., prescription drugs), they generally tend to view chemicals as either safe or dangerous and they appear to equate even small exposures to toxic or carcinogenic chemicals with almost certain harm. This orientation was found to be associated with high levels of concern regarding chemicals, including very small residues of chemicals on food, and a desire to reduce chemical risks at any cost.
854
CHAPTER 39
Perceptions of Pesticides as Risks to Human Health
---+- White Male
-
-+- - White Female
____ Non-White Male - ..... - Non-White Female I _ ______
._ ~.-------.J
Cigarette Smoking Street Drugs AIDS Stress Chemical Pollution Nuclear Waste Motor Vehicle Accidents Drinking Alcohol Suntanning Ozone Depletion Pesticides in Food Outdoor Air Quality Blood Transfusions Coal/Oil Burning Plants Climate Change
'a. I
Bacteria in Food Nuclear Power Plants Food Irradiation Storms & Floods Genet Engr Bacteria Radon in Home Hi-Volt Power Lines VDTs Medical X-Rays Commercial Air Travel
L-_ _ _ _ _ _ _ _ _ _ _ _ _ _ _ _ _ _ _---'
2
3
4
Slight Risk
Moderate Risk
High Risk
Figure 39.9 Mean risk-perception ratings by race and gender. [Flynn, J., Slovic, P., and Mertz, C. K. (1994). Gender, race, and perception of environmental health risks. Risk Anal. 14, 1104.]
Views on the validity of animal studies have been found to be more complex than expected. Consider two survey items that have been studied repeatedly. One is statement SI: "Would you agree or disagree that the way an animal reacts to a chemical is a reliable predictor of how a human would react to it?" The second statement, S2, is a little more specific: "If a scientific study produces evidence that a chemical causes cancer in animals, then we can be reasonably sure that the chemical will cause cancer in humans." When members of the American and Canadian public responded to these items, they showed moderate agreement with SI; about half the people agreed and half disagreed that animal tests were reliable predictors of human reactions to chemicals. However, in response to S2, which stated that the animal study found evidence of cancer, there was a jump in agreement to about 70% among both male and female respondents (see Fig. 39.10). The important point about the pattern of response is that agreement was higher on the second item. What happens when toxicologists are asked about these two statements? Figure 39.11 shows that toxicologists in the United States and toxicologists in the United Kingdom responded sim-
ilarly to the public on the first statement but differently on the second (Kraus et aI., 1992). They exhibited the same rather middling level of agreement with the general statement about
Percent agree 70% 60% 50% 40% 30% 20% 10% 00% Men
Women
Figure 39.10 Agreement among members of the V.S. public with statements SI and S2. [Kraus, N. N., Malmfors, T., and Slovic, P. (1992). Intuitive toxicology: Expert and lay judgments of chemical risks. Risk Anal. 12,215-232.]
39.3 Intuitive Toxicology: Expert and Lay Judgments of Chemical Risks
Percent agree 90% 75% 60% 45% 30% 15% 00% U.S. Public N=262
U.S Toxicologists N= 170
U.K. Toxicologists N=312
Figure 39.11 Agreement among the public and toxicologists with statements SI and S2. [Slovic, P. (1997). Trust, emotion, sex, politics, and science: Surveying the risk-assessment battlefield. In "Environment, Ethics, and Behavior" (M. H. Bazerman, D. M. Messick, A. E. Tenbrunsel, and K. A. Wade-Benzoni, eds.), p. 299. New Lexington, San Francisco.)
animal studies as predictors of human health effects.! However, when these studies were said to find evidence of carcinogenicity in animals, the toxicologists were less likely to agree that the results could be extrapolated to humans. Thus, findings that lead toxicologists to be less willing to generalize to humans lead the public to see the chemical as more dangerous for humans. Figure 39.12 presents the responses for S! and S2 among men and women toxicologists in the United Kingdom (208 men and 92 women). Here, we see another interesting finding. The men agree less on the second statement than on the first, but the women agree more, just like the general public. Among toxicologists, women are more willing than men to say that one can generalize to humans from positive carcinogenicity findings in animals.
Percent agree 90% 75% 60% 45% 30% 15% 00% Men (-23, n = 208)
Women (+11 . n=92)
Figure 39.12 Agreement among men and women toxicologists in the United Kingdom with statements SI and S2. [Slovic, P. (1997). Trust, emotion, sex, politics, and science: Surveying the risk-assessment battlefield. In "Environment, Ethics, and Behavior" (M. H. Bazerman, D. M. Messick, A. E. Tenbrunsel, and K. A. Wade-Benzoni, eds.), pp. 299. New Lexington, San Francisco.) 1This is a rather surprising result, given the heavy reliance on animal studies in toxicology.
855
These studies of intuitive toxicology have yielded a number of intriguing findings that likely pertain to views about pesticides. One is the low percentage of agreement that animal studies can predict human health effects. Another is that toxicologists show even less confidence in equating human cancers with studies that find cancer in animals resulting from chemical exposure. The public, on the other hand, has high confidence in animal studies that find cancer. These studies also help us understand why the public has come to fear pesticides and other chemicals so greatly. As regulators have sought to develop more effective ways to meet public demands for a safer and healthier environment, they have come to rely heavily on quantitative risk assessment based on animal tests. Such tests often find evidence of cancer at high dose levels. Many scientists are skeptical of such evidence, on the grounds that high doses overwhelm the animals' defense mechanisms and produce cancers that would not occur in humans under normal conditions of exposure. This skepticism is seen in the high percentage of toxicologists who lack confidence in evidence for carcinogenicity derived from animal studies. The public, on the other hand, exhibits a high degree of confidence in positive findings from animal studies. Thus, the large number of animal studies performed over the years may have done a better job of scaring the public than of informing science about chemical carcinogenesis. Another contributing factor is that interpretation of the animal data has been based on a linear model that assumes that there is no level of exposure to a carcinogen that is without some degree of risk. Multiplying even very small probabilities of contracting cancer across large numbers of exposed individuals will likely project at least some number of deaths. This frightens people. Using upper 95% confidence bounds in the linear extrapolation makes the scenario even more frightening. Thus, the many people who believe there is no safe level of exposure to a carcinogen may have learned this from hearing about the linearity assumption or seeing risk estimates projected from a linear model. Psychological and anthropological research also helps us understand the nature of the public's fear of exposure to toxic substances that are said (by scientists using a linear model) to be toxic at all levels. For example, Frazer (1959) and Mauss (1972) describe a belief, widespread in many cultures, that things that have been in contact with each other may influence each other through transfer of some of their properties via an "essence." Thus, "once in contact, always in contact," even if that contact (exposure) is brief. Rozin et al. (1986) show that this belief system, which they refer to as a "law of contagion," is common in our present culture. The implication of these notions is that even a minute amount of a toxic substance in one's food (e.g., a pesticide residue) will be seen as imparting toxicity to the food; any amount of a carcinogenic substance will impart carcinogenicity, etc. The "essence of harm" that is contagious is typically referred to as contamination. Being contaminated clearly has an all-or-nothing quality to it-like being alive or pregnant. When a young child drops a sucker on the floor, the brief contact with "dirt" may be seen as contaminating the candy, causing the par-
856
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Perceptions of Pesticides as Risks to Human Health
ent to throw it away rather than washing it off and returning it to the child's mouth. This all-or-nothing quality, irrespective of the degree of exposure, is evident in the observation by Erikson (1990) that "To be exposed to radiation or other toxins ... is to be contaminated in some deep and lasting way, to feel dirtied, tainted, corrupted" (p. 122). A contagion or contamination model is much more likely to hold in a world in which scientists use linear extrapolation to estimate risks than in a world that recognizes the beneficial effects of chemicals at low doses. We do not, for example, view ourselves as being "contaminated" by exposures to prescription drugs. Another relevant psychological tendency is to confound perception of risk with perception of benefit. If an activity or substance conveys some benefit upon us, we are likely to perceive it as less risky (Alhakami and Slovic, 1994) and more acceptable (Starr, 1969).
39.4 CONCLUSION In this brief overview, an attempt has been made to show the depth and complexity of the public's concerns regarding the risks from pesticides and other chemicals. These concerns transcend national borders and seem to have held remarkably constant for several decades, in spite of views of many toxicologists and other scientists that are quite the opposite from public views. Fortunately, the research described here will help us understand why public attitudes are the way they are and why they are so resistant to change. These attitudes are a complex product of human psychology and culture interacting with complex and idiosyncratic sciences such as toxicology, epidemiology, and risk assessment. One thing is clear. Risk communication efforts conducted by public relations specialists cannot turn public views around and may, in fact, exacerbate them. Some investigators have taken the limitations of risk science, the difficulty of creating and maintaining trust, and the subjective nature of risk judgments as signs pointing to the need for a radically different approach to dealing with conflicts regarding pesticides and other chemical products. This "new" approach focuses on introducing more public participation into both risk assessment and risk decision making to make the decision process more democratic, improve the relevance and quality of technical analysis, and increase the legitimacy and public acceptance of the resulting decisions. Work by scholars and practitioners in Europe and North America has begun to lay the foundation for improved methods of public participation within deliberative decision processes that include negotiation, mediation, oversight committees, and other forms of public involvement (English, 1992; Kunreuther et aI., 1993; National Research Council, 1996; Renn et aI., 1991, 1995). Those who are concerned about promoting rational decisions about the use of pesticides would be well advised to give careful consideration to this approach.
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Slovic, P., Flynn, l., Mertz, C. K., and Mullican, L. (1993). "Health Risk Perception in Canada." Report 93-EHD-170, Department of National Health and Welfare, Ottawa. Slovic, P., Kraus, N. N., Lappe, H., Letzel, H., and Malmfors, T. (1989). Risk perception of prescription drugs: Report on a survey in Sweden. Pharm. Med. 4, 43-65. Slovic, P., Malmfors, T., Krewski, D., Mertz, C. K., Neil, N., and Bartlett, S. (1995). Intuitive toxicology. n. Expert and lay judgments of chemical risks in Canada. RiskAnal. 15,661-675. Slovic, P., Malmfors, T., Mertz, C. K., Neil, N., and Purchase, 1. F. H. (1997). Evaluating chemical risks: Results of a survey of the British Toxicology Society. Hum. Exp. Toxicol. 16,289-304.
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Spigner, C., Hawkins, W., and Loren, W. (1993). Gender differences in perception of risk associated with alcohol and drug use among college students. Women and Health 20, 87-97. Starr, C. (1969). Social benefit versus technological risk. Science 165, 12321238. Steger, M. A. E., and Witt, S. L. (1989). Gender differences in environmental orientations: A comparison of publics and activists in Canada and the V.S. Western Political Quarterly 42,627-649. Stem, P. c., Dietz, T., and Kalof, L. (1993). Value orientations, gender, and environmental concern. Environ. Behav. 25, 322-348.
CHAPTER
40 Mammalian Toxicity of Microbial Pest Control Agents Andrew L. Rubin Department of Pesticide Regulations, California Environmental Protection Agency
40.1 INTRODUCTION Concern for the safety of agricultural workers and the public, as well as for the integrity of ecosystems, has fueled an interest in the use of microbes as pest control agents (Siegel and Shadduck, 1992). Following the registration in the United States of Bacillus popillae as an insecticide in 1948, the list of federally registered microbial pest control agents (MPCAs) has grown to include not only bacteria, but also fungi, yeasts, viruses, and protozoa. As of August 2000, there were 245-250 products listing approximately 60 different MPCAs as active ingredients under federal registration [R. Torla, U.S. Environmental Protection Agency (EPA), personal communication]. In California, 91 products containing 33 MPCAs were under active registration as of September 2000 [Department of Pesticide Regulation (DPR) registration database]. Although occupying only a small fraction of the current pesticide market, the number of pounds of MPCA active ingredients applied agriculturally nearly tripled in California between 1990 and 1999 (DPR, 2000) (Table 40.1). Bacillus thuringiensis (Bt) products overwhelmingly dominated this segment of the market, with much of the increase over the 1990-1999 period due to the use of B. thuringiensis o-endotoxins encapsulated in killed Pseudomonas jluorescens. Nonetheless, the use of MPCAs occupying a much smaller market fraction, notably Agrobacterium radiobacter, Ampelomyces quisqualis, Beauveria bassiana, Candida oleophila, Gliocladium virens, Lagenidium giganteum, Metarhizium anisopliae, Myrothecium verrucaria (killed), Pseudomonas jluorescens, and Streptomyces griseoviridis, also rose. This is evident in the greater than 12-fold increase in combined sales of products containing these ingredients in California between 1991 and 1998, reflecting both agricultural and nonagricultural uses (DPR sales database). Greater human exposure to these organisms under both occupational and nonoccupational scenarios is a reasonable expectation. Handbook of Pesticide Toxicology Volume 1. Principles
In this chapter, the current regulatory system in the United States for assessing the potential for toxicity, infectivity, and pathogenicity of MPCAs in humans is reviewed. In addition, toxicologic overviews for several prominent or proposed MPCAs are provided. These overviews are directed primarily at toxicity and pathogenicity issues arising from exposure to viable microbial organisms, though individual microbial toxins are considered in some cases. It is hoped that the reader will gain an appreciation for the unique problems confronting regulators as they assess the possibility of human health impacts resulting from the use of MPCAs.
40.2 TOXICITY TESTING REQUIREMENTS FOR MPCAS By 1981, with publication by the World Health Organization (WHO) of an approach to the safety testing of MPCAs (WHO, 1981), it was clear that differences between conventional chemicals and MPCAs required the development of a separate MPCA toxicity testing scheme. Shadduck (1983) outlined the premises upon which conventional chemical testing was (and is) based and provided reasons why these premises were not applicable to MPCAs. Siegel (1997) rearticulated Shadduck's arguments in the following manner. First, high doses of conventional chemicals generate biological effects which are expressed either as overt toxicity or as cellular or organ system responses designed to detoxify and excrete the chemical. With MPCAs it is often impossible to generate such effects without first killing the host by suffocation or by circulatory or gastrointestinal blockage. Second, metabolic and excretion pathways are often predictive of conventional chemical toxicity. In contrast MPCAs are not known to be degraded or genetically altered during passage through the host. Third, persistence and accumulation of chemicals within host organisms necessitates long-term testing for chronic effects. In general, MPCAs have not been shown to colonize mammals nor to produce chronic
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Copyright © 2001 by Academic Press. All rights of reproduction in any fonn reserved.
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Table 40.1 Total Pounds of MPCA Active Ingredients Applied per Year in California, 1990-1999 1 Year Active ingredient
1990
1991
1992
1993
1994
Agrobacterium radiobacter
0.10
0.00
0.41
2.20
3.84
Ampelomyces quisqualis
1995
1996
1997
1998
1999 6.44
6.03
14.0
27.8
19.8
0.42
2.59
9.05
39.7
17.9
1300
4890
2210
Bacillus sphaericus, serotype H-5A5B, strain 2362 Bacillus subtilis GB03
0.00
0.03
0.03
62,400
72,000
85,900
65,300
257
15,600
12,500
12,800
34,500
73,700
47,800
36,400
Beauveria bassiana, strain GHA
0.53
573
1240
915
Candida oleophila, isolate 1-182
0.00
305
103
55.0
Bacillus thuringiensis2
51,400
50,700
53,300
60,900
67,800
80,600
Bacillus thuringiensis, spp. kurstaki, genetically engineered
Encapsulated 8-endotoxin of Bacillus thuringiensis in killed Psuedomonas jluorescens 3
35.0
1820
7960
14,300
Codling moth granulosis virus
14,500
320
Gliocladium virens, GL-21 (spores)
15.3
144
156
104
86.1
Lagenidium giganteum (California strain)
86.9
151
0.10
134
859
499
Metarhizium anisopliae, var. anisopliae, strain ESFl
1.03
0.75
0.20
3.12
36.8
10.5
1100
8500
18,800
0.00
0.00
0.00
3640
3660
2100
Myrothecium verrucaria, dried fermentation solids and solubles Nosema [ocustae, spores
0.00
0.01
0.01
0.05
Pseudomonas jluorescens, strain A506
0.16
206
3040
Pseudomonas syringae, strain ESC-ll
34.00
Pseudomonas syringae, strain ESC-IO Streptomyces griseoviridis, strain K61
0.14
21.3
Trichoderma harzianum, Rifai strain KRL-AG2
Total pounds applied
51,400
50,700
55,100
68,900
82,200
95,800
15.4
0.01
0.05
1.42
1.74
4.90
1.90
64.5
39.2
60.3
122
100,000
169,000
166,000
139,000
1Data are from the Pesticide Use Report (DPR, 2000). The figures represent the amounts applied predominantly under agricultural conditions. Where no figure appears, there was no reported usage. Where "0.00" appears, it is the result of rounding off very small reported usages. 2The figures for B. thuringiensis include the following subspecies and strains: subspecies aizawai (GC-91 protein), subspecies aizawai (serotype H-7), subspecies aizawai [strain SD-1372, lepidopteran active toxin(s)], subspecies israelensis (serotype H-14), subspecies kurstaki (serotype 3A,3B), subspecies kurstaki (strain EG 2348), subspecies kurstaki (strain EG 2371), subspecies kurstaki (strain SA-l 1), subspecies kurstaki (strain BMP 123), subspecies kurstaki (strain HD-l), and subspecies san diego. 3The figures for encapsulated 8-endotoxin of B. thuringiensis in killed P. jluorescens include endotoxins from subspecies kurstaki and san diego.
effects. (Two caveats should be noted here. Some MPCAs are capable of persisting within mammals for longer than a few days without mUltiplying. This necessitates careful examination of their host clearance pattern, which would allow persistence to be distinguished from active infection (Siegel and Shadduck, 1990a). Also, viral agents targeted at mammalian pests present unique problems due to their mammalian host ranges (see the discussion of rabbit hemorrhagic disease virus, Section 43.3.3.2». Fourth, structure-activity relationships, which are often applicable to conventional chemicals, are not relevant to MPCAs.
The WHO testing scheme for determining the toxicity of MPCAs was based on four principles (WHO, 1981): (1) MPCAs pose "inherently different" risks to humans than conventional pesticides; (2) findings of minimal or no toxicity in laboratory testing ("negative results") are likely; (3) tiered testing, wherein negative results at one level preclude testing at higher levels, is appropriate; and (4) testing protocols should maximize the possibility of generating adverse effects in the host organism. This approach allowed for a more expedient registration process than that in effect for conventional chemicals. For the great majority of agents, which show negative results under the short-term
40.2 Toxicity Testing Requirements for MPCAs
Tier I requirements, longer term and more expensive studies were avoided. In 1982, the United States Environmental Protection Agency published its "Pesticide Assessment Guidelines, Subdivision M: Guidelines for Testing Biorational Pesticides" (U.S. EPA, 1982). This document, which was revised in 1989, incorporated WHO's philosophy and testing schemes into a system that remains in force today in the United States. A brief discussion of the MPCA testing requirements is presented in the following paragraphs. More complete treatments have appeared elsewhere (Betz et aI., 1990; McClintock, 1999; McClintock et aI., 1995; Siegel, 1997; Siegel and Shadduck, 1992). Few, if any, MPCA candidates have been carried beyond Tier I testing. 40.2.1 TIER I
Under Subdivision M, Tier I, three acute systemic exposure routes, oral, pulmonary, and intravenous (intraperitoneal for larger microbes), are required for unformulated MPCAs or for the technical grade active ingredient. Because aerosolization of viable microorganisms is problematic, intratracheal dosing is often used as a surrogate for pulmonary exposure by the inhalation route. Both the intravenous-intraperitoneal and the intratracheal routes are more invasive than those routes generally required for conventional chemicals, fulfilling the WHO principle of maximizing the likelihood of adverse effects. For each of these tests, between 107 and 108 colony-forming units (cfu), or the highest obtainable dose, is administered to mice or rats. The animals are monitored over a 4-week period for mortality, clinical signs of morbidity, body weight, gross pathology, and microbial clearance. Clearance is a measure of the ability of the host to remove invading microorganisms over time; as such, it is an indication of the presence or absence of an active infection (McClintock et aI., 1995). It is usually assessed by culturing the MPCA from homogenates of various organ systems, blood, and excretory products at established time intervals after dosing. Colony-forming units are enumerated and a pattern of clearance established. Although complete clearance can be demonstrated within a few days for most MPCAs (a result not unexpected because these organisms are rarely adapted for life under mammalian body conditions), clearance can take 50 days or more for persistent organisms. In such cases it is considered sufficient to demonstrate a clearance pattern and to show that the organism does not produce an active infection which can colonize and multiply within the host. In contrast to the first three routes, dermal toxicity test guidelines recommend use of the manufacturing use and formulated end use products, applied at 2 g per kilogram of body weight. This allows for an evaluation of the potential for local irritation, as well as that for systemic toxicity by this exposure route. As is often the case with conventional pesticidal products, it is the formulation components, not the active ingredients, which drive irritation reactions. In addition, there is another, though perhaps minor, consideration. Although it is unlikely that the vast majority of MPCA candidates would penetrate the dermal
861
barrier when applied as aqueous pastes or suspensions, such an event may occur when formulation components with high dermal permeability are also present. Tier I also requires testing for ocular irritation. As for dermal irritation, the formulation components are likely to play a primary role in this process; hence the requirement that the manufacturing and end use products be tested. Clearance determinations are not required for either dermal or ocular exposures. Nonetheless, putative MPCAs could conceive establish themselves at least temporarily in the eyes. Ocular applications of B. sphaericus and B. thuringiensis subsp. israelensis in the rabbit eye led to detections at that site for as long as 8 weeks postdosing (Siegel and Shadduck, 1990a). Cell culture tests are required only in the case of viral pest control agents; other classes of MPCAs are not likely to initiate infections of individual cells. A number of tests using both primary mammalian cell cultures and established mammalian cell lines are necessary to evaluate the potential toxicity and infectivity of the form of the virus considered to be most infective in susceptible cell cultures or in whole organisms (e.g., insects). These include a plating efficiency test, an infectivity evaluation, and a cell morphologic transformation assay. The latter assay, done specifically in Syrian hamster embryo cells, would be extended to an examination of viral tumorigenicity in hamsters if the MPCA proved capable of morphologically transforming cells in culture. Details of the specific cell culture tests are provided in McClintock (1999). Finally, Tier I requires testing for hypersensitivity if common use practices lead to repeated dermal or inhalation exposure. However, this requirement is generally waived because injection hypersensitivity studies using MPCAs are expected to be positive, whereas topical exposure studies are expected to be negative. Reports of hypersensitivity incidents occurring during manufacture or testing of MPCA products are nonetheless required (McClintock, 1999). 40.2.2 TIER 11
When acute toxicity is observed in the absence of pathogenicity and infectivity by the oral, dermal, or pulmonary exposure route, an acute LDso study is undertaken. Those exposure routes that produced toxicity in Tier I testing are used to establish the median lethal dose and slope after a 14-day postdose observation period. A subchronic study is required as well; the test article is administered daily for at least 90 days at a dose of at least 108 cfu/animal/day. The animals are monitored throughout for toxicity and pathogenicity-infectivity. In addition, organs, tissues, and body fluids are assayed for the presence of the microorganism. 40.2.3 TIER III
If significant toxicity in Tier 11 studies is observed, or if an ability to overcome natural host barriers to infection is detected, Tier III studies may be necessary for the registration process
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to be continued. Studies designed to detect effects on reproduction, fertility, oncogenicity, immunodeficiency, and primate infectivity may be called for based on the specific signs noted in Tiers I and 11. However, it is unlikely that organisms which show signs of significant toxicity or mammalian infectivity, defined as the ability to multiply within the host, will be carried beyond Tier I testing. Organisms targeted against mammalian pests clearly present a conundrum to this testing scheme, though it appears that none have yet been considered for registration in the United States.
40.3 TOXICITY OF INDIVIDUAL MPCAS The following discussion focuses on specific toxicity issues pertaining to selected prominent microbes which either are in use or have been considered for use as MPCAs. 40.3.1 BACTERIA 40.3.1.1 Bacillus Thuringiensis Bacillus thuringiensis is the most well-known and widely used of all pesticidal microbes. This gram-positive, spore-forming, facultative soil saprophyte was first registered for use in the United States in 1961 (U.S. EPA, 1998), after having been granted a temporary tolerance exemption for use in food and forage crops in 1958 (Fisher and Rosner, 1959). Bacillus thuringiensis subspecies have shown specificity against various orders of insects, including dipterans (B. thuringiensis subsp. israelensis), lepidopterans (B. thuringiensis subspp. kurstaki and aizawai), and coleopterans (B. thuringiensis subsp. tenebrionis). Recent B. thuringiensis isolates are active against nematodes, mites, and protozoa, as well as against other insect orders (Schnepf et aI., 1998). The entomopathogenic activity is primarily based on production during the stationary growth phase of a parasporal protein crystal. The crystal is composed of "Cry" (for "crystal") and, at least in B. thuringiensis subspp. israelensis and morrisoni (among currently commercially relevant B. thuringiensis subspecies) , "Cyt" (for "cytolytic") proteins. Knowledge of the identity, specificity, and structure of these "o-endotoxins" has expanded enormously over the past two decades. The coding sequences are known for over 100 of the relevant genes (Schnepf et aI., 1998). The Cry protoxin is activated by solubilization and proteolytic cleaveage under the alkaline gut conditions prevalent in susceptible insects. The resultant protein causes larval death by receptor-mediated lysis of the midgut epithelium (McClintock et aI., 1995; Schnepf et aI., 1998). Most Cry genes code for proteins in the 65-138 kDa range, with size at least partially dependent on the strain pathotype (Beegle and Yamamoto, 1992; Drobniewski, 1994). Differences in insect toxicity may be a function of different Cry solubilities in the insect gut, in addition to different inherent characteristics such as receptor affinity (Schnepf et aI., 1998). In some cases the spores can contribute to the insecticidal ac-
tivity of the parasporal crystal proteins, though how this comes about is unclear (Beegle and Yamamoto, 1992). The Cyt toxins are hemo1ytic and cytolytic proteins, with protoxin molecular weights in the 25-28 kDa range (Drobniewski, 1994). Cyt proteins do not exhibit sequence homology with Cry proteins (Hofte and Whiteley, 1989). They appear to disrupt insect cell membranes through detergent-like effects (Butko et al., 1997) and/or through the formation of cation-selective channels (Drobniewski, 1994). Thomas and Ellar (1983) found that intravenous injection of solubilized parasporal crystal proteins from B. thuringiensis subsp. israelensis was toxic to mice, in contrast to the lack of toxicity upon injection of a similar preparation from B. thuringiensis subsp. kurstaki. This was probably due to the presence of a 28-kDa Cyt protein in the former preparation. The toxicity of the isolated 28-kDa protein from B. thuringiensis subsp. israelensis was subsequently verified by intraperitoneal injection into mice (Mayes et aI., 1989). Interestingly, neither the solubilized kurstaki nor israelensis preparations provoked a toxic response in mice by the oral route (Thomas and Ellar, 1983). Several other B. thuringiensis molecules deserve mention. The ,B-exotoxin (thuringiensin), a heat-tolerant nonproteinaceous compound which is toxic to houseflies, mammals, and other nontarget organisms, has been demonstrated in B. thuringiensis subsp. thuringiensis and in one B. thuringiensis subsp. aizawai strain (U.S. EPA, 1998). ,B-Exotoxin works by inhibiting DNA-dependent RNA polymerase (Beegle and Yamamoto, 1992). For purposes of registration it is necessary to demonstrate the absence of ,B-exotoxin in B. thuringiensis formulations (McClintock et aI., 1995). In addition, a proteinaceous, heat-labile, insecticidal a-exotoxin with a molecular weight in the 45-50 kDa range has been identified (Beegle and Yamamoto, 1992). a-Exotoxin has properties similar to the 50-kDa enterotoxin of B. cereus (see Section 40.3.1.2). Finally, insecticidal activity can be enhanced by the expression of other proteinaceous toxins, among them phospholipases, proteases, chitinases, and secreted vegetative insecticidal proteins (Schnepf et aI., 1998). Extensive toxicologic testing of intact commercial B. thuringiensis strains has not resulted in appreciable toxicity, pathogenicity, or infectivity (U.S. EPA, 1998). In one early study, ingestion by humans of 3 x 109 B. thuringiensis spores/day (the subspecies was not identified) for 5 days, or by rats of 2 x 10 12 spores/kg produced no toxicity (Fisher and Rosner, 1959). Exceptions may occur when noncommercial subspecies, unusual exposure routes, or, as indicated previously, the use of isolated toxins as opposed to whole organisms are examined. Intracerebral exposure of laboratory rats is lethal if sufficient numbers of organisms are injected (Siegel and Shadduck, 1990b), though human exposure by this route is unlikely. Warren et al. (1984) reported an incident in which local and lymphatic inflammation requiring antibiotic therapy occurred when a laboratory worker sustained an accidental injection with spent medium containing B. thuringiensis subsp. israelensis and Acenitobacter calcoaceticus var. anitratus. In this case, A. calcoaceticus, a skin-dwelling bacterium, could have provided the proteases
40.3 Toxicity of Individual MPCAs
necessary to activate the protoxin, either by releasing them into the spent medium or into extracellular fluids at the injection site. It was, nonetheless, unclear to what extent the pathology was due to intoxication and to what extent it was due to bacterial persistence or infection with an accompanying inflammatory response from the host. Isolated health concerns pertaining to noncommercial B. thuringiensis strains have occasionally surfaced. Bacillus thuringiensis subsp. konkukian (serotype H34) was detected in the wounds of a French soldier injured by a land mine explosion (Hernandez et aI., 1998). The ability of B. thuringiensis subsp. konkukian (serotype H34) to cause tissue damage was demonstrated by cutaneous application of bacterial suspensions isolated from the wounds to normal and immunosuppressed mice. Inflammatory lesions developed in all mice treated with 107 cfu. These healed spontaneously in the normal animals, but progressed in the immunosuppressed animals. In a study reported in the Russian literature, B. thuringiensis subsp. galleriae was shown to cause syndromes in humans similar to those found in B. cereus-related food poisoning (Pivovarov et aI., 1977). Bacillus thuringiensis with cytotoxic characteristics similar to enterotoxin-producing B. cereus (see Section 43.3.1.2) was isolated from stools in a gastroenteritis outbreak occurring in a Canadian chronic care institution (Jackson et aI., 1995). Although the B. thuringiensis subspecies was not identified, the expression of B. cereus traits is not surprising, as some consider B. thuringiensis and B. cereus to be variants of the same species (Schnepf et aI., 1998). The production of Bacillus diarrheal enterotoxin was demonstrated in various commercial preparations of B. thuringiensis by Damgaard (1995), though the levels were generally low compared to those found in a reference culture of B. cereus that had been isolated from a food poisoning outbreak. It was noted, however, that a role for B. thuringiensis in food poisoning may be underestimated due to the requirement for a special staining technique to differentiate B. thuringiensis from the more conventionally assayed B. cereus. Under field conditions, reports of clinically significant symptoms in humans have been rare considering the length of time that B. thuringiensis has been in use as a pesticide. In one case, a farmer developed a corneal ulcer containing B. thuringiensis after being splashed in the eye with a commercial B. thuringiensis product (Samples and Buettner, 1983). A survey of farm workers exposed to commercial B. thuringiensis subsp. kurstaki sprays failed to identify clinical syndromes in the eye, respiratory tract, or skin (Bernstein et aI., 1999). However, positive skin-prick allergy tests and induction of IgG and IgE antibodies were documented in some exposed individuals, suggesting that allergenicity could result from repetitive exposure. Exposure of immunocompromised individuals may pose a special challenge. This was noted in an epidemiologic study conducted in an area of Oregon that had undergone spraying with B. thuringiensis subsp. kurstaki (Green et aI., 1990). Bacterial cultures from patients undergoing routine exams revealed 55 that were B. thuringiensis-positive. Bacillus thuringiensis was ruled out as a specific pathogen in 52 of those patients, but was not ruled out in the remaining 3 cases. Although all of these
863
latter infections may have been opportunistic, having occurred in people with established medical conditions, exacerbation of preexisting symptoms or development of new symptoms as a result of B. thuringiensis exposure was not ruled out. Despite the documented absence of mammalian toxicity in in vivo testing by most economically important B. thuringiensis strains, the ability of these bacteria to persist for extended periods within mammals after injection or intratracheal administration (McClintock et aI., 1995; Siegel and Shadduck, 1990b) has occasioned concern, particularly in light of the close phylogenetic relationship between B. thuringiensis and other medically significant Bacillus species such as B. cereus, B. anthracis, and B. sphaericus. For example, changes in the microenvironment could lead to increases in production of B. cereuslike enterotoxin. Siegel and Shadduck (1990a, 1992) discussed at length their operational definition of infectivity, which requires demonstration of multiplication of the microorganism within the host and disruption of functional or structural homeostasis. This contrasts with their definition of persistence, in which an organism is present either in a multiplying or quiescent state, but does not disrupt the host. Bacillus thuringiensis generally shows a consistent, though in some cases prolonged, clearance pattern (McClintock et aI., 1995). However, in one study using B. thuringiensis subsp. israelensis, splenic bacterial counts following intraperitoneal injection into mice showed no tendency to decline even after 80 days (Siegel and Shadduck, 1990a). This was interpreted as evidence that multiplication had occurred within the host, though there was no sign of toxicity. Combined with the record of safe use, a conclusion that persistence in animals does not translate into toxicity is warranted. 40.3.1.2 Bacillus Cereus Bacillus cereus is a gram-positive, catalase-positive, rod-shaped saprophyte that is closely related to, if not con specific with, B. thuringiensis. Bacillus cereus does not produce the parasporal inclusion body which is so important to the entomopathogenic activity of B. thuringiensis. One B. cereus strain, BP01, is currently registered in the United States as a plant growth regulator. Bacillus cereus is interesting from a medical standpoint because it is implicated in pathologies of the lung, ear, eye, gall bladder, and urinary tract and as an opportunistic invader in trauma or disease cases (Drobniewski, 1994; Goepfert et aI., 1972). It also is reported to have been an opportunistic pathogen in several cancer patients (Banerjee et aI., 1988) and is an epidemiologically significant food-based pathogen. Two food poisoning syndromes, a diarrheal type, associated with consumption of a diversity of food types, and an emetic type, most commonly associated with rice and pasta consumption, are caused by B. cereus (Drobniewski, 1993; Logan and Turnbull, 1999). A heat-labile enterotoxin complex, consisting of two or three protein components with molecular weights of 38-57 kDa, appears to be responsible for the diarrheal symptoms. A heatstable 10-kDa peptide is associated with the emetic symptoms.
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Serotyping has identified the B. cereus strains most likely to produce the two syndromes. Other factors, notably hemolysins and phospholipases, are involved in the establishment of local and systemic infections (Drobniewski, 1993). In view of the medical and epidemiologic significance of B. cereus, vigilance with respect to the dissemination and application of pesticidal products containing this organism is warranted, as well as strict attention to proper strain identification. Demonstration of an inability to produce enterotoxins or emetic toxins also may be useful in the assessment of potential risks posed by proposed commercial B. cereus strains. 40.3.1.3 Bacillus Sphaericus
Several isolates of B. sphaericus, a crystalliferous spore-former similar to B. thuringiensis, have shown promise as mosquito larvicides (Saik et aI., 1990), making them potentially useful for controlling tropical diseases such as malaria and filariasis (Murthy, 1997). However, B. sphaericus has provoked human health concerns. It was implicated in several cases of meningitis (AlIen and Wilkinson, 1969; Siegel and Shadduck, 1990b), in the formation of a lung pseudotumor (Isaacson et aI., 1976), and as an opportunistic pathogen in a cancer patient (Banerjee et aI., 1988). Interestingly, B. sphaericus isolates from the meningitis patients and from the patient with the pseudotumor did not prove pathogenic in rabbits and mice exposed intraperitoneally or intravenously, or in mice after intracerebral exposure (Siegel and Shadduck, 1990b), raising questions about the applicability of animal testing to human pathology in this case. It is worth noting, however, that B. sphaericus is a species complex, divisible into six groups based on DNA homology, with larvicidal strains clustering in homology group IIA (Krych et aI., 1980). Even among entomocidal strains, the particular B. sphaericus strain designation, exposure route, and host animal strain continue to be relevant when analyzing the toxicity of these organisms in mammalian systems. A more detailed examination of two particular studies may serve to illustrate this point. Shadduck et al. (1980) investigated the pathogenicity in mice, rats, and rabbits of three entomocidal B. sphaericus strains, SSII-1, 1404-9, and 1593-4, after exposure by various routes. Neither death nor systemic illness resulted under any exposure scenario. Conjunctival instillation into rabbits of up to 109 infectious units (iu) of B. sphaericus caused local lesions which were severe at the higher doses. Less severe lesions developed after instillation of autoclaved bacteria, suggesting that a portion of the pathologic response was attributable to the presence of heat-stable foreign material. Generally mild brain lesions occurred in some rats upon intracerebral injection of all three strains. In strain SSII -1 this occurred at a dose as low as 1.2 x 106 iu per rat. Intracerebral hemorrhage was commonly noted in mice upon intracerebral injection of ~3 x 108 iu of any of the three strains, though rabbits similarly treated with 1593-4 did not react in this way. Subcutaneous injection of strain 1404-9 resulted in an abscess in one of five mice exposed to the highest dose, 6.7 x 109 iu per animal. Intraperitoneal injection into rats with 3.2 x 108 iu of strain 1404-9 or
4.7 X 108 iu of strain 1593-4, respectively, were without effect. Viable bacilli were detected in eye cultures 10-14 days after conjunctival instillation into rabbits of as few as 1.2 x 103 iu of strain SSII-1. Persistence in the eye was also detected with injections of 109 iu of strain 1404-9 and 108 iu of strain 1593-4, though lower doses were not tested. Similarly, bacilli were detected in brain cultures 10-14 days after intracerebral injections of as few as 108 iu per rat of strain 1593-4, 1.2 x 108 iu of strain SSII -1, and 3.2 x 108 iu of strain 1404-9. Lower concentrations were not tested for strains 1404-9 or SSII-1. These results indicate that the bacterium can persist in mammals, at least under certain conditions. Whether or not bacterial replication occurred was not explored with respect to the ocular exposure route. However, a pattern of cerebral clearance was established after intracerebral injection of 5.5 x 105 iu of strain 1593-4 into rats. Detections of greater than 600 iu per 100 mg wet brain tissue were noted on day 3 postinjection. No more than 30 iu per 100 mg tissue were observed at various times between days 5 and 12. The brains were considered sterile by day 14. In a follow-up study, conjunctival instillation into rabbits of 4.48 x 108 cfu 1 B. sphaericus strain 2362 did not cause local toxicity (Siegel and Shadduck, 1990a), contrasting with the results of the earlier study using other strains. Nonetheless, culturable bacilli were recovered 8 weeks after treatment (longer recovery times were not examined). A clear pattern of splenic clearance was established after intraperitoneal injection of 1.2 x 107 cfu into mice, though some bacilli (165 cfu per gram of spleen) were still present at study termination on day 67. Intraperitoneal injection of 8 x 108 cfu resulted in the death of 42 of 49 mice within 24 hours. Interestingly, injection of 3.8 x 108 cfu per animal autoclaved strain 2362 resulted in the deaths of 3/6 mice between 24 and 48 hours postinjection. These results were surprising in light of the observation of no toxicity by the intraperitoneal route of B. sphaericus strains SSII-1, 1404-9, and 1593-4 in the earlier study (Shadduck et aI., 1980). Production of a soluble extrabacterial toxin was ruled out as a cause because passage through cellulose acetate filters removed the toxicity. The authors speculated that the strain of mouse (outbred CD-I) could have been particularly sensitive to the effects of strain 2362, citing a study from a different laboratory showing no effect of this strain in Swiss mice. Alternatively, they considered that the particular culture conditions of the bacterium, which they did not know, could have generated a more lethal bacterial isolate. In any case, they noted that these were extremely high doses which are not likely to be relevant to human exposures in the field. The possibility that B. sphaericus has unique pathogenic properties (e.g., it may be infective only in health-compromised individuals) which are not evident in conventional animal testlThe designation "colony forming units" (cfu) was used in the later study (Siegel and Shadduck, 1990a) to signify that the bacterial titers were quantitated by culturing appropriate dilutions of the inocula in agar and counting the resultant bacterial colonies at a later time. "Infectious units" (iu) was used in the earlier study (Shadduck et aI., 1980) because titers were quantitated by serial tube dilution; determination of infectious units was by turbidity in brain infusion broth.
40.3 Toxicity of Individual MPCAs
ing cannot yet be excluded. Certainly the direct impacts of entomocidal B. sphaericus strains on human populations should be monitored until sufficient evidence for their safety is obtained under conditions of actual pesticidal use. 40.3.1.4 Burkholderia Cepacia
A gram-negative, nutritionally versatile, and highly antibioticresistant bacterium, B. cepacia has stimulated economic interest due to its ability to inhibit soil-borne plant pathogens and to degrade hydrocarbons associated with sites of environmental contamination (Parke, 1998). It also causes serious opportunistic infections in humans suffering from chronic granulomatous disease and cystic fibrosis (CF) (Butler et al., 1995). As many as 40% of patients in some CF centers develop B. cepacia infections, with 35% of those patients exhibiting "cepacia syndrome" characterized by grave pulmonary pathogenesis, bacteremia, and death (Holmes et al., 1998). Other figures are perhaps less alarming, but serious nonetheless. LiPuma (1998) cited respiratory culture results from 1996 showing that 3.6% of cystic fibrosis patients showed evidence of infection. Of those infected, 20% "succumb to a rapidly progressive necrotizing pneumonia." Burkholderia cepacia also is implicated in nosocomial infections of non-CF patients, as well as in the "foot rot" syndrome experienced by soldiers in swampy terrain (Holmes et al., 1998). The primary mode of transmission to CF patients appears to be from other CF patients, though transmission from non-CF patients is also possible. Social isolation measures have been necessary in some circumstances (WaIters and Smith, 1993), but these have been accompanied by poor psychosocial outcomes (Butler et al., 1995; LiPuma, 1998). Burkholderia cepacia strains are divided among five genomovars, collectively known as the B. cepacia complex. According to the analysis of Vandamme et al. (1997), genomovars 11 and III are the best represented among CF patients. However, they caution that a systematic study of genomovar distribution has not yet been done, nor is the relative significance for cepacia syndrome of the various strains yet understood. Representatives of all genomovars have been detected in CF patients (Vandamme et al., 1997). Attempts have been made to identify markers associated with epidemic transmission of B. cepacia among CF patients. Such markers could simplify the risk assessment process for proposed B. cepacia pesticidal strains. Cable pili, peritrichous appendages that facilitate binding to CF mucin and airway epithelial cells (Goldstein et al., 1995; Sajjan and Forstner, 1993), may comprise one marker phenotype. These structures were identified in an epidemic strain transmitted among CF patients in clinics in Toronto and Edinburgh (Sun et al., 1995). Another candidate, a l.4-kb DNA fragment known as the "B. cepacia epidemic strain marker" (BCESM), was identified in seven epidemic strains of the bacteria, but was not present in nonepidemic strains (Mahenthiralingam et al., 1997). Unfortunately, the presence of neither BCESM nor of cable pili is considered a certain indicator of transmission potential. The cable pilus gene was detected in only one of the seven epidemic strains
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examined in the Mahenthiralingam study. And as pointed out by LiPuma (1998), the presence of nonepidemic strains in the respiratory tracts of CF patients demonstrates their colonizing ability even as they lack BCESM. In addition to these markers, other virulence factors and epidemic strain-associated markers have been considered. At this point, it appears that no single marker will provide absolute predictive ability for clinically important B. cepacia strains. However, they may provide some initial screening capability when considering potential pesticidal strains. George et al. (1991) investigated the effect on CD-1 mice of intranasal instillation of 5.3 x 108 cfu B. cepacia strain AC1100. Ruffled fur, weight loss, and inactivity were noted during the first two days following treatment, but recovery was evident thereafter. Increased lung weights, apparent between days 2 and 14 (study termination), were attributed to macrophage influx and endotoxin-mediated edema. Declining numbers of B. cepacia were evident through day 7 in the lungs and through day 2 in the nasal cavity. Burkholderia cepacia was also present in the gastrointestinal tract for the first two days after treatment. This was attributed to mucociliary evacuation from the lung to the mouth and thence to the stomach, with bacterial survival afforded by the mucus coating acquired in the respiratory system. In a later study (George et al., 1999), endotoxinresistant C3H1HeJ mice were subjected to intranasal instillation with B. cepacia strain ATCC 25416. Lethality was observed at as low as 2.2 x 108 cfu per mouse, with an approximate LDso of 7 x 108 cfu per mouse. Although this appears inconsistent with the relative lack of effect in mice seen at a similar dose in the 1991 study, the considerable splay evident in the mortality data should be recognized, as well as the fact that different bacterial and mouse strains were used. Other differences from the earlier study were evident. In the 1999 study, no changes in lung weights were detected through 14 days posttreatment, possibly reflecting the endotoxin-resistant status of the mouse strain used in that study or the lower bacterial dose applied (~107 cfu per mouse vs. 5.3 x 108 cfu per mouse in the earlier study). Also in contrast to the 1991 study, B. cepacia appeared to be stably established in the lungs, small intestine, cecum, and large intestine through study termination on day 14. Moreover, B. cepacia was cultured from the liver and spleen at 3 hours, and from the mesenteric lymph nodes through 10 days. These data imply that, under some circumstances, B. cepacia could gain a more tenacious hold in a mammalian system. Subdivision M testing of B. cepacia isolate M36, submitted for registration as an active ingredient to the State of California, indicated that clearance had occurred through the feces within 7 days of oral dosing of rats with 2.85 x 108 cfu. However, severe fibrous adhesions between pleural surfaces of the thoracic cavity, lungs, and pericardial sac were noted in 1/3 males and grey lung coloring in 2/3 females (DPR, 1994a). Intratracheal dosing of 1.9 x 108 cfu per rat resulted in pulmonary clearance by day 22, with lung discoloration evident through day 8 (DPR, 1994b). Infectivity of this strain by all routes of exposure was discounted because there was no evidence of multi-
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CHAPTER 40 Mammalian Toxicity of MPCAs
plication. Significantly, use of M36 has been eliminated due to the presence of BCESM. Use of other pesticidal strains currently are restricted to agricultural applications such as seed treatment and drip irrigation to minimize exposure to susceptible populations (D. Gurian-Sherman, U.S. EPA, personal communication). Finally, because of its large and highly adaptable genome, there is concern that pathogenic strains of B. cepacia could be generated through gene transfer or recombination if large numbers of putative nonpathogenic organisms are artificially introduced into the environment (Holmes et aI., 1998). For this reason, cystic fibrosis advocacy groups have expressed serious misgivings about the registration of B. cepacia products before adequate testing and assurance of nontransformability is available (PTCN, 1997).
40.3.2 FUNGI
40.3.2.1 Metarhizium Anisopliae In use since the late 1800s, M. anisopliae is a Deuteromycete (Fungi imperfecta) used in the United States largely for the control of cockroaches, though it also is effective against other orthopterans and against coleopterans. It has a wide geographic distribution, existing as an insect or nematode parasite, or in various soils, sediments, spoil heaps, and other environments (Domsch et aI., 1980). Death of the host insect results when contact with conidia, the environmentally stable asexual spore stage, leads to infection. This is followed by enzyme-mediated exoskeletal degradation, mycelial development, and sporation (Ward et aI., 1998). Massing of conidia in affected insects lends a characteristic green color, hence the name "green muscardine" for the insect disease (Ferron, 1981). Insecticidal activity also may reside in a family of cyclodepsipeptides known as destruxins, 15 of which were identified in M. anisopliae as of 1989 (Gupta et aI., 1989), as well as in other toxic substances (Domsch et aI., 1980). Standardized laboratory studies have not demonstrated toxicity or infectivity of M. anisopliae in rats, mice, or rabbits, though persistence without multiplication was reported (Siegel and Shadduck, 1990b). Allergenicity was demonstrated in mice following intraperitoneal sensitization and intratracheal challenge with crude protein extracts of mycelia and conidia (Ward et aI., 1998). Although applicability of the mouse model to humans exposed solely via inhalation was not experimentally addressed, evidence for allergy under occupational situations has been reported (Kaufman and Bellas, 1996). A single case of keratomycosis, an increasing problem with fungi in general due to increased use of antibacterial drugs, immunosuppressants, and corticosteroids (Ishibashi et aI., 1986), was reported in an 18-year old man (Cepero de Garcia et al., 1997). A single case of hyphomycotic rhinitis was reported in a cat (Muir et aI., 1998). Disseminated infection with severe morbidity was reported in an immunocompromised child (Burgner et aI., 1998), highlighting the need for great caution where exposures of sick individuals are possible.
40.3.2.2 Beauveria Bassiana Beauveria bassiana, a Deuteromycete long known for its entomopathogenic properties, causes an insect disease known as white muscardine. The organism produces a number of cyclodepsipeptides such as beauvericin which may account for at least part of its insect toxicity (Miller et al., 1983). Beauveria has been used as a medicant in Japan for over a millenium (Ignoffo, 1973). Allergic responses have been reported in humans following inhalation of spore preparations, though repeated handling of cultures did not reveal adverse effects in another study (Ignoffo, 1973). A study from China noted hypersensitivity-like pulmonary reactions in mice and rats after a single exposure to B. bassiana. However, the low room temperatures may have constituted a significant stress to the animals (Song et al., 1989, cited in Semalulu et al., 1992). Russian investigators have reported the LDso to be greater than 1.1 x 1010 and greater than 2.2 x 1010 fungal cells in albino rats exposed intragastrically and intraperitoneally, respectively, and greater than 4 x 1010 fungal cells in rabbits exposed intravenously (Mel'nikova and Murza, 1980). Beauveria bassiana has been implicated in at least two cases of keratomycosis, though both patients had long histories of antibiotic and corticosteroid use (Ishibashi et al., 1986). A study in the French literature isolated organisms from the genus Beauveria from an individual with bronchopneumonia (Freour et al., 1966, cited in Semalulu et al., 1992). Direct inoculation into rabbit corneas of B. bassiana isolated from a patient with keratitis resulted in inflammation, corneal ulcers, corneal haze, injection of the iris, and sparse-to-moderate fungal growth in the cornea, though the severity was less than that seen in parallel eyes treated with Candida albicans and tended to resolve itself with time (Ishibashi et al., 1986). Injection of B. bassiana into the quadriceps muscles of CD-1 mice led to focal muscle necrosis, edema, and inflammation, with the severity of the responses dependent on the number of organisms injected (Semalulu et aI., 1992). Muscle regeneration was visible by 7 days. Viable spores capable of initiating colonies in artificial media were not detected after 3 days. Thus it is unlikely that this organism, which does not grow at temperatures above 32 QC, can infect or colonize mammals under normal circumstances.
40.3.2.3 Gliocladium Wrens Gliocladium virens is a common, soil-dwelling saprophyte that is useful in controlling Pythium ultimum and Rhizoctonia solani, organisms that cause damping-off disease in greenhouses. Such antifungal activity may be due partially to production of gliotoxin, a relatively nonselective antibiotic that is also immunosuppressive and moderately toxic to mammals'(Lumsden and WaIter, 1995). However, studies using G. virens isolate GL-21 did not reveal unusual toxicity, though conjunctival irritation for at least 72 hours in the rabbit eye irritation study and death by capillary obstruction in the rat intravenous study were noted (DPR, 1993). The potential for G. virens-induced allergenicity is not known.
40.3 Toxicity of Individual MPCAs
40.3.2.4 Lagenidium Giganteum The ability of this aquatic saprophyte to parasitize, and eventually kill, mosquito larvae demonstrates its potential usefulness in mosquito control programs. Conventional toxicity testing with an organism as large as L. giganteum, which can produce cells greater than 200 ).tm in length, was impossible because intratracheal instillation of as few as 1.16 x 105 oospores per rat resulted in the prompt death of many of the treated animals from acute pneumonia, airway obstruction, or severe inflammation (Siegel and Shadduck, 1987). A similar result was obtained upon intravenous injection of 1.78 x 106 cfu into mice, where embolism killed several animals within 36 hours of treatment (Kerwin et al., 1990). Lowering the numbers either of active or autoclave-inactivated organisms still resulted in tissue damage after intratracheal or intraperitoneal exposures (Siegel and Shadduck, 1987). In all likelihood, these lesions represented local inflammatory responses to large amounts of foreign biological material. Although there was histologic evidence for persistence, multiplication within the mammalian hosts did not occur. Oospore treatment of rat skin or rabbit eyes did not result in irritation. The potential for allergenicity remains untested. 40.3.3 VIRUSES Interest in the use of viruses as pest control agents is based on their promise as specific vectors, either in native form or as genetically engineered constructs designed as delivery agents for pesticides or as immunocontraceptives for mammalian wildlife control. Establishment of the range of target species susceptible to infection by a given virus is perhaps the major issue in the assessment of risks associated with use of that virus as a pest control agent. For obvious reasons, human health concerns are magnified when the virus in question is infective in, or pathogenic toward, mammalian species. 40.3.3.1 Baculoviruses Baculoviruses are double-stranded DNA viruses that are highly specific pathogens of insects and other arthropods (Huber, 1995). Two major groups of baculoviruses, the nucleopolyhedrosis viruses (NPVs) and the granuloviruses (GVs), are enveloped by proteinaceous occlusion bodies that protect the virions from adverse environmental impacts (Saik et al., 1990). Infection in susceptible insects occurs when the occlusion bodies are ingested from leaf surfaces and dissolved under the alkaline conditions of the insect gut (Black et aI., 1997). Death of the host results from the multiplication of freed virions within bodily tissues (Huber, 1995). Based on studies both in humans and in laboratory animals, human safety concerns appear to be minimal. Dietary consumption over a 5-day period of 6 x 109 Helicoverpa zea NPV polyhedra was without effect (Heimpel and Buchanan, 1967), as was occupational exposure to the same organism over a 26month period in a virus production facility (Huang et aI., 1977). Exposure of mice and guinea pigs to H. zea NPV inclusion
86"
bodies, virus rods, and polyhedral protein by several routes (inhalation, oral gavage, and intradermal, intraperitoneal, intracerebral, and intravenous injection) revealed no effects (Ignoffo and Heimpel, 1965). Similar negative results were obtained using the Trichoplusia ni NPV (Heimpel, 1966). No unusual responses were detected in rats following acute subcutaneous injection of 1.2 x 109 polyhedral inclusion bodies (PIB) of H. zea NPV into neonates or acute intravenous injection of the same quantity into adults (Bames et aI., 1970). In addition, no effects were identified with dietary exposures of 90 days or 2 years using feed preparations containing viral loads between 6 x 107 and 6 x 109 PIB per 100 g. Allergenic responses were not detected in guinea pigs after 3 weeks of inhalation exposure at 1 hour/day, 5 days/week, to H. zea NPV free viral rods (obtained from 3 x lOll PIB/day) or intact PIB (3 x lOll/day), or after dermal exposure to 1.5 x 1011 PIB for 5 days followed by intradermal injections of 1.2 x 108 PIB in each of four bodily sites (Meinecke et aI., 1970). Pigs force-fed with Mamestra brassicae larval NPV at a dose level of 5 x 107 polyhedra per gram of body weight showed a slight, transitory rise in body temperature (Doller et aI., 1983). There was, however, no evidence for lymphatic involvement, no effect on leukocyte counts, and no evidence for viral replication or organ infection within the hosts. In one accounting published in 1965, over 26 baculoviruses had been tested in 10 mammalian species without indication of toxicity (Ignoffo and Heimpel, 1965). Indeed, baculoviruses do not appear capable of multiplying within mammalian hosts (Black et aI., 1997). Finally, the ubiquity of baculoviruses in the human food supply attests to their harmlessness (Black et al., 1997; Heimpel et aI., 1973). 40.3.3.2 Rabbit Hemorrhagic Disease Virus Rabbit hemorrhagic disease virus (RHDV) is a single-stranded RNA virus belonging to the calicivirus family. It is linked to a syndrome of necrotising hepatitis, hemorrhage, and death in European rabbits (Oryctolagus cuniculus) that has become known as rabbit hemorrhagic disease. This disease was first recognized in Angora rabbits exported from Germany to China in 1984. It subsequently spread from China to Korea, Europe (including the British Isles), Mexico, Israel, and North Africa (Nowotny et aI., 1998). Efficacy of the virus is not completely understood and may depend on such host factors as rabbit density, viral resistance, and age distribution, as well as nonhost factors, including the presence or absence of vertebrate or invertebrate vectors and the quality of the virus preparations used (Parkes et aI., 1999). Nonetheless, RHDV has been purposefully spread in Australia, where introduction of European rabbits in the 19th century left the island continent with a monumental and ongoing rabbit infestation. The rabbits' prodigious appetite, burrowing practices, and reproductive capacity have wreaked ecological and agricultural devastation, with serious consequences for the Australian economy. New Zealand has also considered using RHDV to control its own rabbit infestation, though importation was prohibited in 1997 due to uncertainties surrounding the epidemiology and efficacy of the virus
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Mammalian Toxicity of MPCAs
[Ministry of Agriculture and Forestry (MAF), 1997]. Despite this ruling, the virus became established in New Zealand by virtue of illegal importation and release (Sissons and Grieve, 1998). Subsequent, deliberate spread by farmers, although initially of unclear legality, eventually was legalized through an amendment to the law. Recent regulations provide for the importation of the virus, as long as it is done with appropriate permits (J. Parkes, Landcare Research, New Zealand, personal communication). Following government-sanctioned tests of the efficacy of rabbit-to-rabbit transmission on an uninhabited Australian island in 1995, RHDV jumped to the mainland by an unknown mechanism. This led to the approval of hundreds of deliberate releases on the mainland, and brought into focus the issue of human safety in the affected areas. The wide host ranges of other calicivirus types, along with the apparent virulence of some caliciviruses in humans [representatives of four of five calicivirus categories are considered to be human pathogens (Smith et al., 1998)], and the high mutation rates characteristic of RNA viruses in general, militate against premature judgements on the issue of possible direct human impacts. Claims of high RHDV specificity toward European rabbits are based primarily on serologic analysis of blood from various experimentally infected species [Bureau of Resourse Sciences (BRS), 1996]. These studies have been criticized for applying an inappropriately high standard for a positive judgement of infection. It has been argued that the data actually support the opposite conclusion, that many species do indeed show evidence of infection (Smith, 1998). Ultimately, when dealing with serologic data, rigorous tests will be required to distinguish active infection from simple exposure. In any case, regardless of controversies over the extent of the RHDV host range, or whether there is a credible risk to humans, it seems clear that the rapid spread of the disease puts rabbit populations at risk in areas where eradication of those rabbits may not be desired. Where the possibility of human infectivity or illness resulting from contact with infected rabbits has been examined directly, divergent conclusions have been drawn, even when they are based on the same data set. Mead et at. (1996), in a report to the Australian government, as well as Carman et at. (1998) in the follow-up report in the open literature, found no evidence for infection or for symptoms of disease in over 250 people, many of whom had high exposure to the virus through direct handling of diseased rabbits. Smith et at. (1998), using the same data but categorizing putative exposure levels differently, saw correlations between exposure and incidence of a number of pathologies, including flu or fever, diarrhea or gastroenteritis, neurologic symptoms, rashes, and hepatitis. Matson (1998) also differed from Mead in his analysis of the human serologic data provided in the original Mead report, seeing plausible evidence for infection in some people from South Australia who had handled infected rabbits. Because of these conflicting data and their interpretations, clarification of the issue of RHDV specificity and safety to humans will depend on the rigorous analysis of well-planned future studies, both of an epidemiologic and experimental nature.
40.4 GENERAL CONCLUSIONS The promise of microbial pest control agents resides in their host organism specificity and in their relatively benign ecosystem and human health impacts. The requirements for toxicity testing of these agents in the United States under Subdivision M of the Federal Insecticide, Fungicide, and Rodenticide Act are designed to provide a fast and efficacious means to identify problematic MPCAs, while moving the rest toward registration. In general, this approach appears to have functioned well. Nonetheless, as is clearly recognized in those guidelines, MPCAs pose unique challenges to humans. They are, after all, alive, and thus carry at least a theoretical potential for adaptation and survival in novel microenvironments. It is not inconceivable that hazards to humans posed by living organisms could be missed in animal studies. For example, it might be argued that particular attention must be paid to the welfare of sensitive human subpopulations such as those who are diseased or immunocompromised, or who might be allergic to specific microorganisms. Attention to strain type is also very important, particularly when the microbe in question belongs to a medically significant species or genus. In the final analysis, continued monitoring of the health effects of MPCAs under conditions of actual pesticidal manufacture and use is warranted to ensure the safety of this interesting and viable approach to pest control.
ACKNOWLEDGMENTS For helpful discussions and comments on the manuscript, the author thanks Drs. Derek Gammon, Doug Gurian-Sherman, J. Thomas McClintock, John Parkes, Joel Siegel, and Alvin Smith. For help gaining access to the Pesticide Use Report of the Department of Pesticide Regulation (State of California), the author thanks Dr. Larry Wilhoit.
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Mammalian Toxicity of MPCAs
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Saik, J. E., Lacey, L. A., and Lacey, C. M. (1990). Safety of microbial insecticides to vertebrates---Domestic animals and wildlife. In "Safety of Microbial Insecticides" (M. Laird, L. A. Lace, and E. W. Davidson, eds.), pp. 115132. CRC Press, Boca Raton, PL. Sajjan, U., and Forstner, J. (1993). Role of a 22-kilodalton pilin protein in binding of Pseudomonas cepacia to buccal epithelial cells. Infect. Immun. 61, 3156--3163. Samples, J. R., and Buettner, H. (1983). Corneal ulcer caused by a biologic insecticide (Bacillus thuringiensis). Am. 1. Ophthalmo!. 95,258-260. Schnepf, E., Crickmore, N., van Rie, J., Lerecius, D., Baum, J., Feitelson, J., Zeigler, D. R., and Dean, D. H. (1998). Bacillus thuringiensis and its pesticidal crystal proteins. Microbio!' Mol. BioI. Rev. 62, 775-806. Semalulu, S. S., MacPherson, J. M., Scheifer, H. B., and Khachatourians, G. G. (1992). Pathogenicity of Beauveria bassiana in mice. 1. Vet. Med. B 39, 81-90. Shadduck, J. A. (1983). Some considerations on the safety evaluation of nonviral microbial pesticides. Bull. World Health Organization 61,117-128. Shadduck, J. A., Singer, S., and Lause, S. (1980). Lack of mammalian pathogenicity of entomocidal isolates of Bacillus sphaericus. Environ. Entomol. 9,403-407. Siegel, J. P. (1997). Testing the pathogenicity and infectivity of entomopathogens to mammals. In "Manual of Techniques in Insect Pathology" (L. A. Lacey, ed.), pp. 325-336. Academic Press, San Diego. Siegel, J. P., and Shadduck, J. A. (1987). Safety of the entomopathogenic fungus Lagenidium giganteum (Oomycetes: Lagenidiales) to mammals. 1. Econ. Entomol. 80(5),994--997. Siegel, J. P., and Shadduck, J. A. (1990a). Clearance of Bacillus sphaericus and Bacillus thuringiensis ssp. israelensis from mammals. 1. Econ. Entomol. 83,347-355. Siegel, J. P., and Shadduck, J. A. (1990b). Safety of microbial insecticides to vertebrates-humans. In "Safety of Microbial Insecticides" (M. Laird, L. A. Lacey, and E. W. Davidson, eds.), pp. 101-113. CRC Press, Boca Raton, PL. Siegel, J. P., and Shadduck, J. A. (1992). Testing the effects of microbial pest control agents on mammals. In "Microbial Ecology. Principles, Methods, and Applications" (M. A. Levin, R. J. Seidler, and M. Rogul, eds.), pp. 745759. McGraw-Hill, New York. Sissons, c., and Grieve, J. (1998). Introduction. In "Rabbit Control, RCD: Dilemmas and Implications, Proceedings of the Rabbit Control, RCD: Dilemmas and Implications Conference," Wellington, New Zealand, 30--31 March 1998 (compiled by B. D. W. Jarvis), pp. v-vi. New Zealand Association of Scientists, supported by The Royal Society of New Zealand. Smith, A. W. (1998). Calicivirus models of emerging and zoonotic diseases. In "Rabbit Control, RCD: Dilemmas and Implications, Proceedings of the Rabbit Control, RCD: Dilemmas and Implications Conference," Wellington, New Zealand, 30--31 March 1998 (compiled by B. D. W. Jarvis), pp. 37-42. New Zealand Association of Scientists, supported by The Royal Society of New Zealand. Smith, A. w., Skilling, D. E., Cherry, N., Mead, J. H., and Matson, D. O. (1998). Calicivirus emergence from ocean reservoirs: Zoonotic and interspecies movements. Emerging Infectious Diseases 4, 13-20. Song, J. v., Houng, S. M., Lin, G. H., Lou, C. Z., Hou, M. S., Tzan, L. F., and Chang, O. Z. (1989). Experimental study of farmers' lung-like lesions caused by Beauveria bassiana. Chung-hua-Ping-Li-Hsuch-Tse-Chin 18, 111-114. Sun, L., Jiang, R.-Z., Steinbach, S., et al. (1995). The emergence of a highly transmissible lineage of cbl+ Pseudomonas (Burkholderia) cepacia causing CF centre epidemics in North America and Britain. Nature Med. 1,661. Thomas, W. E., and Ellar, D. J. (1983). Bacillus thuringiensis var. israelensis crystal o-endotoxin: Effects on insect and mammalian cells in vitro and in vivo. 1. Cell Sci. 60, 181-197. U.S. Environmental Protection Agency (U.S. EPA) (1982). "Pesticide Assessment Guidelines. Subdivision M, Biorational Pesticides." Office of Pesticide and Toxic Substances, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA) (1998). "Reregistration Eligibility Decision (RED). Bacillus thuringiensis." Prevention, Pesticides,
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CHAPTER
41 The Influence of Age on Pesticide Toxicity Carey Pope Oklahoma State University
41.1 GENERAL CONCEPTS IN DIFFERENTIAL SENSITIVITY TO PESTICIDES Age-related differences in sensitivity to pesticides can occur for a wide variety of reasons. Toxicokinetic differences among different age groups can contribute to differential sensitivity, with differences in biotransformation often being a major factor. In other instances, toxicodynamic differences may exist which lead to age-related differences in sensitivity. For example, during development and maturation, a critical time of exposure or "window of opportunity" during which a developmental process occurs may impart selective sensitivity. At the other end of the spectrum, changes associated with aging may alter sensitivity to pesticide toxicity. Moreover, the relative contributions of toxicokinetic and toxicodynamic factors in age-related sensitivity may differ markedly among the various classes of pesticides, and even among members of the same class of toxicants. Exposures to pesticides are often age-related, based on age-specific behaviors, diets, or other factors. Thus, the nature of age-related differences in sensitivity to pesticides is complex, and broadbased generalities are typically unjustified. See Table 41.1. With the common routes of exposure (i.e., oral, dermal, and inhalation), a pesticide must first be absorbed before systemic toxicity can occur. In a comparative study of 14 different pesticides, 11 of these exhibited age-related differences in percutaneous absorption (Shah et al., 1987). Interestingly, four of the 14 showed greater absorption in young (33 days old), while seven of the 14 showed more extensive absorption in adults (82 days old). Moreover, even within the same class of pesticide (e.g., the organophosphorus toxicants parathion and chlorpyrifos), no clear age-related pattern of dermal absorption was evident, that is, chlorpyrifos showed greater absorption in young while parathion showed greater absorption in older animals. Hall and co-workers (1992) reported that dermal absorption of the dinitrophenol pesticide dinoseb was lower (about 20%) in 33-day-old rats compared to adults (82 days of age). Thus, Handbook of Pesticide Toxicology Volume 1. Principles
differences in rates or extent of absorption can contribute to differential sensitivity among age groups. Once absorbed, differences in tissue distribution or rates of elimination between age groups can contribute to differential sensitivity. Older animals (and people) typically have a higher fat content than younger individuals, which can have an important effect on distribution, accumulation, and storage of highly lipophilic pesticides, for example, organochlorines. Deichmann (1972) reported that DDT was eliminated from the body most efficiently in neonates and less so in older rats, at least partially because of differences in partitioning of the pesticide into fatty tissues. Changes in biotransformation during maturation and aging can often contribute to age-related differences in sensitivity. Immature and very old animals generally have lower biotransformation capacities, for example, lower levels of CYP450 (Benke and Murphy, 1975; Mehendale, 1980; Wynne et al., 1987). If a pesticide is activated by CYP450 to a more toxic metabolite, lower levels of CYP450 could be associated with lower sensitivity to that pesticide. In contrast, pesticides which are inactivated by monooxygenases may be relatively more toxic in groups with lower CYP450 levels. Lower activities of phase 11 reactions in neonatal or aged animals may also increase sensitivity to certain pesticides (Borghoff et al., 1988; Das et al., 1981; Egaas et al., 1995). Because of the complexity of pathways and the multiplicity of reactions generally involved in xenobiotic metabolism, however, differences in individual metabolic processes between age groups have to be considered in context to appreciate the net consequences of biotransformation on age-related toxicity. For example, while young rats exhibit lower rates of CYP450-mediated activation of the organophosphorus pesticide parathion to its active metabolite, paraoxon, lesser capacity in neonates for detoxification of paraoxon appears to be a prominent difference contributing to higher sensitivity (Benke and Murphy, 1975). Toxicodynamic differences can also contribute to age-related sensitivity. The ability to restore function following toxic ant exposure may be higher in some age groups than in others. For
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CHAPTER 41
Age and Pesticide Toxicity
Table 41.1 General Factors Contributing to Age-Related Differences in Sensitivity to Pesticides Toxicokinetic
Toxicodynamic
Exposure-based
Differences in absorption
Age-related expression of target molecules
Age-related behaviors
Differences in distribution/elimination Differences in biotransformation
or sensitive processes
Age-related diets
Differential capacities to recover from or adapt to toxicant insult
example, young rats challenged with hepatotoxic ants recover much better than older animals, apparently because of more rapid and robust synthesis of new cells following the initial tissue damage (Dalu and Mehendale, 1996). More rapid recovery of acetylcholinesterase (AChE) activity in younger animals (Chakraborti et aI., 1993; Moser, 1999; Pope et al., 1991; Pope and Liu, 1997) and slower recovery in aged animals (Michalek et aI., 1990) following acute exposure to an organophosphorus anticholinesterase may make those age groups differentially sensitive to subsequent anticholinesterase exposures. Thus age-related differences in sensitivity to pesticides can be influenced by multiple toxicokinetic and toxicodynamic factors. An important consideration in the differential sensitivity to pesticides is the time available for toxicity to develop. Children have a longer time to live than adults; thus, if pesticide exposure leads to the development of some form of delayed toxicity, for example, tumor formation, a child has more time for this adverse effect to be exhibited. Conversely, older individuals have experienced a longer time to accumulate residues or damage from long-term exposures. As noted before, the critical time-dependent nature of developmental stages is also an important consideration in age-related differences in response to pesticides. The endogenous metabolite bilirubin, for example, induces encephalopathy in the developing nervous system only at certain early timepoints when the blood-brain barrier is deficient (Lee et aI., 1995; Wennberg et aI., 1993). Another factor of particular importance to age-related differences in susceptibility is differential exposures. Age-related behaviors may contribute to differential exposure and sensitivity. For example, young children tend to sample the environment by taste. If the opportunity arises for oral "sampling" of a pesticide container, the young child may be much more susceptible to toxicity based on a greater likelihood of such exposure. In general, young children tend to be more exploratory and inquisitive than adults, which can sometimes lead to contact with inappropriately stored chemicals. Many lipophilic xenobiotics concentrate in breast milk; thus, breastfed infants may be preferentially exposed to such toxicants (Mussalo-Rauhamaa et aI., 1984; Schildkraut et aI., 1999). Young children eat more in proportion to their body size and they tend to eat more frequently than adults. When pesticide residues are consumed with the food, the relative frequency of exposure can be important if recovery takes longer than the time between exposures. Toddlers are also in contact with the floor more than are adults. With a higher surface-area-to-body-weight ratio, dermal contact may
Differences in time available for cumulative exposures and/or expression of toxicity
be more extensive than in adults. When pesticide residues fall to the floor after household applications, or become associated with carpeting or furniture, there is a higher probability of direct dermal contact in children playing on those surfaces (Fenske et aI., 1990; Lu and Fenske, 1999). Conversely, adults can be exposed to chemicals in the workplace, an exposure possibility which is generally missing in young children and older adults. Obviously, there are many reasons why exposures to pesticides can be age-related. The role of differential exposure in age-related sensitivity to pesticides is a critical issue and is discussed in more detail in Chapter 16. It is apparent, however, that age-related differences in sensitivity to pesticides can be caused by differences in inherent sensitivity to the pesticide, differences in exposure, or both. Clearly, multiple factors can contribute to differential susceptibility to pesticides throughout life. Risk assessment for pesticides relies heavily on data generated from animal studies. In fact, the United States Environmental Protection Agency recently limited the use of human data in the pesticide registration process (EPA, 1998a). The use of rodent animal models to estimate age-related differences in sensitivity in humans has some inherent problems, however, in particular when modeling the effects of early postnatal exposures. Developmentally, the maturational states of experimental animals and humans at parturition and perinatal periods can be quite different (Romijn et al., 1991). If neonatal rodents are more sensitive than adults to a particular pesticide, but only briefly during the early postnatal period, they may not represent a valid model for children because of the species differences in maturation relative to the timing of exposure. The comparative development, maturation, and aging of organ systems between man and experimental animals must be kept in mind when extrapolating age-related differences in sensitivity from animal models.
41.2 CHILDREN'S HEALTH AND REGULATION OF PESTICIDES IN THE UNITED STATES Ideally, regulatory policies governing the use of pesticides should be conservative enough to allow for protection of all members of the population. With noncarcinogenic toxicants, an uncertainty factor of ten is generally incorporated into the risk assessment process for such purposes (Bames and Dourson,
41.3 Age-Related Differences in Sensitivity to Pesticides
1988), assuming that variability in sensitivity to a particular agent within subpopulations is no greater than an order of magnitude. For this to be true and for all members of the population to be protected, all possible extrinsic and intrinsic modifiers of toxicity, for example, nutrition, disease, physiological stressors, or genetic polymorphisms, must together contribute to less than a ten-fold variation in sensitivity in the entire population. One intrinsic modifier of toxicity, age, has received considerable attention in recent years. In particular, the relative sensitivity of developing infants and children to pesticides has been the focus of concern (Bearer, 1995; Fenner-Crisp, 1995; Garrettson, 1997; Goldman, 1995; Little, 1995; Tilson, 1998). In 1988, the National Academy of Sciences (NAS) initiated a concerted effort to evaluate pesticide exposures in infants and children and to determine if the health of children was adequately addressed in the regulation of pesticides. The Committee on Pesticides in the Diets of Infants and Children, composed of scientists from industry, government, and academia, was established within the National Research Council of NAS in 1988 to evaluate the relative sensitivity of infants and children to pesticides. The conclusions eventually reached by this select committee had far-reaching consequences (see later). In 1989, public attention in the United States was focused by media coverage of a report from the Natural Resources Defense Council (NRDC) entitled Intolerable Risk: Pesticides in our Children's Food (NRDC, 1989) on the possibility that children were being exposed to excessive levels of pesticide residues in food products. The executive summary of this report begins "Our nation's children are being harmed by the very fruits and vegetables we tell them will make them grow up healthy and strong." While the basis of many claims in the NRDC report may have been inaccurate (Wilkinson and Ginevan, 1989), the public attention raised by this report had a significant impact; that is, it strengthened the commitment to ensure that children's health was adequately considered in the risk assessment of pesticides. Four years later, the National Academy of Sciences published the report Pesticides in the Diets of Infants and Children (NAS, 1993), which detailed conclusions from the NRC committee with the same name. Major findings of this committee included (i) that both quantitative and qualitative differences in toxicity of pesticides can occur between children and adults, but quantitative differences are usually less than a factor of ten, (ii) that infants and adults differ quantitatively and qualitatively in the types of pesticide exposures in the diet, a factor generally more important than differences in inherent sensitivity, (iii) that assessment of pesticide exposures should consider dietary as well as nondietary sources, and (iv) that "in the absence of data to the contrary, there should be a presumption of greater toxicity to infants and children" (NAS, 1993). The findings from this committee provided impetus for federal legislation addressing pesticide regulation, in particular regarding potential problems with differential exposure and sensitivity in children. In 1996, the Food Quality Protection Act (FQPA), containing sections relating to the protection of infants and children from pesticide exposures, was passed into law. The FQPA
875
amended the Federal Insecticide, Fungicide and Rodenticide Act and the Federal Food, Drug and Cosmetic Act (FFDCA). Section 408(b)(2)(C) of FFDCA states that with "threshold" adverse effects, "an additional tenfold margin of safety for the pesticide chemical residue ... shall be applied for infants and children to take into account potential pre- and post-natal toxicity and completeness of the data with respect to exposure and toxicity to infants and children." Further, this section of FFDCA states that the "Administrator may use a different margin of safety for the pesticide chemical residue only if, on the basis of reliable data, such margin will be safe for infants and children." In October of 1995, the United States Environmental Protection Agency (EPA) announced that it would explicitly evaluate risks to infants and children in all regulatory actions, and in April of 1997, Executive Order 13045 directed Federal agencies to identify and assess environmental health and safety risks to children (EPA, 1998b). Thus, the default position of EPA in pesticide regulatory decisions was to use an additional 10 x uncertainty factor (the FQPA factor) for threshold effects to ensure the protection of infants and children from pesticide toxicity. The EPA Office of Pesticide Programs proposal included, however, the possibility of either removing or reducing the magnitude of the FQPA factor if "reliable data" were available that suggested infants and children would be adequately protected under those conditions (EPA, 1999). Thus, risk assessment procedures for pesticides registered with the U.S. EPA now incorporate an additional uncertainty factor for infants and children unless sufficient data indicates that young are not at higher risk. The conclusions from the NAS report (NAS, 1993) regarding risks to infants and children were based on two parameters, that is, differences in sensitivity and differences in exposure. The following is a brief summary of evidence pertaining to age-related differences in response to pesticides. It should be noted that while the recent focus of concern in the United States has been on the possibly higher susceptibility of infants and children, because of the demographics of societal aging, elderly individuals and their relative susceptibility to pesticides may become a more important issue in the future (Overstreet, 2000). Alterations in cholinergic neurotransmission with aging and associated neurological disorders such as Alzheimer's disease may be particularly important in contributing to differential sensitivity to the cholinesteraseinhibiting agents and with other pesticides which may alter cholinergic neurotransmission.
41.3 AGE-RELATED DIFFERENCES IN SENSITIVITY TO PESTICIDES It is apparent that, as with other types of xenobiotics (Done, 1964; Goldenthal, 1971), there is no consistent effect of age on acute sensitivity to pesticides across all classes or even within a class of compounds. There are various factors that could contribute to differential toxicity, whether one compares different age groups, different species, different sexes, different strains,
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Table 41.2 Studies Reporting Age-Related Differences in Sensitivity with the Major Classes of Pesticides Type of study: Pesticide class Organophosphorus
relative sensitivity
References
Animal study:
Brodeur and DuBois (1963); Gagne and Brodeur (1972);
Immature more
Benke and Murphy (1975); Mendoza (1976); Long et al. (1986);
sensitive than adults
Pope et al. (1991); Atterberry et al. (1997); Moser and Padilla (1998); Karantb and Pope (2000)
Animal study: Adults
Lu et al. (1965); Chakraborti et al. (1993); Pope and Liu (1997);
more sensitive than
lohnson and Barnes (1970); Moretto et al. (1991);
immature
Peraica et al. (1993); Pope et al. (1992, 1993); Harp et al. (1997)
Animal study: Aged
Veronesi et al. (1990); Karantb and Pope (2000)
adults more sensitive than young adults Human study:
Diggory et al. (1977)
Children more sensitive than adults Organochlorines
Animal study:
Eriksson (1997); Samanta and Chainy (1997);
Immature more
linna et al. (1989)
sensitive than adults Animal study: Adults
Lu et al. (1965); Kiran and Varma (1988)
more sensitive than immature Carbamates
Animal study:
Moser (1999) (based on lethality)
Immature more sensitive than adults Animal study: Aged
Knisely and Hamm (1989); Takahashi et al. (1991)
adults more sensitive than young adults Human study:
Lifshitz et al. (1997) (depending on endpoint)
Children more sensitive tban adults Pyrethroids
Animal study:
Cantalamessa (1993); Sheets et al. (1994)
Immature more sensitive than adults
or any other factor. These contributing factors will be examined in more detail with specific examples of pesticides potentially capable of eliciting age-related effects. See Table 41.2. 41.3.1 ORGANOPHOSPHORUS PESTICIDES
Organophosphorus pesticides (OPs) elicit toxicity through inhibition of AChE (Mileson et aI., 1998). Age-related differences in sensitivity to OPs have been reported in many experimental studies (Benke and Murphy, 1975; Brodeur and DuBois, 1963; Gagne and Brodeur, 1972; Gaines and Linder, 1986; Mendoza, 1976; Moser and Padilla, 1998; Pope et aI., 1991). In general (but not always), neonatal animals are more sensitive to the acute toxicity ofOPs. Lu and co-workers (1965) reported a maturational decrease in sensitivity to malathion among newborn,
14- to 16-day-old, and adult rats. Mendoza (1976) reported that 1-day-old rats were about nine times more sensitive to lethality from acute exposure to malathion. Mortality in newborn pigs following dermal application of chlorpyrifos (2.5% aerosol) was markedly higher when exposure occurred within the first three hours of life than at 30-36 hours after birth, suggesting a rapid change in sensitivity in the first days following parturition (Long et aI., 1986). Pope and co-workers (1991) reported that 7-day-old rats were 2-9 times more sensitive than adult (90 days of age) rats to the acute toxicity of methyl parathion, parathion, and chlorpyrifos. In contrast, methamidophos appears to elicit little age-related toxicity (Moser, 1999; Padilla et aI., 2000). Several factors could contribute to age-related differences in response to acute OP exposures. Gagne and Brodeur (1972)
41.3 Age-Related Differences in Sensitivity to Pesticides
investigated potential metabolic factors in the higher sensitivity of weanling rats to parathion and concluded that limited detoxification of parathion and its metabolite paraoxon were at least partially responsible. Later, Benke and Murphy (1975) evaluated metabolic contributions to age-related differences in sensitivity to parathion and methyl parathion. When metabolic rates were compared to LDso values in different age groups, high correlations were noted between lethality and liver and plasma A-esterase activity, oxon dealkylation and dearylation, and binding to "noncritical tissue constituents" in liver and plasma. They concluded that more robust metabolic inactivation of the active oxons of these two pesticides in more mature animals was primarily responsible for decrease in sensitivity with age. Atterberry and co-workers (1997) compared the toxicity and biotransformation of parathion and chlorpyrifos in neonatal and adult rats and concluded that differences in liver carboxylesterase activity and CYP450-dependent dearylation were important in differential age-related sensitivity to these pesticides. Moser and colleagues (1998) concluded that differences in liver carboxylesterase and A-esterase activities formed the basis for age-related differences in sensitivity to acute chlorpyrifos exposures. Other studies have indicated that maturational differences in the capacity for detoxification of organophosphates by A-esterases and carboxylesterases may contribute to higher sensitivity to these pesticides in immature animals (Costa et aI., 1990; Li et aI., 1993, 1995; Maxwell, 1992; Pond et aI., 1995). Thus, considerable evidence suggests that immature animals are more sensitive to the acute toxicity of several OP pesticides because of limited detoxification of either the parent compound or its active metabolite. Young children also appear to be more sensitive to acute toxicity from OP exposure. In a case of parathion-contaminated food in Jamaica, the highest incidence of lethality was in children less than five years of age (Diggory et aI., 1977). Differences in metabolic capacities between very young children and older children or adults may also be primary determinants in age-related sensitivity to acute OP exposures. Augustinsson and Barr (1963) showed that serum arylesterase (A-esterase) activity was very low in newborn children but increased steadily during the first six months of life. Ecobichon and Stephens (1973) reported that plasma cholinesterase and A-esterase activities increased dramatically in children during the first year of life, after which no further increases occurred. Any active anticholinesterases in the blood of very young children would therefore be less likely to bind to nontarget cholinesterases or to be hydrolyzed by A-esterases; thus, more inhibitor would be available to reach target tissues. As detoxification of active OP anticholinesterases is thought to be a prominent factor in age-related sensitivity (Atterberry et aI., 1997; Benke and Murphy, 1975; Mortensen et aI., 1996; Moser and Padilla, 1998), these studies suggest that dramatically higher acute sensitivity in children may only exist in the very young (:sI year of age), however, when these detoxification processes appear most limited. In addition to metabolic differences that may contribute to age-related sensitivity to OP pesticides, some toxicody-
877
namic differences among age groups could also be important. Organophosphorus and carbamate pesticides are toxic by virtue of their ability to inhibit AChE (Fukuto, 1990). Species differences in sensitivity of AChE to inhibition by some OP anticholinesterases have been reported (Kemp and Wallace, 1990). Thus, there could be a degree of selective toxicity among age groups based on the molecular interaction between the toxic ant and its "receptor," AChE. Several studies have reported, however, that AChE sensitivity to the inhibitors is not a contributing factor to age-related differences in sensitivity (Atterberry et aI., 1997; Benke and Murphy, 1975; Mortensen et aI., 1998). Upon extensive inhibition of AChE in the nervous system, the neurotransmitter acetylcholine accumulates in synapses, causing excessive stimulation of cholinergic receptors on postsynaptic cells, leading to cholinergic toxicity. It is known that feedback inhibition of acetylcholine release can occur through activation of muscarinic acetylcholine receptors located on presynaptic terminals (Allgaier et aI., 1993; Vickroy and Cadman, 1989; Weiler et aI., 1989). Activation of these presynaptic muscarinic receptors diminishes further acetylcholine release and thereby may reduce the excessive stimulation of postsynaptic cholinergic receptors following extensive AChE inhibition. Pedata and co-workers (1983) reported that muscarinic autoreceptor function was absent in 7-day-old rat brain but viable in brain from 21-day-old animals. Thus, with extensive AChE inhibition, very young rats do not have an adaptive mechanism which limits further neurotransmitter release in times of excess (e.g., when AChE is inhibited). Pedata and co-workers (1983) and Meyer and Crews (1984) reported that evoked acetylcholine release was lower in tissues from both neonatal and aged brain compared to animals 1-6 months of age. Differences in the amount of acetylcholine released upon stimulation between the age groups may therefore contribute to differences in response to AChE inhibitors. The function of muscarinic autoreceptors appears markedly reduced with aging in some rat brain regions (Araujo et aI., 1990). A deficit or lack of feedback inhibition of acetylcholine release in some age groups may limit their adaptation to synaptic AChE inhibition and contribute to higher sensitivity (Pope, 1999). Differences in acute sensitivity to OP anticholinesterases between neonatal and adult rats may therefore have both a toxicokinetic and a toxicodynamic basis. It should be stressed, however, that the studies cited above generally used lethality as the endpoint for estimating age-related sensitivity. By definition, dosages at or near those causing death would have to be considered "high"-level exposures. Less information is available regarding age-related differences in sensitivity to lower levels of exposure. While of prominent importance with acute, high-level exposures where detoxification systems may be saturated, differential detoxification capacities in neonatal and adult animals may have lesser importance when repeated, lower-level exposures occur. With lower, nonlethal dosages, less AChE activity would be inhibited, with lesser signs of cholinergic toxicity. At even lower dosages, some degree of AChE inhibition could occur in the absence of any overt toxicity (Nostrandt et aI., 1997).
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Under these conditions, feedback inhibition of acetylcholine release (or lack of that adaptive mechanism in neonatal animals) would have little consequence. Thus, with acute dosages of pesticide high enough to cause some level of AChE inhibition but with no alteration of cholinergic neurotransmission, two factors which apparently make younger animals more sensitive (lower detoxification capabilities, lesser adaptive regulation of neurotransmitter release) may have no relevance. With such repeated lower-level exposures, however, another toxicodynamic factor (i.e., recovery of AChE activity following inhibition) may play a more prominent role. As mentioned before, AChE activity following OP exposure recovers much faster in neonatal tissues (Pope et ai., 1991) and much slower in aged animals (Micha1ek et al., 1990) than in adult tissues. While neonatal rats were more sensitive to single, high dosages of chlorpyrifos, adults exhibited more extensive changes in cholinergic neurochemical markers (i.e., AChE inhibition, muscarinic receptor binding) following repeated, intermittent dosing (40 mg/kg, every four days for a total of four exposures) (Chakraborti et al., 1993; Pope and Liu, 1997). Apparently, while young rats are more sensitive to the acute effects of chlorpyrifos, they can recover much faster than adults from the biochemical insult. When exposures are sufficiently separated in time, neonatal animals can regain AChE activity faster and avoid cumulative inhibition with repeated exposures. In contrast, in particular with OPs such as chlorpyrifos which produce long-term inhibition of AChE, activity recovers more slowly in adult tissues, allowing accumulative inhibition with subsequent exposures. Thus, one could argue that with acute chlorpyrifos dosing, young animals are more sensitive than adults, but with repeated dosing, age-related sensitivity is reversed. Clearly, the nature of the exposures (acute vs. repeated, high level vs. low level) can influence age-related differences in sensitivity to these toxicants. Few studies have evaluated the relative sensitivity of aged animals to organophosphates. Acetylcholinesterase activity in some brain regions (e.g., hippocampus, cortex) but not others (e.g., pons-medulla) of rats declines with aging (Bisso et ai., 1991; Meneguz et aI., 1992). As mentioned before, recovery of AChE activity as well as muscarinic receptor binding following repeated organophosphate exposures was impaired in aging brain, in particular in the cerebral cortex (Michalek et ai., 1990). Age-related differences in baseline activity of cholinergic neurochemical processes or their adaptive responses to pesticide exposure could therefore influence sensitivity to some anticholinesterases. Veronesi and co-workers (1990) evaluated the effects of chronic fenthion exposure (25 mg/kg, three times a week for 10 months) in either young (2 months old) or aged (12 months old) rats. Using this dosing treatment schedule, chronic (10 months) fenthion exposures initiated in young rats produced gliosis and necrosis in the dentate gyrus and CA!, CA3, and sometimes CA2 regions of the hippocampus. Aged rats treated with the same regimen of fenthion exhibited similar degrees of hippocampal degeneration earlier during the progression of exposure, that is, by 2 months, and much more extensive pathol-
ogy than noted in the younger animals when evaluated following 10 months of exposure. These studies suggest that persistent acetycholinesterase inhibition by fenthion can produce neuropathological changes in the rat hippocampus and that aged rats are more sensitive than younger rats to such effects. Karanth and Pope (2000) compared acute sensitivity to chlorpyrifos and parathion in neonatal (7 days old), juvenile (21 days old), adult (90 days old), and aged (24 months old) Sprague Dawley rats. Neonatal and juvenile rats were more sensitive than adults to both toxicants. Adult and aged rats were similar in sensitivity to chlorpyrifos, but aged animals were markedly more sensitive than adults to parathion. Moreover, plasma carboxylesterase activity among groups was highly correlated with acute sensitivity to parathion, further suggesting a toxicokinetic basis for the age-related differences in sensitivity to this pesticide. The above discussion pertains to differences in sensitivity among different age groups to the cholinergic toxicity of OP pesticides. Some recent reports suggest that OP pesticides may affect macromolecular synthesis and cell viability in the brain following early postnatal exposures, independent of AChE inhibition (Slotkin, 1999). Whitney and co-workers (1995) reported that DNA and protein synthesis could be affected by chlorpyrifos in a time-dependent and brain regional-dependent manner. When postnatal rats (11-14 days of age) were given chlorpyrifos (1 mg/kg/day), a delayed reduction in DNA concentration and content in forebrain was noted at 15-20 days of age (Campbell et aI., 1997). Reductions in cellular RNA concentration and content were also reported in the brainstem and fore brain following repeated postnatal chlorpyrifos exposures in rats (Johnson et aI., 1998). Song and co-workers (1997) reported that repeated postnatal exposures to chlorpyrifos in rats affected multiple components of the adenylyl cyclase cascade system (e.g., activity of adenylyl cyclase, G-protein function, expression of neurotransmitter receptors coupled to adenylyl cyclase). Moreover, changes in these processes were noted in the cerebellum, a brain region with only sparse cholinergic innervation. Other studies suggest that anticholinesterases may affect neuronal adhesion and neurite extension (Bigbee et ai., 1999; Dupree and Bigbee, 1994; Small et aI., 1995; Song et aI., 1998). Thus, OP pesticides may be capable of altering macromolecule synthesis, intracellular signaling, and neuronal adhesion/outgrowth in the developing brain, apparently independent of catalytic inhibition of AChE. Some organophosphorus toxicants can induce a delayed neuropathological response referred to as organophosphorusinduced delayed neurotoxicity (OPIDN) (Abou-Donia, 1981). This form of neurotoxicity is not associated with AChE inhibition but has been correlated with the inhibition of another enzyme in the nervous system called neurotoxic esterase (NTE) (Johnson, 1976, 1980). Individuals affected by this delayed neurotoxicity exhibit gait disturbances (incoordination and difficulties in walking) and sensory deficits (numbness and tingling, particularly in the fingers and toes), which mayor may not follow signs of toxicity characteristic of AChE inhibition. Degeneration of certain nerve tracts in both the central and the
41.3 Age-Related Differences in Sensitivity to Pesticides
peripheral nervous systems has been demonstrated in OPIDN. It has more recently been observed that some compounds [e.g., the common protease and NTE inhibitor phenylmethylsulfonyl fluoride (PMSF)], while not being capable of inducing delayed neurotoxicity, can potentiate or promote delayed neurotoxicity caused by an OP (Lotti et aI., 1991; Pope and Padilla, 1990; Pope et aI., 1993). The sequence of administration of the two compounds is of paramount importance; that is, for delayed neurotoxicity to be exacerbated, OP exposure must precede exposure to the potentiating agent. Young animals are resistant to delayed neurotoxicity (Johnson and Barnes, 1970; Moretto et aI., 1991). Before the age of about 6-7 weeks, chickens (the animal model of choice for studies of delayed neurotoxicity) are completely resistant to functional and morphological signs of OPIDN. From about 7-10 weeks of age, sensitivity develops, and at about 12-14 weeks of age, they become completely sensitive (Moretto et aI., 1991; Pope et aI., 1992). Studies have also examined the potentiation of OPIDN in young animals (Peraica et aI., 1993; Pope et aI., 1992). As stated above, five-week-old chickens are normally resistant to the clinical and morphological changes associated with delayed neurotoxicity. If OP exposure is followed by treatment with PMSF, however, overt delayed neurotoxicity can be demonstrated. On the other hand, clinical and morphological changes typical of OPIDN are generally not elicited in very young chickens (e.g., two weeks of age) regardless of the dose of the OP or whether a potentiating agent is given after the OP (Harp et aI., 1997). Just as the mechanism(s) underlying OPIDN itself has not been elucidated, the basis for age-related differences in sensitivity to delayed neurotoxicity remains unknown. In contrast to age-related sensitivity to acute toxicity from most OPs, however, young are less sensitive to the delayed neurotoxicity of OPs. 41.3.2 ORGANOCHLORINE INSECTICIDES At one time, organochlorines (OCs) constituted the highest-use pesticide class in the world. With increased awareness of ecological damage, global contamination, and insect resistance, the use ofOCs has decreased. The most well-known OC, DDT, has been extensively studied. In acute toxicity studies, DDT is actually less toxic to neonatal rats than to adults (Lu et aI., 1965). In this same study, dieldrin, another OC, was also reported to be less toxic in neonatal rats. Later studies have suggested that early neonatal exposure to DDT (0.5 mg/kg, po) can have long-lasting consequences (Eriksson et aI., 1984, 1993). Total cholinergic muscarinic receptor (eH]QNB) density was increased in cortex one week after DDT exposure in lO-day-old rats, but no effect was noted in hippocampus. Moreover, muscarinic receptor binding was still altered at 4 months of age following this single treatment with DDT, but at this time there was a reduction in binding density. Functional alterations (deficits in locomotor habituation) were also noted in rats four months after acute DDT exposure (Eriksson, 1997). Of particular interest in these studies
879
was the observation that neonatal (10 days old) rats treated with DDT (0.5 mg/kg) showed an increase in cortical muscarinic receptor binding one week after exposure, whereas adult rats treated similarly showed a decrease in receptor binding. Moreover, neither 3-day-old rats nor 19-day-old rats showed the same response (i.e., up-regUlation of muscarinic receptors) when treated similarly with DDT (Eriksson, 1997). Subsequent studies have shown that 1O-day-old mice treated with DDT and then challenged at five months of age with bioallethrin showed increased expression of the m4 subtype of muscarinic receptors in selected brain regions (cortex and striatum) (Talts et aI., 1998b). Thus, there appears to be a critical developmental window in which alteration of the cholinergic system can occur following early DDT exposure, and changes in muscarinic receptor density induced by DDT appear specific for the m4 subtype. While most OCs have been banned from use in the United States, their use continues in other countries. A few OCs are still used in the United States, for example, lindane (y-hexachlorocyclohexane). Lindane is still commonly prescribed for treatment of scabies and pediculosis. Rivera et al. (1990) reported that repeated, relatively low-level exposures to lindane (10 mg/kg/day for seven days) during postnatal week 1 or 2 induced transient changes in reflex behaviors (e.g., surface righting, cliff avoidance) and locomotor hyperactivity, in the absence of overt signs of toxicity. Serrano and co-workers (1990) reported that early postnatal lindane exposure reduced the level of myelin basic protein and 2',3' -cyclic nucleotide 3' -phosphodiesterase activity, an enzyme in high concentrations in myelin and myelin-forming cells, in a dose-dependent manner. Lindane exposure (either acute [20 mg/kg] or repeated [10 mg/kg/day for seven days]) in rats 15 days of age caused complex behavioral changes (improvement in passive avoidance behavior, alterations in locomotor activity) and apparent enhanced turnover of brain monoaminergic neurotransmitters (Rivera et aI., 1998). While these studies only evaluated toxicity in postnatally maturing animals, the endpoints evaluated and the changes noted suggested that higher sensitivity may exist in younger individuals. Samanta and Chainy (1997) reported that acute lindane exposure (50 mg/kg, i.p.) caused only minimal lipid peroxidation in liver of 30-day-old chickens but more extensive oxidative changes in 7 -day-old animals. Furthermore, superoxide dismutase was inhibited and glutathione levels were elevated by lindane in 7-day-old but not 30-day-old chickens. Thus, lindane can cause diverse age-related changes that generally target younger animals. Endosulfan is another OC that is still commonly used in the United States. Kiran and Varma (1988) studied the toxicity of endosulfan in different age groups of rats (12.5 mg/kg/day for four days beginning at 15,30,70, and 365 days of age). Hyperglycemia and glycogen depletion were most extensive in 365day-old animals and least affected in the youngest age group. Liver aldolase activity was also reduced more in older rats than in younger animals. In contrast, red blood cell Na+IK+ ATPase activity was inhibited more in the youngest age group. These
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CHAPTER 41
Age and Pesticide Toxicity
results indicate complex age-related differences in response to endosulfan. Chlordecone is an organochlorine that causes hyperexcitability, tremors, incoordination, and other signs of neurotoxicity (Tilson and Mactutus, 1982). Several studies have evaluated the effects of early postnatal exposure to chlordecone. Tilson and co-workers (1982) reported that rats exposed acutely on postnatal day 4 to chlordecone had markedly altered responses during reversal of visually cued nose poke behavior when tested at about four months of age. Neonatal chlordecone exposure was also reported to alter passive avoidance performance (Mactutus et aI., 1982). Jinna and co-workers (1989) reported that chlordecone inhibited rat brain ATPases (Na+ /K+ ATPase, Ca++ ATPase) in an age-related manner, that is, neonatal enzyme activity was more sensitive to inhibition by chlordecone in vitro. Chlordecone has been shown to potentiate the hepatotoxicity of halogenated solvents, for example, carbon tetrachloride (Soni and Mehendale, 1998). Twenty and 45-day-old rats were resistant to chlordecone-enhanced hepatotoxicity relative to 60-day-old animals, however (Dalu and Mehendale, 1996). Dosages of chlordecone (10 ppm in the diet for 15 days) and carbon tetrachloride (0.1 ml/kg, i.p.) that caused 100% lethality in the adult rats caused 0% and 25% lethality in 20- and 45-dayold animals. It was concluded from these studies that the relative ability of the liver to recover from injury was the prominent factor underlying age-related differences in toxic outcome, with immature animals being more competent than adults at restoring tissue integrity and function. Thus, while these studies do not indicate age-related differences in sensitivity to chlordecone alone, they suggest that the modulation of solvent hepatotoxicity by chlordecone can occur in an age-related manner. Many of the OCs, for example, DDT, chlordecone, methoxychlor, chlordane, and endosulfan, have also been noted to interact directly with hormonal receptors (Tilson, 1998). The DDT analog, methoxychlor, is still used in the United States and has been shown to both alter sex-related hormones and reproductive function in rats treated postnatally (Chapin et aI., 1997). The endocrine-disrupting capacity of these agents could be cause for concern with early exposures (Cassidy et aI., 1994; Chapin et aI., 1997; Davis et aI., 1993). The reader is referred to Chapter 16 for more information on endocrine disruption by pesticides. 41.3.3 CARBAMATES
Knisely and Hamm (1989) investigated the comparative actions of physostigmine on nociception in different age groups of rats (3, 17, and 25 months of age). Tail-flick latencies were dose-dependently altered in all age groups by physostigmine, but more extensive increases in latency were noted in the 17- and 25-month-old animals with higher dosages, suggesting higher sensitivity in the aged animals to this carbamate anticholinesterase. Such changes could be an indication of upregulation of cholinergic receptors due to loss of cholinergic innervation with aging.
Takahashi and co-workers (1991) compared the motor, sensory, and thermoregulatory responses of young adults (3 months of age) and older adults (12 months of age) to carbaryl (10 or 50 mg/kg, p.o.). Carbaryl affected nociception primarily in the older animals. Hypothermia was also affected in an agerelated manner. Locomotor changes following carbamate exposure, however, were similar between the two age groups. Again, these data illustrate the potential for age-related differences in response to a pesticide when one endpoint is used and conversely, lack of age-related differences in sensitivity when based on another endpoint. Moser (1999) reported that aldicarb was about twice as toxic in preweanling rats compared to adults using the acute maximum tolerated dose as the endpoint of sensitivity. Interestingly, preweanling rats exhibited fewer signs of functional toxicity than older animals, in the presence of similar levels of brain and blood cholinesterase inhibition. Furthermore, the young rats were resistant to locomotor alterations noted in older animals following aldicarb administration. Lifshitz and co-workers (1997) retrospectively compared the clinical course of poisoning in children (1-8 years of age) and adults (17--41 years of age) following carbamate pesticide exposures. In all cases, blood serum cholinesterase inhibition was approximately the same (10-30% below the lower limit of normal). Interestingly, signs of coma/stupor and hypotonia were noted in 100% of the children but in none of the adults. While miosis was noted in 92% of the adults, this sign was only recorded in 55% of the children. Moreover, muscle fasciculations were observed in 83% of the adults and in only 6% of the children. While the relative level of AChE inhibition in the target tissues was unknown, these results suggest that children may respond differently than adults following acute anticholinesterase exposures producing relatively similar degrees of blood cholinesterase inhibition. 41.3.4 PYRETHROID INSECTICIDES
Eriksson and Nordberg (1990) studied the effects of early postnatal exposures to one of two different pyrethroid insecticides, bioallethrin (a type I pyrethroid) and deltamethrin (a type 11 pyrethroid), on cholinergic receptors in mouse brain. With lower levels of exposure, bioallethrin (0.72 mg/kg/day from postnatal day 10-16) reduced high-affinity muscarinic receptor binding in brain, whereas deltamethrin (0.71 mg/kg/day) increased high-affinity binding, both in the absence of overt signs of toxicity. Deltamethrin also increased cortical [3H]nicotine binding. Higher levels of repeated exposure (72 and 1.2 mg/kg/ day for bioallethrin and deltamethrin, respectively) caused overt toxicity (tremor, choreoathetosis), but only deltamethrin affected cholinergic receptor binding under these conditions. Early exposure to bioallethrin in mice (0.7 mg/kg/day from postnatal day 10-16) was also reported to increase sensitivity to bioallethrin when administered at 7 months of age, suggesting long-term changes in sensitivity following exposure during postnatal maturation (Talts et aI., 1998a). These studies, similar
41.3 Age-Related Differences in Sensitivity to Pesticides
to those with early postnatal exposures to DDT (Eriksson et aI., 1984), indicate that development of some components of the cholinergic system may be sensitive to alteration by early postnatal exposure to "noncholinergic" pesticides (i.e., pesticides not having a primary action on some aspect of cholinergic neurotransmission). Cantalamessa (1993) compared the acute toxicity and metabolism of cypermethrin and permethrin in neonatal and adult rats. With both pesticides, an age-related decrease in acute toxicity was noted. Cypermethrin and permethrin were 16.8 and 4.4 times more toxic (based on 24-hour oral LDso values) in 8-day-old animals compared to adults. Carboxylesterase inhibition (by tri-ortho-cresyl phosphate) in neonatal animals failed to alter acute toxicity, but lethality was increased in adults by this pretreatment, suggesting that neonatal animals may be more sensitive to acute toxicity of these pyrethroids at least partially because of incomplete development of this detoxification system. Sheets and co-workers (1994) evaluated the sensitivity of preweanling, wean ling, and adult rats to a wide dose range of deltamethrin. Younger rats (11 and 21 days of age) were markedly more sensitive than adults (72 days of age) to the acute lethality of deltamethrin (LDso: 11 days = 5.1 mg/kg; 21 days = 11 mg/kg; 72 days = 81 mg/kg, p.o.). In contrast, using acoustic startle response to evaluate functional toxicity of lower-level exposures, the ED50 was the same between Il-dayold and 72-day-old animals. Based on these studies, age-related differences in sensitivity to deltamethrin could be considered substantial (if based on acute lethality) or nonexistent (if based on the acoustic startle response). Clearly, the selection of the endpoint used to define sensitivity, as well as the exposure conditions, can qualitatively influence determination of age-related susceptibility to these pesticides. 41.3.5 MISCELLANEOUS PESTICIDES
Gaines and Linder (1986) examined the comparative acute sensitivity of weanling (4-6 weeks of age) and adult rats to 34 pesticides from different chemical classes. The immature rats were more sensitive to only four of those pesticides. Moreover, differences in acute sensitivity to pesticides were generally only two- to threefold in magnitude. One problem with this study, however, was the age of the younger animals used, that is, 4- to 6-week-old rats. Similar studies using less mature animals (e.g., 1- to 3-week-old rats) may have yielded different conclusions. Watkinson studied the cardiotoxicity of the formamidine pesticide chlordimeform. Weanling (22-30 days of age [Watkinson, 1985]) and aged (24 months of age [Watkinson, 1986]) rats were treated sequentially with 5, lO, 30, 60, and 120 mg/kg chlordimeform (i.v.) or vehicle, and mean arterial blood pressure and heart rate were monitored. While chlordimeform reduced heart rate and blood pressure in both age groups, the magnitude of the changes was greater in the aged animals. Arrythmias were also less pronounced in younger animals and required higher thresholds of chlordimeform. In addition, while
881
a single injection of chlordimeform (60 mg/kg, i.v.) was lethal to all aged rats tested, only 23% of the weanling rats died following this level of exposure. Lower sensitivity of young rats to the lethality of chlordimeform had been previously reported (Robinson and Smith, 1977). Thus, it appears that younger animals are less sensitive than aged rats to the toxicity of the formamidine insecticide, chlordimeform. Ivermectin is a broad-spectrum antiparasitic agent (Campbell and Benz, 1984). Relatively low-level exposure to ivermectin (4 mg/kg/day) during gestation (GD 6-20) and lactation (postnatal days 2-20) caused lOO% lethality in pups with no apparent toxicity in dams (Poul, 1988). When exposure was limited to gestation, only 22% lethality was noted in the offspring. Lower exposure levels (1 mg/kg/day) had no effect on survival but delayed some developmental endpoints, including cliff avoidance and locomotion. Lankas and co-workers (1989) reported that newborn rodents were particularly sensitive to the neurotoxicity of ivermectin. Following application of ivermectin to control an ectoparasite infestation, Skopets and coworkers (1996) noted evidence of higher sensitivity of young mice to ivermectin. While all adults tolerated the ivermectin exposures, preweanling mice developed seizures or tremors, and lethality was observed in some cases. Together, these data suggest that neonatal rodents are more sensitive than adults to the acute toxicity of ivermectin. As ivermectin is typically prevented from access to the central nervous system in adults (Lovell, 1990), incomplete blood-brain barrier formation appears to contribute to these age-related differences in sensitivity (Lankas et aI., 1989). Age-related differences in sensitivity were noted following acute dibromochloropropane exposure (250 mg/kg, s.c.) in 4- and 9-week-old rats (Saegusa, 1987). It was noted that the older animals exhibited a higher incidence of lethality, more extensive body weight reductions, and more extensive tissue damage in kidney, intestine, and testes. Dithiobiuret (DTB, thioimidodicarbonic diamide) was originally proposed as a rodenticide and is a prototypical motor neuron toxicant that produces a flaccid weakness following repeated exposures (Atchison et aI., 1982). Using failure of the rotorod test as an indication of neuromuscular toxicity, Atchison and co-workers (1982) studied the sensitivity of weanling (25 days old), juvenile (50 days old), and adult (80 days old) rats to DTB (1 mg/kg/day). In females, the mean time to onset of rotorod failure was about six days in wean ling rats, four days in juveniles, and only about three days in adults. Neither differences in total accumulation of DTB nor distribution appeared to contribute to the differences in DTB toxicity among the age groups. These data provide another example of higher sensitivity to neurotoxic ants in adults compared to younger animals. Using a series of immunotoxicity assays, Smialowicz and co-workers (1989) reported that preweanling rats (3-24 days of age) were somewhat more sensitive than adults to tributyltininduced immune alterations. In addition, natural killer cell activity was only affected in the neonatal animals. Furthermore, some immune responses were altered in lO-week-old animals
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treated prior to weaning, suggesting long-term changes in immune function could occur with early exposure to tributyltin. Children may be more sensitive to the insect repellent, DEET (diethyl m-toluamide) (Couch and Johnson, 1992). DEET is used safely by an estimated 200 million people each year around the world (Brown and Hebert, 1997), but severe neurological manifestations have occasionally been associated with its use (Osimitz and Murphy, 1997). Four boys (age 3-7 years) had seizures within 48 hours of applying DEET to the skin. Six young girls (ages 1.5-8 years) exhibited seizures, ataxia, and/or coma after dermally applying DEET, and three of those children later died. These types of neurological signs have been reported in adults following oral consumption of large amounts of DEET (Tenenbeim, 1987). Thus, while occurrence is rare, children may exhibit serious signs and symptoms of neurotoxicity and can die following dermal application of this widely used repellent. Because of the scarcity of data on absorption, metabolism, or elimination of DEET in children, it is unclear why children may be more sensitive to this compound (Garrettson, 1997).
41.4 CONCLUSIONS Changes in sensitivity to pesticides can occur throughout the life span from early postpartum to senescence. Recently, there has been considerable concern that children may be at higher risk than adults to pesticides. Enactment of the Food Quality Protection Act in 1996 was in response to this concern, and called for consideration of additional safety in the risk assessment of pesticides to protect infants and children. It is clear from review of both experimental and clinical data, however, that there is no hard-and-fast rule regarding age-related differences in sensitivity to pesticides. While neonates may be more sensitive to the acute toxicity of some pesticides, adults may be more sensitive to others. Even within a class of toxic ants, for example, within the organophosphorus anticholinesterases, examples of higher sensitivity in both age groups can be demonstrated. In fact, even when a single pesticide is considered, age-related differences in sensitivity may change qualitatively depending on the conditions of exposure (e.g., acute vs. repeated dosing, high- vs. low-level exposures) or the endpoint measured. While maturational differences in biotransformation capacity may be limiting in some cases, for example, with acute, high-level exposures where detoxification enzymes could become saturated, such metabolic differences may be of lesser importance with repeated, lower levels of exposure to the same pesticides. Similarly, differences in the ability to recover following pesticide exposure may be much more important when repeated exposures occur than following acute exposures. Storage and clearance of pesticides may also be more important with repeated, long-term exposures. Age-related sensitivity to pesticides should therefore be evaluated on a case-by-case basis, recognizing both the factors which influence age-related differences in response and the critical importance of appropriate endpoint selection for establishing differential sensitivity.
Relative sensitivity can be expressed in one of two ways; that is, a subpopulation exhibits differences in sensitivity to a particular form of toxicity or a subpopulation exhibits qualitatively different forms of toxicity to the same pesticide. Young animals may be more sensitive to the acute lethality of some pesticides (e.g., chlorpyrifos), but this does not necessarily mean that young animals will be more sensitive to the same pesticides when sensitivity is based on nonlethal endpoints of toxicity. Risk assessments are typically performed using a "critical" endpoint, generally the most sensitive end point to the toxicant in question derived from a series of toxicity studies. Thus, even if a pesticide causes a qualitatively different form of toxicity in different age groups, the risk assessment and estimation of tolerable exposure levels will not change unless this response occurs at dosages lower than those defining the critical endpoint. There will always be uncertainty in risk assessment. One factor which contributes to that uncertainty is age and its influence on the response to a particular toxic ant. If the critical endpoint for a particular pesticide is well established based on "reliable" data derived from studies across all age groups, the contribution of age-related differences in sensitivity to such uncertainty can be minimized. Knowledge of mechanisms which contribute to such age-related differences in response to pesticides will ultimately aid in the safer use of these chemicals.
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Moser, V. C., and Padilla, S. (1998). Age- and gender-related differences in the time course of behavioral and biochemical effects produced by oral chlorpyrifos in rats. Toxicol. Appl. Pharmacol. 149, 107-119. Mussalo-Rauhamaa, H., Pyysalo, H., and Moilanen, R. (1984). Influence of diet and other factors on the levels of organochlorine compounds in human adipose tissue in Finland. 1. Toxicol. Environ. Health 13,689-704. National Academy of Sciences (NAS) (1993). "Pesticides in the Diets ofInfants and Children." National Academy Press, Washington, DC. Natural Resources Defense Council (NRDC) (1989). "Intolerable Risks: Pesticides in Our Children's Food." Natural Resources Defense Council, New York, NY. Nostrandt, A. c., Padilla, S., and Moser, V. C. (1997). The relationship of oral chlorpyrifos effects on behavior, cholinesterase inhibition, and muscarinic receptor density in rat. Pharmacol. Biochem. Behav. 58, 15-23. Osimitz, T. G., and Murphy, J. V. (1997). Neurological effects associated with use of the insect repellent N,N -diethyl-m-toluamide (DEET). 1. Toxicol. Clin. Toxicol. 35,435--441. Overstreet, D. H. (2000). Organophosphate pesticides, cholinergic function and cognitive performance in advanced age. NeuroToxicology 21, 75-81. Padilla, S., Buzzard, J., and Moser, V. C. (2000). Comparison of the role of esterases in the differential age-related sensitivity to chlorpyrifos and methamidophos. NeuroToxicology 21, 49-56. Pedata, E, Antonelli, T., Lambertini, L., Beani, L., and Pepeu, G. (1983). Effect of adenosine, adenosine triphosphate, adenosine deaminase, dipyridamole and aminophylline on acetylcholine release from electrically-stimulated brain slices. Neuropharmacology 22, 609-614. Peraica, M., Capodicasa, E., Moretto, A., and Lotti, M. (1993). Organophosphate polyneuropathy in chicks. Biochem. Pharmacol. 45, 131-135. Pond, A. L., Chambers, H. w., and Chambers, J. E. (1995). Organophosphate detoxication potential of various rat tissues via A-esterase and aliesterase activities. Toxicol. Lett. 78, 245-252. Pope, C. N. (1999). Organophosphorus pesticides: do they all have the same mechanism of toxicity? 1. Toxico!. Environ. Health B. Crit. Rev. 2, 161181. Pope, C. N., Chakraborti, T. K., Chapman, M. L., Farrar, J. D., and Arthun, D. (1991). Comparison of in vivo cholinesterase inhibition in neonatal and adult rats by three organophosphorothioate insecticides. Toxicology 68, 5161. Pope, C. N., Chapman, M. L., Tanaka, D. J., and Padilla, S. (1992). Phenylmethylsulfonyl fluoride alters sensitivity to organophosphorus-induced delayed neurotoxicity in developing animals. NeuroToxicology. 13,355-364. Pope, C. N., and Liu, J. (1997). Age-related differences in sensitivity to organophosphorus pesticides. Environ. Toxicol. Pharmacol. 4, 309-314. Pope, C. N., and Padilla, S. (1990). Potentiation of organophosphorus-induced delayed neurotoxicity by phenylmethylsulfonyl fluoride. 1. Toxico!. Environ. Health 31, 261-273. Pope, C. N., Tanaka, D., Jr., and Padilla, S. (1993). The role of neurotoxic esterase (NTE) in the prevention and potentiation of organophosphorusinduced delayed neurotoxicity (OPIDN). Chem. BioI. Interact. 87, 395406. Poul, J. M. (1988). Effects of perinatal ivermectin exposure on behavioral development of rats. Neurotoxico!' Teratol. 10,267-272. Rivera, S., Rosa, R., Martinez, E., Sunol, C., Serrano, M. T., Vendrell, M., Rodriguez-Farre, E., and Sanfeliu, C. (1998). Behavioral and monoaminergic changes after lindane exposure in developing rats. Neurotoxicol. Teratol. 20, 155-160. Rivera, S., Sanfeliu, C., and Rodriguez-Farre, E. (1990). Behavioral changes induced in developing rats by an early postnatal exposure to lindane. Neurotoxicol. Teratol. 12,591-595. Robinson, C. P., and Smith, P. W. (1977). Lack of involvement of monoamine oxidase inhibition in the lethality of acute poisoning by chlordimeform. 1. Toxico!. Environ. Health 3, 565-568. Romijn, H. J., Hofman, M. A., and Gramsbergen, A. (1991). At what age is the developing cerebral cortex of the rat comparable to that of the full-term newborn human baby? Early Hum. Dev. 26,61-67. Saegusa, J. (1987). Age-related susceptibility to dibromochloropropane. Toxico!. Lett. 36, 45-50.
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CHAPTER
42 Emerging Issues: Children's Exposure to Pesticides in Residential Settings John L. Adgate and Ken Sexton University of Minnesota School of Public Health
42.1 INTRODUCTION During the 1990s, there has been increasing concern about the potential effects of pesticides on children's health, much of it driven by mounting evidence from both animal toxicological studies and epidemiological investigations that children may suffer adverse health effects (e.g., neurobehavioral) from exposure to organophosphate (OP) pesticides. It is now widely recognized that health risk assessments should take special account of children because they may be both more exposed and more biologically susceptible than adults [Guzelian et aI., 1992; National Research Council (NRC), 1993; U.S. Environmental Protection Agency (U.S. EPA), 1995a]. Among the reasons children may be at potentially greater risk are their lower body weights, developing organs, higher metabolic rates, and unique behavior patterns. For example, the differences in body weight between children and adults are illustrated in Table 42.1, which summarizes average body weights of residents of the United States (U.S. EPA, 1999a). The importance of understanding children's exposure to pesticides was highlighted by the National Research Council in its 1993 report, Pesticides in the Diet of Infants and Children. The NRC recommended" ... that certain changes be made in current regulatory practice. Most importantly, estimates of expected total exposure to pesticide residues should reflect the unique characteristics of the diets of infants and children and should account also for all nondietary intake of pesticides" (NRC, 1993, p. 7). Three years later, Congress passed the Food Quality Protection Act (FQPA) of 1996 (P.L. 104-170, P.L. = Public Law), which requires that children's exposure to pesticides be evaluated for all potential pathways, both dietary (i.e., consumption of food and beverages) and nondietary (i.e., intake of pesticides in air, water, and soil or dust). The FQPA codified the need for more and better exposure data to help in the process of risk-based decision making, and mandated an examination of children's aggregate exposure, which means analysts must conduct assessments of nondietary pesticide exposure by mUltiple routes: that is, exposures by inhalation of airborne chemicals; Handbook of Pesticide Toxicology Volume 1. Principles
dermal absorption of chemicals in contact with the skin; and ingestion of chemicals in nonfoods such as soil and house dust. This chapter surveys existing and emerging methods for assessing children's nondietary exposures to pesticides in residential settings, and it comments on the implications of these methods for health risk assessment. We begin by describing basic principles of exposure assessment; then we examine the sources and pathways for children's exposure to pesticides. Methods for measuring pesticides are discussed next, followed by a discussion of biomarker measurements to estimate exposure and dose. The chapter concludes by looking ahead at the research needed to reduce uncertainties in children's risk assessments.
42.2 PRINCIPLES OF EXPOSURE ASSESSMENT Although the terms "exposure" and "dose" are well-established concepts familiar to all environmental health scientists, their meaning can vary depending on the nature of the discussion. For the purposes of exposure assessment, however, it is important that these and related terms be defined precisely in the context of an environmental health paradigm. 42.2.1 EXPOSURE AND DOSE Exposure is a deceptively simple concept, defined as contact at a body boundary between a person and an environmental stressor (biological, chemical, or physical) over time (Sexton et aI., 1995b; U.S. EPA, 1992). This simple definition masks the fact that a rigorous exposure analysis can be a complex endeavor, requiring collection and analysis of multiple variables, such as concentration and duration of exposure, as well as descriptive factors that influence contact rates and, therefore, determine the magnitude of exposure. A minimal description of exposure for a particular route (i.e., inhalation, ingestion, dermal absorption)
887
Copyright © 2001 by Academic Press. All rights of reproduction in any fonn reserved.
888
CHAPTER 42 Emerging Issues: Children's Exposure Table 42.1 Body Weight Values for Specified Age Groups in the United Statesa Body weight Age (years)
(kg)
Comments
Infants (0.5-1.5)
10
Mean of values for males and females in the 6-11
Toddlers (3)
15
Mean of values for male and female 3-year-olds
Children (6)
23
Mean of values for male and female 6-year-olds
Youth (10--12)
41
Mean of values for males and females age 10, 11,
Adult reproductive females
60
Mean for females age 13-54 years
Adults
72
Mean for males and females 18 years and older
month and I year age groups
and 12 years
aSource: Data from U.S. EPA (l999a).
must include at least two related attributes: concentration of the stressor in the carrier medium (exposure concentration); and the time of contact (duration). If the exposure concentration is integrated over the duration of contact (Eq. 1), the area under the resulting curve is the magnitude of the exposure in units of concentration times time (e.g., mg/l-day, mg/kg-day, Il-g/m3 -hr). This is the method of choice to describe and estimate short-term exposures, where integration times are on the order of minutes, hours, or days (NRC, 1991; Sexton et a!., 1995a; U.S. EPA, 1992): 12
E=
1
C(t)dt
(1)
11
where E is magnitude of exposure, t2 - t1 is duration of exposure, and C(t) is exposure concentration as a function of time. Over periods of months, years, or decades, exposures to most stressors, including pesticides, occur intermittently rather than continuously. Yet long-term health effects, such as cancer, are customarily evaluated based on average exposure or dose over the period of interest (typically years), rather than as a series of intermittent exposures. Consequently, long-term exposures or doses are usually estimated by summing across discrete exposure episodes and then calculating an average for the period of interest. Although the integration approach can also be used to estimate long-term exposures or doses, its application to time periods longer that about a week is usually difficult and inconvenient (Sexton et a!., 1995a). Duration of exposure is a key element of pesticide exposure assessment because it is directly related to adverse outcomes (U.S. EPA, 1997). Accordingly, pesticide exposures are typically divided into four general categories by duration: Acute exposures are less than a day in duration; short-term exposures last between 2 and 7 days; intermediate-term exposures last from one week to several months; and chronic exposures persist over a substantial portion of an individual's lifetime. Examples of pesticide uses that could result in acute or short-term exposures include treatments to turf in parks or applications indoors at schools; intermediate-term or chronic exposures might occur
because of agricultural pesticide residues in the food supply or continual use of pesticides inside the residence. Dose is a more complicated concept intimately related to exposure. Once a stressor, such as a chemical pesticide, enters the body it is described as a dose. Several different types of dose are relevant to exposure estimation. Potential (or administered) dose is the amount of a chemical that is actually ingested, inhaled, or in contact with the skin. Applied dose is the amount of a chemical directly in contact with the body's absorption barriers, such as the skin, respiratory tract, or gastrointestinal tract. Internal (or absorbed) dose is the amount of the chemical absorbed and, therefore, available to undergo metabolism, transport, storage, or elimination. Delivered dose, sometimes referred to as body burden, is the portion of the absorbed dose that reaches a tissue of interest, such as blood, urine, or hair. Biologically effective (or target) dose is that portion of the delivered dose that reaches a site of toxic action, such as the liver or brain (Sexton et a!., 1995a; U.S. EPA, 1992). All these dose types can be represented by a general equation (Eq. 2) that incorporates the intake and uptake factors that modify these distinct dose types (Lioy, 1990): D
=
10
= 10
1
D(t)dt 1
J(x)g(ab)p(as, rd, me, el)C(t) dt
(2)
where D is the integrated dose at a target tissue; D(t) is the time-varying function for dose; J (x) is the contact rate; g( ab) is the target organ or system specific bioavailability that determines absorption (ab); and p (as, rd, me, el) describes the nature of a contaminant's assimilation (as), cell repair or damage (rd), elimination (el), and metabolism (me).
42.2.2 THE DOMAIN OF EXPOSURE ASSESSMENT The chain of events depicted in Fig. 42.1 is an "environmental health paradigm"-a simplified representation of key steps be-
42.2 Principles of Exposure Assessment
Intake and Uptake Processes
Environmental Health Paradigm
889
Important Determinants (Mechanisms)
- eompensatlon -damage • repair
Figure 42.1 The domain of exposure assessment in relation to the environmental health paradigm. (Reproduced with permission Sexton et aI., 1995a.)
tween release of hazardous agents into the environment and potential morbidity or mortality in humans (Sexton et al., 1995b). This sequential series of steps serves as a useful construct to aid in understanding and evaluating environmental health risks. The figure also illustrates the domain of exposure assessment, which extends from identifying emission sources to characterizing the biologically effective dose, and includes developing an understanding of important events (e.g., contact between people and pollution), mechanisms (e.g., pharmacokinetics), and processes (e.g., intake, uptake). An important aspect of exposure assessment typically involves characterizing the critical "exposure pathway," which refers to the specific course of movement for a particular chemical from its source through various environmental media (i.e., air, water, soil) to ultimate contact with people. Common residential nondietary pathways for exposure to pesticides occur indoors (e.g., a consumer product applied to
a carpet results in child being exposed through hand-to-mouth activity-dose occurs via ingestion) and outdoors (e.g., application of consumer product to lawn or garden results in child being exposed through dermal contact-dose occurs by dermal absorption). It is important to remember that children can also be exposed to pesticides in many nonresidential settings, such as day care centers, schools, and playgrounds. As shown schematically in Fig. 42.1, exposure assessment is necessarily broad-based and complicated, involving acquisition and interpretation of information about key steps in the environmental health paradigm and related data on exposure factors, intake, uptake, and pharmacokinetics. Basically, assessing exposure to chemical pesticides involves the qualitative description and quantitative estimation of the chemical's contact with and entry into the body. Although no two exposure assessments are exactly alike, most address three areas that are critical for re-
890
CHAPTER 42
Emerging Issues: Children's Exposure
alistic risk assessments and contribute substantially to informed risk management decisions: (1) the number of people exposed at specific concentrations for the time period of interest; (2) the resulting dose, especially to the target tissue; and (3) the relative contributions of important sources and pathways to exposure or dose. In addition to these areas, a comprehensive exposure assessment should include an analysis of variability (e.g., within and between individuals) and uncertainty (e.g., statistical error in measurements and model parameters) (Sexton et aI., 1995a; U.S. EPA, 1992).
drawbacks of this approach are that it tends to be intrusive, resource intensive, and generally does not provide information on pathways and routes of exposure. This approach is also constrained by the lack of specific physiologically-based pharmacokinetic models for the pesticides of interest. Without a good understanding of the relevant pharmacokinetics, reconstruction of internal dose and calculation of previous exposures are highly uncertain (Sexton et aI., 1995a; U.S. EPA, 1992). These problems are even more acute for studies involving children.
42.2.3 EXPOSURE ASSESSMENT APPROACHES
42.2.3.3 Scenario-Based Assessment
In practice, exposure assessment for environmental chemicals involves use of both qualitative and quantitative data to describe contact with and entry into the human body. The quantitative estimation of chemical pesticide exposure can be approached in three ways: personal measurements, reconstructive analysis, or scenario-based assessment (Sexton et aI., 1995a; U.S. EPA, 1992). 42.2.3.1 Personal (Point-of-Contact) Measurement
Personal (or point-of-contact) measurements document exposures as they occur by measuring the pesticide concentration at the point of contact between the person and the environmental (or carrier) medium. Examples include use of pumps and filters to measure airborne concentrations near the breathing zone, duplicate diet food samples to measure dietary levels, or skin patch samples to measure dermal concentrations. The major strength of this approach is that it measures exposure directly during the monitoring period, which typically is on the order of minutes, hours, or, at the most, days. The problems with personal measurements are that they are costly and timeconsuming, can be burdensome for the study participants, and suitable monitoring devices are not available for all pesticides and pathways of interest (Sexton etal., 1995a; U.S. EPA, 1992). Because these problems are exacerbated in the case of children, personal monitoring has rarely been attempted in this subpopulation (Weaver et aI., 1998). 42.2.3.2 Reconstructive Analysis
Reconstructive exposure analysis uses measurements of dose (i.e., body burden, elimination levels), in conjunction with information or assumptions about rates of intake and uptake, to derive (or reconstruct) estimates of past pesticide exposure. Use of this approach requires valid measures of exposure biomarkers in accessible human tissues so that internal dose can be realistically reconstructed, and adequate information to estimate rates of intake, uptake, and metabolism accurately. As discussed in more detail in the section on biomonitoring, urine levels are the biomarker of choice for assessing children's pesticide exposures, largely because of the relative ease of collection. The strength of this approach is its capacity to demonstrate unequivocally that exposure and uptake have occurred. The primary
In principle, personal (point-of-contact) measurements and reconstructive analysis are complementary methods, which, because they are based on direct measurements in the exposed population, are the preferred approaches for pesticide exposure assessment. In practice, however, they are rarely used because necessary data are not available for the situations and populations of interest. Consequently, the most common approach is scenario-based exposure assessment, which requires the analyst to use available facts (e.g., environmental measurements, databases), in combination with inferences and professional judgment, to construct a plausible set of assumptions (i.e., a scenario) that describes quantitatively how contact occurs between people and chemical pesticides (Sexton et aI., 1995a). A typical scenario-based approach estimates pesticide exposure by merging two separate but essential components of exposure: (1) concentration of the chemical in the environmental (carrier) medium, estimated by using monitoring data or making assumptions about source-path way-exposure interactions; and (2) people's contact time with the carrier medium, estimated by using existing demographic, geographic, and timeactivity data or by making reasonable assumptions about activity patterns, lifestyle characteristics, residential proximity to sources, and other factors. The related doses are estimated by using knowledge and assumptions about relevant pharmacokinetic processes. Variations of the scenario-based approach include both (a) "microenvironmental" methods, which combine measurements in important microenvironments (e.g., inside the residence, outdoors in the community) with data on timeactivity patterns and (b) pathway--exposure factor ("PEF") methods, which combine measurements in important environmental media (e.g., air, water, food, soil, dust) with "off-theshelf" exposure factors (U.S. EPA, 1999a) (e.g., volume of air breathed or water consumed per day, body weight, skin surface area). Examples of different types of data used in constructing exposure scenarios are listed in Table 42.2. The primary advantage of scenario-based approaches is that they enable assessors to estimate pesticide exposure and dose in cases where data are limited or lacking (a common occurrence). On the other hand, the uncertainty introduced by the need to make assumptions and inferences in the face of inadequate or inappropriate information is also their major disadvantage. Scenario-based assessments typically do not include a complete description of the exposure and dose distribution for
42.2 Principles of Exposure Assessment
891
Table 42.2 Examples of Types of Measurements Useful in Construction of Scenarios for Pesticide Exposure Assessmentsa Additional information needed to Type of measurement
Element estimated
Examples
characterize exposure
Fixed location
Environmental medium
Air and water quality
Population location and activities
monitoring
samples used to establish
networks
relative to monitoring location; fate of pesticides over distance between
long-term media levels
monitor and variation in concentration
and trends
at point of exposure Short-term media monitoring
Environmental or ambient medium; samples used to
Special studies of spray drift, indoor air
of pesticides between measurement
establish a snapshot of
point and point of exposure; time
quality of medium over
variation of pesticide concentration at point of exposure
relatively short time Food samples
Population location and activities; fate
Concentrations of
U.S. Food and Drug
Dietary habits of various age, sex, or
contaminants in food
Administration Total
cultural groups; relationship between
supply
Diet Study, market
food items sampled and groups
basket studies, shelf
(geographic, ethnic, demographic)
studies, cooked-food
studied; relationships between
sampling
concentrations in uncooked versus prepared foods
Drinking water samples
Concentrations of
Ground Water Supply
Fate and distribution of pesticides
pesticides in drinking
Survey, Community
from point of sample to point of
water supply
Water Supply Survey, tap
consumption; population served
water samples
by specific facilities and consumption rates; for exposure due to other uses (e.g., cooking, showering); need to know activity patterns and volatilization and degradation rates
Consumer product samples
Concentration levels of the products
Shelf surveys, e.g.,
Establish use patterns and market
solvent concentration in
share of particular products,
household cleaners
individual exposure at various usage levels, extent of passive exposure
Breathing zone measurements
Exposure to airborne
Indoor air studies
Location, activities, and time spent relative to monitoring locations;
chemicals
protective measures or avoidance Microenvironmental measurements
Ambient medium in a defined area, e.g., kitchen
Special studies of indoor air, house dust, and contaminated surfaces
Surface soil samples
Degree of contamination of soil available for contact
Soil samples at contaminated sites
Activities of study populations relative to monitoring locations and time exposed Fate of pesticide on or in soil; bioavailability, contact, and ingestion rates as a function of activity patterns and age
Fish tissue samples
Extent of contamination of edible fish tissue
National Shellfish Survey
Relationship of samples to food supply of individuals or population of interest; consumption habits; preparation habits
aBased on U.S. EPA (1992).
892
CHAPTER 42
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the population of interest, instead providing only "point estimates" of specific locations on the population distribution of exposures. Historically, the points of interest on the exposure distribution have been the median or average exposures, the "high-end" exposure (the 90th percentile and above is defined as the high end of the exposure distribution), and the most exposed individual in the population. Although the scenario-based approach is used in the vast majority of pesticide exposure assessments, it is most useful when the analyst has some insight into the completeness, soundness, and validity associated with underlying assumptions and inferences, and understands their overall effect on the uncertainty of estimated values for exposure and dose. Yet despite their obvious limitations, scenariobased approaches remain the only viable method for estimating pesticide exposure and dose in the absence of direct measurements. Scenario-based approaches are specified in the Environmental Protection Agency's standard operating procedures (SOPs) for pesticide exposure assessment in residential settings, which provide standard default methods for assessment of both handler (i.e., the individual applying the pesticide) and postapplication exposures when chemical or site-specific data are limited (U.S. EPA, 1997). These SOPs are for "high-end" exposures so that the residential lawn scenario, for example, is assumed to represent the upper end of the distribution of exposures that could occur from lawns, parks, playgrounds, recreational areas, athletic fields, and other turf areas. These scenarios typically rely on one or more upper-percentile assumptions, such as the 90th percentile value for exposure duration and the 90th percentile value for skin surface area. They are intended to represent "Tier 1" assessments that can be used to indicate whether a more detailed assessment is warranted, possibly including collection of chemical-specific or site-specific data. A Tier 1 assessment uses a "conservative" screening scenario to make "worst-case" or "bounding" estimates (e.g., maximum pesticide application to 100% of turf) (U.S. EPA, 1995b, 1997). If this analysis suggests a possible problem, estimates can then be refined in subsequent tiers using progressively more realistic assumptions and values (e.g., average pesticide residues on percentage of turf actually treated). Typically, Tier 1, scenario-based exposure assessments combine information and assumptions about environmental medium concentrations, contact rate, and exposure duration using the general equation E
= ex CR x ED
(3)
where E = exposure to the chemical pesticide (mass per unit time, e.g., J.lg/day); C = concentration in environmental media (mass of substance per unit volume, e.g., J.lg/m3); CR = contact rate (volume per unit time, e.g., m3/day); and ED = exposure duration (fraction of a day exposed, e.g., minutes exposed per total minutes in a day). Better and more complete data are often available for inhalation exposures and these assessments frequently involve relatively straightforward assumptions about inhalation rates and
absorption of the target chemicals. Adequate and appropriate data are available less often for both ingestion and dermal exposures, which are also intrinsically more complicated because of difficulties quantifying relevant behaviors, activities, and physical processes that determine actual exposures. Scenario-based approaches also typically go on to calculate dose using the general equation D
=
E x AF/BW
(4)
where D = dose (in micro grams per kilogram of body weight per day, J.lg/kg-bw-day); E = exposure to a chemical pesticide (mass per unit time, e.g., J.lg/day); AF = absorption factor (unitless); BW = body weight (kg). Because population-specific and situation-specific data are rarely available to estimate parameters in the simplified exposure and dose equations, values are typically obtained by using EPA-approved "exposure factors," which are derived from an amalgamation of sources (U.S. EPA, 1999a).
42.3 SOURCES AND PATHWAYS FOR CHILDREN'S RESIDENTIAL EXPOSURE TO PESTICIDES The most repeated yet underappreciated precept in children's environmental health is the factual statement that "children are not little adults," which means that it is not appropriate to estimate exposure and dose in children by merely scaling down data from adults. Postnatally, there are at least three important differences between children and adults that are crucial for assessing exposure: factors relating to (a) sources, (b) intake, and (c) uptake (Plunkett et aI., 1992). Children tend to eat less varied and different diets than adults, they may consume food that has come in contact with the floor and other contaminated surfaces, and they are more likely to spend time on or near treated surfaces. Children have higher intake rates of air, water, soil, and food per unit of body weight and surface area than do adults. Moreover, differences in uptake and pharmacokinetics can result in children receiving a higher proportionate dose for a given exposure than adults. Although it is common to speak of children as a homogeneous group, there are important differences in exposure-related attributes by age. For example, dietary and nondietary exposures are likely to be substantially different for a 2-year-old toddler, a 6-year-old elementary school student, and a 14-yearold adolescent, even if they live in the same household. In fact, experts have identified at least seven distinct and sometimes overlapping developmental periods in the life of a child that are likely to have important effects on exposure-related attributes. These stages and their duration, important developmental milestones, and related behaviors, activities, and potential exposures are summarized in Table 42.3 (Carmichael, 1946; Freeman, 1999; Kelly, 1983; Larsen and Pascal, 1998; NRC, 1993). Although this chapter focuses primarily on exposures outside the womb, exposures prior to birth can be an important
42.3 Sources and Pathways for Children's Residential Exposure to Pesticides
893
Table 42.3 Developmental Stages in Children and Associated Milestones and Exposure-Related Behaviorsa Important developmental milestones and
Examples of potential exposures and related
stage
Time period
average age of occurrence
behaviors and activities
Embryonic
8 days to 8
Human organogenesis approximately days
Mother's acute or short-term exposures to current
Developmental
weeks of
use pesticides or release of persistent pesticide stores
20-60 of gestation
pregnancy Fetal
8 weeks of pregnancy to
Control of autonomic nervous system at 24 weeks
current use pesticides or release of persistent
birthb
Perinatal
29 weeks of pregnancy to 7
pesticides Preterm viability outside the womb 24-38
End transplacental exposure; begin breast-feeding transfer of persistent compounds and recent maternal
weeks
days after birth Neonatal
Birth-28 days
exposure to compounds with short half-lives Beginning of ocular control and head
Continued breast-feeding
motion Infancy
Birth-I 2 months
Transplacental transfer from mother's exposures to
Begin potential inhalation and dermal exposures
Rolling over at 2-3 months
Breast-feeding usually completed by end of 1st year
Sitting starts at 3 months
Transition to solid food begins at 6-9 months: low
Maturation of xenobiotic metabolism and
diet diversity leads to higher relative consumption of
elimination processes at 3 months
some foods compared to adults and older children
Standing supported at 6 months
Mouthing nonfood objects begins as early as 4
Crawling usually mastered by 9 months
months and continues to at least age 2 Dermal and inhalation exposures from sitting on
Walking begins at 10-17 months Weight triples and height increases to
~20
cm
treated surfaces
during first year Childhood
I year-12 years
Receptive to adult language and self-feeding at
~1
year
Bladder and bowel control at
Hand to mouth behavior common from ages 1-3 year olds
~2
years
Mastery of common motor skills by 5 years
Periodic consumption of nonfood objects (pica) exhibited in half of 1- to 3-year-olds Increasing time spent outside of the home
Adolescence
> 12-18 years
Maturation of organ systems to adult size and weight
Potential occupational exposures in young farm workers
aCompiled from Carmichael (1946), Freeman (1999), Guzelian et at. (1992), Kelly (1983), Larsen and Pascal (1998) and NRC (1993). bNormal term birth is at 40 ± 2 weeks of pregnancy.
health issue for which only limited exposure data exist. Typically, embryonic and fetal in utero exposures are estimated based on the mother's body burden for a particular chemical. Many children are exposed in utero to low levels of persistent pesticides (i.e., chlorinated organic compounds such as DDT and its metabolites). Although many of these pesticides have been banned in the United States for many years, they are still present in the food web due to both past and current use worldwide and their tendency to bioaccumulate as a result of their persistence in the environment. Recent studies using highly sensitive measurement techniques indicate that many of these compounds are present in adipose tissue (Kutz et aI., 1992) of U.S. residents and in the breast milk of women from both developed and less developed nations (Banerjee et aI., 1997; BasriUstunbas et aI., 1994; Dogheim et aI., 1996; Furst et aI., 1994; lohansen et aI., 1994; Klopov et aI., 1998; Kroger, 1972; Quinsey et aI., 1995; Saleh et aI., 1996; Schade and Heinzow, 1998). Residue levels measured in breast milk have been steadily declining in the United States and no studies have demonstrated adverse effects in breast-fed children. Current scientific consen-
sus supports the conclusion that the benefits of breast-feeding outweigh the risks from exposure to persistent organic pesticides in human milk (NRC, 1993). Beginning with the neonatal stage and progressing through infancy (birth to 12 months), children are influenced by a special set of pesticide sources and pathways as they experience rapid increases in mobility and important changes in behaviors and activities. For example, activities such as crawling, rolling, climbing, and finally walking during the first year or so of life have profound effects on young children's contact with surfaces, which increases the potential for dermal absorption of pesticide residues. Infants often put their hand and other objects in their mouth, resulting in non dietary ingestion from contaminated surfaces. Teething occurs largely during the first year of life and coincides with very high levels of hand-to mouth and object-to-mouth activities (Freeman, 1999). From an exposure perspective, children between ages 1 and 12 can be subdivided into four age groups [young toddlers (12-24 months), older toddlers (24-36 months), preschoolers (3-5 years), and
894
CHAPTER 42
Emerging Issues: Children's Exposure
school-aged children (>5 years)], based on behavior and activity patterns likely to influence pesticide exposures. 42.3.1 SOURCES OF NONDIETARY RESIDENTIAL PESTICIDE EXPOSURES FOR CHILDREN
For children less than 12 years old, their residence is likely to be the primary setting in which they come into contact with pesticides. Children spend an average of 21 hours a day indoors (D.S. EPA, 1999a), most of it at home, in school or in daycare. A wide variety of consumer pesticide products are used in the home, including preparations for flies, ants, roaches, and fleas, as well as products for care of pets and houseplants (Nigg et aI., 1990). Pesticide application inside residences falls generally into four categories: (1) broadcast (i.e., pressurized fan spray with dilute pesticide formulation on open surfaces), (2) crack and crevice (i.e., application method similar to broadcast but applied to these enclosed locations), (3) total aerosol releases (e.g., foggers), and (4) structural applications (e.g., treatments of walls and foundations for termites) (Fenske et aI., 1991). In addition, many cleaners and disinfectants, room deodorizers, and laundry aids also contain pesticides, though users are seldom aware of this. Typical active ingredients found in indoor application products include chlorpyrifos, propoxur, diazinon, malathion, dichlorvos, and pyrethroids (Ness, 1994). Other pesticide products are used outdoors around the residence. For example, it is common for building occupants to use insecticides on lawns and gardens, fungicides for structural protection and lawn treatments, and herbicides for weed control. Prevalent active ingredients in these products are the insecticides carbaryl, chlorpyrifos, diazinon, and malathion, fungicides captan and chlorothalonil, and herbicides such as 2,4-D and glyphosate (Ness, 1994). Probability-based sampling studies of the prevalence of pesticide storage for the D.S. population indicate that >90% of households have at least one pesticidecontaining product. In 1988, a national in-home survey found an average of 3.8 pesticide products (95th percent confidence interval 3.3-4.3) per home (Whitmore et aI., 1992). A recent probability-based survey conducted in more than 300 Minnesota households with children, representing more than 49,000 households in the census tracks sampled, reported a mean of 6.0 pesticide products stored, and a mean of 3.1 products used in the previous year (Adgate et al., 2000). Exposure and dose may be influenced by the formulation of the pesticide itself. In quantifying exposure to these compounds it is important, therefore, that the physical, chemical, and biological properties of the specific compound be examined. For example, pesticides are made and distributed in numerous formulations, some of which are designed to improve their efficacy or enhance storage or safety (Fenske, 1997; Ware, 1994). Common formulations are aerosols, dusts, or granular materials, which have varying persistence in the environment. Insecticides or fungicides applied indoors or tracked in from the outdoors on shoes or by pets can become embedded in carpets
and other surfaces indoors. Although data are limited, it is possible that pesticides in these indoor locations are protected from natural processes that might otherwise degrade them (e.g., sunlight, extremes of temperature, rainfall, and microbial action) and, therefore, may represent a long-term potential source of exposure. 42.3.2 NONDIETARY RESIDENTIAL EXPOSURE PATHWAYS FOR PESTICIDES
The sUbpopulations of children thought to be at greatest risk from residential exposures are children whose families use pesticides inside or outside the home, and those whose parents or siblings work with pesticides. Children of applicators and agricultural workers are at risk for higher exposures because of their contact with work-contaminated clothing and the increased likelihood of pesticides being tracked indoors from work applications (Fenske, 1997). Several studies have shown that airborne concentrations of pesticides outside residences are typically an order of magnitude lower than indoor levels (Nigg et aI., 1990). Investigations of children's nondietary pesticide exposures in the residential setting have focused principally on two pathways: potential indoor exposures occurring upon reentry after a pesticide has been applied ("reentry exposures"), and potential outdoor exposures as a result of garden and turf treatments. An intrinsic problem in these kinds of exposure assessments is determining the amount of pesticide product that is "dislodgeable," which is defined as the amount that can be removed from lawn and garden foliage or indoor surfaces and is therefore available for subsequent ingestion or skin absorption. Examples of possible residential exposure scenarios for children are shown in Table 42.4. Although the table is not exhaustive, it does include the residential exposure scenarios deemed most likely to cause elevated pesticide exposures in children.
42.4 PESTICIDE MEASUREMENT
METHODS RELEVANT TO SCENARIO-BASED ASSESSMENTS The most commonly used environmental sampling methods relevant to assessing exposures in children are air monitoring, measurements of house dust, and measures of dislodgeab1e residues on both indoor and outdoor surfaces. Personal (e.g., breathing zone) air samples are rarely collected for young children (1-12 years old) because of their inability to comply with extensive study protocols and the relatively large size of the monitoring equipment. Instead, air samples have been collected in important microenvironments (e.g., main activity room of the residence) using area monitors. It has been demonstrated, for example, that airborne chlorpyrifos levels after broadcast application are higher near the floor, which is in much closer proximity to a toddler's breathing zone than to an adult's (Fenske et aI., 1990).
42.4 Pesticide Measurement Methods Relevant to Scenario-Based Assessments
895
Table 42.4 Examples of Nondietary Residential Exposure Scenarios and Important Routes of Exposure for Children Exposure routes Exposure scenarios
Dermal
Investigation
Inhalation
References
x
x
x
Byme et al. (1998), Fenske et al. (1990);
x
x
Post application contact indoors Residues from crack and crevice, broadcast, or
Gurunathan et al. (1998); Ross et al. (1991)
aerosol treatments on surfaces House dust contaminated from indoor or
Bradman et al. (1997); Loewenherz et at. (1997);
structural treatments, track-in, or spray drift
Nishioka et at. (1996); Simcox et at. (1995);
from outdoor sources
Whitmore et at. (1994)
Transfer from pet treatments or flea collars
x
Insect repellents applied to clothes or skin
x
Mouthing or consumption of impregnated
Ames et at. (1989); Chambers (1996)
x
Lipscomb et at. (1992); Osimitz and Murphy (1997)
x
V.S. EPA (1997)
x
Nishioka et at. (1996); V.S. EPA (1997)
materials, such as antimicrobial shower curtains Postapplication contact outdoors Pesticide product or residues on treated turf
x
or soils Fogging or spraying for mosquitoes and other
x
Moore et at. (1993)
x
Bradman et at. (1997); Bradman et al. (1994);
pests near homes Spray drift from large-scale treatments near homes or schools
Because the behaviors and activities of young children are more likely to bring them into contact with contaminated surfaces, it is important to understand the levels of pesticides in these locations. Available methods for sampling indoor surfaces include (1) deposition pad samples, typically used on any indoor surface, (2) wipe sampling techniques, used on relatively smooth surfaces such as floors, counter tops, and window sills, and (3) vacuum techniques, which have been used to collect house dust samples from both hard floor surfaces and carpets. Deposition sampling is performed postapplication using aluminum foil (Fenske et aI., 1991), gauze (Ross et aI., 1991), or cotton cloth pads (Byrne et aI., 1998; Krieger et aI., 1997) as the collection medium. Wipe sampling techniques have been used to collect chlorpyrifos samples from deposition samplers using water as a solvent (Fenske et aI., 1991) and from surfaces using octadecylbonded silica disks and methanol and hexane as solvents (Gurunathan et aI., 1998). Choice of method used to remove pesticide residues from surfaces can have a significant effect on estimated exposures. For example, investigators (Byrne et aI., 1998) using deposition pads wiped on surfaces and toys subsequently extracted with isooctane report either nondetectable or lower chlorpyrifos exposures compared to investigators (Gurunathan et aI., 1998) using hexane to remove pesticides directly from surfaces and toys. Use of organic solvents directly on a surface apparently results in more complete removal of chlorpyrifos residues, but may overestimate doses obtained in dermal contact or hand-to-mouth activities, where the solvent would be saliva, sweat, or the sebum layer on the skin.
Loewenherz et at. (1997); Marty et at. (1994)
A number of vacuum sampling systems have been developed to collect house dust samples from carpets, rugs, and bare floors, mostly for sampling lead (Rinehart and Rogers, 1995). A specialized high-volume vacuum sampler (HVS3) was developed specifically to obtain samples of semivolatile pesticides in house dust [American Society of Testing and Materials (ASTM), 1993; Roberts et al. (1991)]. The HVS3 has been used in several field studies to collect carpet and smooth-surface samples (Bradman et aI., 1997; Lewis et al., 1994; Nishioka et aI., 1996; Simcox et aI., 1995). In a recent field test comparing this method with samples from 15 home vacuum cleaner bags, the HVS3 obtained higher upper bound concentrations for up to 26 pesticides, but prevalence of measurable pesticides and median dust concentrations were similar for both methods (Colt et aI., 1998). Three techniques have been developed to measure dislodgeable residues on indoor and outdoor surfaces and to characterize transfer from one location to another. The California Department of Food and Agriculture (CD FA) developed a 12-kg roller made of polyvinyl chloride (PVC) pipe covered with polyurethane foam (PUF), which is used to roll over cotton sheets placed on a treated surface (Ross et al., 1991). Results of this method were compared to residue levels on the clothes of adult subjects wearing dosimeter clothing while performing a set of standardized aerobic dance routines (Jazzercise) on treated surfaces (Ross et aI., 1990). The upper bound estimate of total dislodgeable residues, 1-3% of residues present on surfaces, was comparable between the CDFA roller and lazzercise methods. A second PUF roller method has been developed to simulate the force a crawling 9-kg child applies to a
896
CHAPTER 42
Emerging Issues: Children's Exposure
surface (Lewis et aI., 1994). Using this method, investigators estimated 2,4-D and dicamba track-in rates onto carpets after outdoor turf applications (Nishioka et aI., 1996). Transfer was estimated to be 3% of dislodgeable residues, which were J.1-0.2% of overall turf application levels. A subsequent study by these same investigators in 13 homes after homeowner and commercial applications indicated that children and pet activities were the most significant factors determining residue levels indoors (Nishioka et aI., 1999). A "drag sled" method has also been developed that uses a 100-cm2 patch of denim affixed to the bottom of a sledlike device, whose weight approximates the force exerted by a 1O-kg child on a surface (Byrne et aI., 1998; Vaccaro et aI., 1996). In theory, these three methods should give similar results, but no studies have directly compared these methods. Dermal absorption is a major occupational route of pesticide exposure in adults and a variety of direct and indirect methods have been developed using various dosimeters and indirect estimation methods to measure pesticide loading on the skin. Surrogate skin (Fenske, 1997) and fluorescent tracer methods (Fenske, 1988, 1990; Fenske et aI., 1986) have been used to monitor occupational exposures in adults, but their use in additional studies on children has been limited because of concerns about intentional exposure to subjects and the introduction of fluorescent tracer into the pesticide formulation prior to application is required. Hand wipe methods have been developed that use either isopropanol and gauze wipes (Geno et aI., 1996; Lewis et aI., 1994; Lu and Fenske, 1999) or a 10% isopropanoldistilled water mixture wrapped around a subject's hand using a plastic bag (Edwards and Lioy, 1999; Fenske and Lu, 1994) to remove pesticides. Although some data suggest that hand wipes may remove deeply embedded compounds that may not be removable by typical soap-and-water washing, data from controlled mass-balance experiments also suggests that dermal wash methods may significantly underestimate exposure because they typically remove approximately 20-40% of the available compound, with the remaining amounts likely absorbed through the skin (Fenske and Lu, 1994). Only one study has compared pesticide residue levels measured by wipe, roller, and hand wipe or press methods. Indoor chlorpyrifos levels after broadcast application indicate that wipe and PUF roller measurements estimate a dermal loading that is 23-36 times greater than estimates based on hand press or drag samples (Lu and Fenske, 1999). Overall, the relation between residue levels measured by these methods and the actual exposures experienced by children remain unclear, although evidence suggests the methods may provide a reasonable upper bound estimate of surface transferable residues. Information about activity patterns is an important component of scenario-based assessments that is combined with data about environmental concentrations to estimate exposure. Several small field studies have collected time-activity data for children (Lewis et aI., 1994), and most existing time-activity databases are summarized in the U.S. EPA Exposure Factors Handbook (U.S. EPA, 1999a). This important reference describes how older children (e.g., 9-11 year olds) spend their
time (Schwab et aI., 1991), but has little information about the frequency of certain important exposure-related activities for younger age groups, such as toy mouthing, consumption of nonfood items, or dermal contact rates. The National Human Activity Pattern Survey (NHAPS) is the largest probabilitybased survey ever conducted in the contiguous United States (n = 9386), but relatively few children (approximately 500) less than 4 years old were surveyed (Nelson et aI., 1994). Although NHAPS does not provide information on variability in activities by age or season of the year, it contains data important for assessing upper bound exposures associated with time spent in specific locations. These data can be useful for estimating inhalation exposures, but provide little insight for developing estimates of dermal exposure or nondietary ingestion by handto-mouth behaviors. Recently, videotaping has been used to provide a more detailed examination of children's activity patterns, including quantifying specific behaviors, such as hand-to-mouth rates, using video translation software (Zartarian et aI., 1995, 1997a, b) or trained video observers (Reed et aI., 1999). Both methods derive similar results for hand-to-mouth rates in preschool children over approximately a 1O-hr period. Combining this information with environmental media concentrations, skin surface concentrations, and data on contact and removal rates will improve exposure estimates for the hand-to-mouth ingestion pathway. Videotaping methods may eventually assist in developing more rigorous estimates for both nondietary ingestion of pesticides and dermal contact rates.
42.5 USING BIOMARKER MEASUREMENTS TO ESTIMATE PESTICIDE EXPOSURE AND DOSE Many of the dose-related steps in the environmental health paradigm (Fig. 42.1) occur at inaccessible sites in the body (e.g., liver, brain). Biological markers (biomarkers) are indicators of these significant but inaccessible events that can be measured in accessible human tissues (e.g., blood) and excreta (e.g., urine), using either invasive or noninvasive (sample collection does not require penetration of the body) methods. For example, invasive methods may involve collection of blood, lung tissue, bone marrow, amniotic fluid, liver tissue, or adipose tissue, whereas noninvasive techniques might involve collection of expired air, saliva, semen, urine, hair, feces, breast milk, skin, or fingernails. A particular biomarker may be an indicator of exposure (e.g., pesticide metabolites in urine), effect (e.g., DNA hyperploidy caused by benzidine exposure), or susceptibility (e.g., immunoglobulin levels in blood) (Sexton et aI., 1995a). For pesticide exposures, biomarker measurements can serve as exposure and dose indicators and, when linked with physiologically based pharmacokinetic models, can assist in estimating past exposure and related doses (Cashman et aI., 1996; Dong et aI., 1994, 1996). Biomarker measurements can also provide a "reality check" to verify the realism and reasonableness of scenario-based exposure assessments. The value of a
42.5 Using Biomarker Measurements to Estimate PesticideExposure and Dose
particular biomarker for exposure assessment depends generally on the half-life of the marker in the body, the specificity of the marker for the pesticide(s) of interest, the relative ease of sample collection, and the difficulty or cost of chemical analysis.
42.5.1 STUDIES MEASURING BIOMARKERS OF PESTICIDE EXPOSURE IN HUMAN POPULATIONS A 1998 review by Fenske (Fenske, 1998) of the literature on biological monitoring for pesticides found relatively few references directly relevant for assessing nonoccupational exposures. Fenske divided the relevant publications into five groups: (1) 9 review articles, reports, or compendia published in the past 10 years; (2) 8 small-scale or controlled urine monitoring studies; (3) 12 general or subpopulation urine monitoring studies; (4) 3 blood or serum monitoring studies; and (5) 7 studies measuring pesticides in breast milk and 1 in adipose tissue (Fenske, 1998). Urinary metabolites are the most commonly used biomarkers of organophosphate exposure in both adults and children. Although these samples are relatively easy to collect and analyze, their utility is limited by the lack of validated pharmacokinetic models for most compounds of interest (Woollen, 1993), and the fact that urinary metabolites are elimination products rather than direct markers of exposure or internal dose (Fenske, 1997). Urine samples are typically collected either as "spot" samples, which can be first morning voids or convenience samples taken as they become available during the day, or as aggregate 24-hr or week-long samples. Although repeat spot samples are desirable because they provide data on rate of excretion and furnish a measure of within-person variability, insuring subject compliance can be complicated and timeconsuming. A variety of methods exist for quantitative analysis of urine samples (Hill et aI., 1995b; Nigg et aI., 1990), which are normally adjusted by dividing measured biomarker concentrations by creatinine mass to normalize for differences in metabolic rate and urine dilution. Because children typically have higher creatinine excretion rates than adults due to higher metabolic and tissue turnover rates, comparing adjusted data between adults and children is not straightforward. Relatively few studies involving measurement of urinary pesticide biomarkers have been conducted in human popu1ations, and most have not collected data on children. Fenske (1998) identified 16 published studies that use urine monitoring to estimate nonoccupational pesticide exposures in the United States, Israel, and Canada (Table 42.5). Most studies have been small-scale, convenience samples of at-risk populations, such as occupants of treated residences (Esteban et aI., 1996; Richter et aI., 1992b), people likely to be exposed by aerial sprayings (Dong et aI., 1994), and families of farm workers (Shealy et aI., 1997). Larger-scale, urinary biomarker studies in adults have been conducted in conjunction with the National Health and Nutrition Examination Survey (NHANES) 11 (1976-1980)
897
(Kutz et aI., 1992) and III (1988-1991) (Hill et aI., 1995a, c), which uses a probability-based sampling scheme for the U.S. population. Results from 1000 adults in NHANES III have been used to establish "reference ranges" for 12 pesticide biomarkers, where a reference range is defined as the distribution of biomarker concentrations in a population with no known exposure or only minimal exposure to the toxic ant of interest (Hill et aI., 1995c). A reference range provides an indication of median and high-end biomarker levels in the adult D.S. population, and serves as a basis for comparison with data from children that might be collected in the future. The reference ranges show that p-dichlorobenzene is ubiquitous and most adults 29-59 years of age have been exposed to naphthalene, chlorpyrifos, and pentachlorophenol. Over the 1O-year period between samples, time trend data indicate exposures to chlorpyrifos are increasing whereas exposures to pentachlorophenol are apparently decreasing (Hill et aI., 1995c). Although a few biomarker studies have been conducted in children, virtually all utilize convenience samples of potentially at-risk groups, such as children of pesticide workers (Bradman et aI., 1997; Loewenherz et aI., 1997). One of the largest is an Arkansas study using single first-morning-void urine samples to examine 197 children thought to be exposed to chlorinated phenols and phenoxy herbicides (Hill et aI., 1989). As in adults, p-dichlorobenzene and pentachlorphenol were detected in nearly 100% of the urine samples, with median concentrations of 9 and 14 ppb, respectively. Measurable levels of some herbicide metabolites were found in more than 50% of the samples, with others present in only 10-20% of the samples. In another study of at-risk children, investigators examined dimethylphosphate metabolite levels in 160 spot urine samples from children of 48 families of pesticide applicators and 14 reference families in central Washington's applegrowing region (Loewenherz et aI., 1997). Findings demonstrated that children of pesticide applicators had significantly higher exposures than children from the reference families, even though they all lived in the same community. Results also suggest that decreasing age and closer residential proximity to orchards were associated with higher exposures. Although the study does not use a probability-based sampling frame, it does involve collection of repeat biomarker measurements over a 3- to 7-day period in each child, thereby allowing for examination of both within and between individual variability in exposures. Because neither of the two studies just described reported measurements of pesticide concentrations in all relevant environmental media, it was not possible to determine the primary pathways and routes of exposure. Persistent chlorinated pesticides or their metabolites can be measured readily in blood and serum, and these biomarkers have been used to provide an estimate oflong-term body burden (Pirkle et aI., 1995). However, most active ingredients in current use, such as organophosphate compounds and pyrethrins and pyrethoids, have relatively short half-lives in the body and are not easily measurable in blood or serum. Measurable levels of the pesticides ethion and carbaryl and the herbicide atrazine
CHAPTER 42 Emerging Issues: Children's Exposure
898
Table 42.5 Published Studies Using Urine Monitoring to Estimate Nonoccupational Exposures in the United States and Canadaa Study
Scenario
Population
Pesticides
Harris and Solomon (1992)
Controlled exposure to treated turf
Adults
2,4-D
Harris et al. (1992)
Residential exposure to treated turf
Adult applicators and bystanders
2,4-D
Richter et al. (1992b)
Indoor residential exposure from
Family: 2 adults, 2 children
Diazinon
Dong et al. (1994, 1996)
Modeling of population exposure from
Adults
Malathion
Adults
Disodium octaborate tetrahydrate
commercial application aerial application Krieger et al. (1996)
Controlled exposure to indoor carpet treatment
Shealy et al. (1997)
Farm pesticide application
Farm families
Carbaryl, dicamba, 2,4-D, others
Hill et al. (1989)
Community exposure to hazardous
Children
Phenoxy herbicides
waste site Kutz et al. (1992)
NHANES II subsample
U.S. adults
Pentachlorophenol, chlorpyrifos,
Richter et al. (1992a)
Residential treatments and agricultural
Adult kibbutz workers and
OP pesticides
Thompson and Treble (1994);
Population survey
parathion, others drift
residents Various ages
Pentachlorophenol
Treble and Thompson (1996) Hill et al. (1995a)
NHANES III subsample
1000 adults
p-Dichlorobenzene
Hill et al. (1995c)
NHANES III subsample
1000 adults
Chlorpyrifos, pentachlorophenol,
Esteban et al. (1996)
Indoor residential pest control
Residents
Methyl parathion
Davies and Peterson
Community exposure from agricultural
General population of Dade
OP pesticides, chlorpyrifos
others treatments (1997) Loewenherz et al. (1997)
and residential uses of pesticides Community exposure from agricultural pesticide use
County, FL Children 1--6 years old
Azinphos-methyl, phosmet, parathion, chlorpyrifos
aSource: Data from Fenske (1998). Reproduced with permission.
have been observed in saliva, but few human studies have been conducted and none in children (Fenske, 1997; Nigg and Wade, 1992). Beyond the NHANES reference range data, few populationbased studies of pesticide exposures, using either environmental or biomarker measurements, have been conducted and, until the late 1990s, none had been done in a sample of children. In the 1980s, the Non-Occupational Exposure Assessment Study (NOPES) was one of the first population-based studies of pesticide exposures in the U.S. (Whitmore et aI., 1994). It focused on estimating adults' exposure (using outdoor, indoor, and personal air samples, as well as some surface residue samples) to 32 pesticides in two U.S. cities: Jacksonville, Florida, selected to representative high pesticide usage; and Springfield, Massachusetts, selected to represent lowto-moderate pesticide usage. Results demonstrated that a number of widely used home and garden pesticides were detectable in indoor residential air in the two cities. Subsequent indirect analysis suggested that exposures via the dietary pathway were likely to be greater than exposures occurring by either inhalation or dermal absorption for most of the 32 pesticides examined, with cyclodiene termiticides being the notable exception.
42.5.2 THE NEXT GENERATION OF CHILDREN'S PESTICIDE EXPOSURE STUDIES
Fenske (1998) has identified 24 ongoing studies in the United States and Canada that involve biological measurements of nonoccupational pesticide exposures. The 23 that use urinary biomarkers are summarized in Table 42.6. In the near future, this next generation of biomarker studies will begin providing more and better data on children's pesticide exposures. Many of the studies focus particularly on children, including children offarm families, children living along the U.S.-Mexico border, and children living in urban and nonurban settings. Several of these studies, especially the Phase 1 field studies that are part of the National Human Exposure Assessment Survey (NHEXAS), focus on measuring multipathway exposures to mUltiple pesticides using a combination of environmental, personal, biological and time-activity measurements. The goal is to measure or estimate the contributions to exposure from important pathways and to collect and analyze data so that it is possible to assess both aggregate exposures (total exposure from ingestion, inhalation, and dermal contact for a single pesticide) and cu-
42.5 Using Biomarker Measurements to Estimate PesticideExposure and Dose
8~
Table 42.6 Current Pesticide Biological Monitoring Studies in the United States and Canadaa Study
Location
Study design
Target pesticides
Univ.ofMN
Minnesota
State population-based probability
Atrazine, malathion, chlorpyrifos, carbaryl
design Univ.ofMN
Minneapolis-St.
Urban-rural comparison of
Paul and rural
Atrazine, malathion, chlorpyrifos, carbaryl, 2,4-[
3- to 12-year-old children
counties Univ.ofMN
Minneapolis
Comparison of K-5 school
Malathion, chlorpyrifos, carbaryl
children from low-income neighborhoods Univ. of AZ
Arizona
State population-based probability
Atrazine, malathion, chlorpyrifos, carbaryl
design Univ. of AZ
U.S.-Mexico border
Geographic-based and pop.-based
Atrazine, malathion, chlorpyrifos, carbaryl
probability design Univ. of AZ
Yuma County, AZ
Subpopulation of farmworker
Chlorpyrifos, diazinon
children EmoryUniv.
Maryland
Population-based probability
Atrazine, malathion, chlorpyrifos, carbaryl
design Univ.ofWA
Chelan-Douglas and
Clinic-based urban-rural
King Counties, WA Univ.ofWA
Chelan-Douglas
Subpop. defined by proximity to
Counties, WA Univ.ofWA
Yakima County, Monterey County,
Probability-based within defined
OP pesticides (dialkylphosphates)
communities Clinic-based enrollment during
CA OR Health Science
OP pesticides (dialkylphosphates)
treated farmland
WA Univ. CA Berkeley
OP pesticides (dialkylphosphates)
comparison
OP pesticides (dialkylphosphates)
prenatal care
Oregon
Subpop. of farmworker children
OP pesticides (dialkylphosphates)
Controlled study
Adult volunteers
Chlorpyrifos
Riverside, CA
Convenience sample of pesticide-
Chlorpyrifos
Univ. Univ.ofCA Riverside Univ.ofCA
using population
Riverside Univ. of Guelph
Ontario, Canada
Convenience sample of home
Chlorpyrifos
users of pesticides National Cancer
Iowa, North
ATSDR
Subpop of farm families from
Carolina
Institute
Texas border area
Extensive pesticide screen
Agricultural Health Study Component of epidemiological study of neural tube defects
ATSDR
U.S.-Mexico border
Comparison of high and low
EPAlORD
Ohio
Convenience sample of pesticide-
Methyl parathion, atrazine, 2,4-D, chlorpyrifos, carbaryl, malathion OP pesticides (dialkylphosphates)
potential exposure (children) 2,4-D
using population EPAlORD
California and North
Clinic-based high-risk pediatric
Carolina Health Canada
Ontario, Canada
OP pesticides (dialkylphosphates)
population Probability-based sample of farm family population
a Source: Data from Fenske (1998). Reproduced with permission.
Phenoxy herbicides
900
CHAPTER 42
Emerging Issues: Children's Exposure
mulative risks (sum of health risks from exposure to multiple pesticides). The completed Minnesota Children's Pesticide Exposure Study (MNCPES) is an example of the new generation of exposure monitoring studies (Adgate et aI., 2000; Quackenboss et aI., 2000). The primary objective was to characterize children's exposure to selected pesticides through a combination of complementary monitoring approaches, including concurrent biomarker, personal, environmental, and activity-pattern measurements. The design strategy focused on testing the hypothesis that ingestion of organophosphate pesticides from the dietary pathway is the primary route of exposure for children. Emphasis was placed on measuring exposures to four pesticide compounds (i.e., chlorpyrifos, diazinon, malathion, atrazine) that were selected based on their frequent use, presence in multiple environmental media, expected population exposures, and related human toxicity. The study was conducted during the summer of 1997 and involved a stratified random sample of households with children ages 3-12 years located in either (a) the cities of Minneapolis and St. Paul (urban households) or (b) Rice and Goodhue Counties (nonurban households) just south of the metropolitan area. The results from a residential survey documenting storage and use of products containing target pesticides were used to screen homes and preferentially select households where children were likely to have higher exposures. The study suc-
cessfully obtained pesticide exposure data on lO2 children, including measurements of personal exposures (air, hand rinse, duplicate diet), environmental concentrations (residential indoor and outdoor air; drinking water; residential surfaces; soil), activity patterns (questionnaire, diary, videotaping), and internal dose (metabolites in urine and blood). MNCPES demonstrates that it is feasible and practical to use a population-based sample to obtain environmental, personal, and biological measurements, as well as time-activity data, from children and their households as part of an intensive, six-day monitoring protocol. The extensive exposure monitoring (multiple pathways, routes, and pesticides) conducted for each of the lO2 children in the MNCPES is unprecedented, and the resulting database will provide the most comprehensive and in-depth characterization of children's exposures to organophosphate pesticides ever undertaken.
42.6 LOOKING AHEAD As we enter the twenty-first century there are legitimate reasons to be concerned about possible adverse health effects in children from pesticide exposures. The overarching need right now is for sound science to help answer critical questions. Which pesticides are of primary concern based on health-based criteria? Who is exposed to levels above health-related benchmarks? Personal (point of Contact) Measurements
Environmental Concentration Measurements
I I
Outdoor
Air
I I I
t Soil (Entranceway)
Inhal~tion
Breathing Zone Air
I
I I I I I I I I
8 - - - -1---------------, I I
: ~
I
I I
r----------------------Indoor Air
Dust on Non-floor Surfaces
----r~~~~±~~~~~~:
l_ _ _ _ _ _ _ _ _ _ __ _ _ _.._ --
---------- ---- --- - ----- ---- --- --
----
Ingestion
Food, Beverages
I I I I I I I I Dermal
...
i
1
!
1
r----------'
Dermal Contact
Dust on Floors, Carpet, etc. I
Biomarker Measurements
,
Ab~orption
I
I I I I
Figure 42.2 Schematic representation of (a) important residential pesticide exposure pathways, (b) potential cross-media transfers, and (c) the relationship between environmental, personal, and biomarker measurements. Source U.S. EPA (l999b).
Body Burden (e.g., urine, blood, hair)
References
How, why, and when are they exposed to elevated levels? Which actions will be most effective and efficient in preventing or reducing exposures? Unfortunately, the data necessary to answer these and related questions with an appropriate degree of certainty are scarce, uneven, and fragmented. The 1993 NRC report Pesticides in the Diets of Infants and Children and the Food Quality Protection Act of 1996 has focused attention on the critical need to better understand children's exposure to pesticides from both dietary and nondietary pathways. Nevertheless, because of the existing paucity of adequate and appropriate data virtually all assessments of children's exposure to pesticides must resort to scenario-based approaches. The major scientific uncertainties inherent in residential pesticide exposure assessments for children include uncertainties about transfer of pesticide residues from surfaces to skin, rates of dermal absorption, ingestion of pesticide residues adhering to hands, and quantitative relationships between exposure, dose, and biomarkers. To address these and other important data needs, a variety of research approaches are required. Biomarkers must be linked to improved pharmacokinetic models that can be used to calculate internal dose and exposure, and which serve to validate exposure estimates from scenario-based approaches. Predictive multimedia, multipathway, and multichemical models need to be developed and validated to assist in characterizing aggregate exposure and cumulative risk. Firstgeneration aggregate exposure methods have been developed for some chemicals (International Life Sciences Institute, 1998; Shurdut et aI., 1998), but much more work is needed to apply them to children. These models must also address key exposure issues, such as the potential for cross-media transfer to nontarget surfaces or sinks, the timing of multiple exposures relative to one another, and the role of developmental or behavioral characteristics in determining high-end and maximally exposed individuals. The schematic in Fig. 42.2 provides a simplified representation of nondietary pesticide exposure pathways and potential cross-media transfers in the residential setting. It summarizes our current understanding of how children can be exposed to pesticides in and around their dwellings. The individual boxes indicate locations where exposure-related measurements can be made. Solid lines between or within media, such as those between outdoor air and indoor air, represent pathways driven largely by physical or chemical processes; dotted lines represent pathways and exposures that are largely influenced by human activities. Pesticides in outdoor and indoor air, tap water, dust, and soil can come into contact with children during their normal daily activities, and cause exposures through inhalation, ingestion, or dermal contact. Once a pesticide enters the body by either intake or uptake, that portion which reaches an accessible tissue or excreta of interest (i.e., delivered dose or body burden) can often be measured using sophisticated analytical techniques. To adequately characterize these various pathways and determine their relative contributions to actual exposures and doses under real-life conditions and situations, it is necessary to undertake a systematic research program involving well-
901
designed and complementary studies. As shown in Fig. 42.2, this will necessarily entail three interrelated types of measurements: (1) measurements of pesticide concentrations in environmental media; (2) measurements of personal pesticide exposures; and (3) measurements of exposure biomarkers in urine or other appropriate biological material. Although each category of information is important in its own right, the value of these data increase dramatically when they are interconnected. If, for example, the three types of measurements are conducted concurrently or in a complementary sampling frame, then the data can be used to elucidate the entire sequence of events depicted in Fig. 42.2. By collecting and analyzing matched data on environmental concentrations, personal exposures, and body burden, it becomes possible to understand the residential conditions, including critical pathways, which cause elevated pesticide exposures in children. We can then make informed choices about whether associated health risks are acceptable and, if not, which mitigation strategies are most cost-effective. Instead of asking "how much do these studies cost?" we should be asking the more important question, "can we afford not to obtain the necessary information?"
ACKNOWLEDGEMENT During the writing of this Chapter Drs. Adgate and Sexton were supported in part by U.S. EPA STAR Grant R825283 to the University of Minnesota.
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sure to organophosphate pesticides in workers and residents in Israel. Israel J. Med. Sei. 28, 584-598. Richter, E. D., Kowalski, M., Leventhal, A., Grauer, F., Marzouk, J., Brenner, S., Shkolnik, I., Lerman, S., Zahavi, H., et al. (1992b). Illness and excretion of organophosphate metabolites four months after household pest extermination. Arch. Environ. Health 47(2), 135-138. Rinehart, R., and Rogers, J. (1995). "Sampling House Dust for Lead. Basic Concepts and Literature Review." Westat Inc., RockviIIe, MD (and EPA 747-R-95-007, U.S. EPA, Washington, DC). Roberts, J. W., Budd, W. T., Ruby, M. G., Bond, A. E., Lewis, R. G., Wiener, R. W., and Camann, D. E. (1991). Development and field testing of a high volume sampler for pesticides and toxics in dust. J. Expo. Anal. Environ. Epidemiol. 1(2), 143-155. Ross, J., Fong, H. R., Thongsinthusak, T., Margetich, S., and Krieger, R. (1990). Measuring potential dermal transfer of surface pesticide residue generated from indoor fogger use: An interim report. Chemosphere 20(3-4),349-360. Ross, J., Fong, H. R., Thongsinthusak, T., Margetich, S., and Krieger, R. (1991). Measuring potential dermal transfer of surface pesticide residue generated from indoor fogger use: Using the CDFA roller method. Interim report n. Chemosphere 22(9-10), 975-984. Saleh, M., Afify, A., Ragab, A., EI-Baroty, G., Kamel, A., and EI-Sebae, A. (1996). Breast milk as a biomarker for monitoring human exposure to environmental pollutants. In "Biomarkers for Agrochemicals and Toxic Substances," ACS Symposium Series, Vol. 643, pp. 114-125. Am. Chem. Soc., Washington, DC. Schade, G., and Heinzow, B. (1998). Organochlorine pesticides and polychlorinated biphenyls in human milk of mothers living in northern Germany: Current extent of contamination, time trend from 1986 to 1997 and factors that influence the levels of contamination. Sci. Total Environ. 215(1-2), 3139. Schwab, M., Terblanche, A. P., and Spengler, J. D. (1991). Self-reported exertion levels on time/activity diaries: Application to exposure assessment. J. Expo. Anal. Environ. Epidemiol. 1(3),339-356. Sexton, K, Callahan, M. A., and Bryan, E. F. (1995a). Estimating exposure and dose to characterize health risks: The role of human tissue monitoring in exposure assessment. Environ. Health Perspeet. 103(Suppl. 3), 13-29. Sexton, K, Callahan, M. A., Bryan, E. F., Saint, e. G., and Wood, W. P. (I 995b). Informed decisions about protecting and promoting public health: Rationale for a National Human Exposure Assessment Survey. 1. Expo. Anal. Environ. Epidemiol. 5(3), 233-256. Shealy, D. B., Barr, J. R., Ashley, D. L., Patterson, D. G., Jr., Camann, D. E., and Bond, A. E. (1997). Correlation of environmental carbaryl measurements with serum and urinary I-naphthol measurements in a farmer applicator and his family. Environ. Health Perspeet. 105(5), 510-513. Shurdut, B. A., Barraj, L., and Francis, M. (1998). Aggregate exposures under the Food Quality Protection Act: An approach using chlorpyrifos. Regul. Toxieo/. Pharmacal. 28(2),165-177. Simcox, N. J., Fenske, R. A., Wolz, S. A., Lee, I.-e., and Kalman, D. A. (1995). Pesticides in household dust and soil: Exposure pathways for children of agricultural families. Enviran. Health Perspeet. 103(12), 1126-1134. Thompson, T. S., and Treble, R. G. (1994). Preliminary results of a survey of pentachlorophenol levels in human urine. Bull. Environ. Contam. Taxiea/. 53(2), 274-279. Treble, R. G., and Thompson, T. S. (1996). Normal values for pentachlorophenol in urine samples collected from a general population. J. Anal. Taxieol. 20(5),313-317. U.S. Environmental Protection Agency (U.S. EPA) (1992). Guidelines for exposure assessment. Federal Register 57(104),22888-22938. U.S. Environmental Protection Agency (U.S. EPA) (1995a). "Guidance for Risk Characterization." Science Policy Council, U.S. Environmental Protection Agency, Washington, De. U.S. Environmental Protection Agency (U.S. EPA) (l995b). "Series 875: Occupational and Residential Exposure Test Guidelines, Group B-Post Application Exposure Monitoring Test Guidelines (Working Draft, Version 5.1)." Office of Prevention, Pesticides, and Toxic Substances, U.S. EPA, Washington, De.
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Emerging Issues: Children's Exposure
V.S. Environmental Protection Agency (V.S. EPA) (1997). "Standard Operating Procedures (SOPs) for Residential Exposure Assessments." Office of Pesticide Programs, Health Effects Division, V.S. EPA, Washignton, DC., and Versar, Inc. V.S. Environmental Protection Agency (V.S. EPA) (1999a). "Exposure Factors Handbook." EPA/600/C-99/00I, Office of Research and Development, V.S. Environmental Protection Agency, Washington, DC. V.S. Environmental Protection Agency (V.S. EPA) (1999b). "FIFRA Scientific Advisory Panel Meeting, February 23-24, 1999. Session Ill-A Set of Scientific Issues Being Considered by the Environmental Protection Agency Regarding: Consultation on Development of Draft Aggregate Exposure Assessment Guidance Document for combining Exposure from Multiple Sources and Routes." SAP Report 99-02C, Scientific Advisory Panel, Washington, DC. Vaccaro, J., Nolan, R., Murphy, P., and Berbrich, D. (1996). "The Vse of a Vnique Study Design to Estimate Exposure to Adults and Children to Surface and Airborne Chemicals," pp. 166-183. ASTM Spec. Tech. Pub!. 1287, Am. Soc. Testing and Materials (ASTM), Philadelphia. Ware, G. W. (1994). 'The Pesticide Book." Thomson, Fresno, CA. Weaver, V. M., Buckley, T. J., and Groopman, J. D. (1998). Approaches to environmental exposure assessment in children. Environ. Health Perspect. l06(Supp!. 3), 827-832.
Whitmore, R. W., Immerman, F. W., Camann, D. E., Bond, A. E., Lewis, R. G., and Schaum, J. L. (1994). Non-occupational exposures to pesticides for residents of two V.S. cities. Arch. Environ. Contam. Toxicol. 26(1),47-59. Whitmore, R. w., Kelly, J. E., and Reading, P. L. (1992). "Executive Summary, results, and Recommendations," National Home and Garden Pesticide Vse Survey. Final Report, Vo!. 1. NTIS PB92-1747471NZ, prepared by Research Triangle Institute for Office of Pesticides and Toxic Substances, V.S. EPA, Washington, DC. Woollen, B. H. (1993). Biological monitoring for pesticide absorption. Ann. Occup. Hyg. 37(5), 525-540. Zartarian, V. G., Ferguson, A. C., and Ledde, J. O. (1997a). Quantified dermal activity data from a four-child pilot field study. J. Expo. Anal. Environ. Epidemiol. 7(4), 543-552. [Erratum 1. Expo. Anal. Environ. Epidemiol. (1998) 8(1), 109.] Zartarian, V. G., Ferguson, A. c., Ong, C. G., and Ledde, J. O. (1997b). Quantifying videotaped activity patterns: Video translation software and training methodologies. J. Expo. Anal. Environ. Epidemiol. 7(4), 535-542. Zartarian, V. G., Streicker, J., Rivera, A., Cornejo, C. S., Molina, S., Valadez, O. E, and Leckie, 1. O. (1995). A pilot study to collect micro-activity data of two- to four-year-old farm labor children in Salinas Valley, California. J. Expo. Anal. Environ. Epidemiol. 5(1), 21-34.
CHAPTER
43 Pesticide Percutaneous Absorption and Decontamination Ronald C. Wester and Howard 1. Maibach University of California, San Francisco
Human skin is a primary body organ that contacts the environment and is a route by which pesticides enter the body. Pesticides, in actuality, are designed poisons intended to protect agricultural production and home, work, and play environments. Contact with humans is not a question of if, but of when and how much, a question regulators and risk assessors have been addressing. Human skin is designed to retain body fluids and act as a barrier to the environment. This barrier can be crossed by a process called percutaneous absorption. Human absorption differs in regions of the body, and clothing can be a partial barrier. However, pesticides have managed to cross those barriers, leading to illness and death. The major processes and outcomes are discussed here.
Figure 43.1 shows human systemic parathion absorption from dermal exposure. Parathion is predicted to be lethal not only for total systemic absorption but also for exposure to limited regions. The LD 50 used for parathion is 14 mg/kg. Given a body weight of 70 kg, systemic absorption of 980 mg might result in 50% mortality. Thus, parathion lethal toxicity levels can be reached at 8-hr and longer exposures. This was unfortunately validated in the agricultural fields of California and elsewhere.
43.2 PERCUTANEOUS ABSORPTION METHODOLOGY 43.2.1 ABSOLUTE TOPICAL BIOAVAILABILITY
43.1 INTRODUCTION Percutaneous absorption is a primary focal point for dermatotoxicology and dermatopharmacology. Local and systemic toxicity depend on a chemical penetrating the skin. The skin is both a barrier to absorption and a primary route to the systemic circulation. The skin's barrier properties are impressive. Fluids and precious chemicals are reasonably retained within the body; at the same time hundreds of foreign chemicals are restricted from entering the systemic circulation. Even with these impressive barrier properties, the skin is a primary body organ that contacts the environment and is a route by which many chemicals enter the body. Some chemicals applied to the skin have proved to be toxic. These include pesticides which in actuality are designed poisons. Table 43.1 summarizes the 30-year lesson with parathion. Absorption of parathion was established for human skin contact, but other species similarly absorb the compound. Mathematical models based on quantitative structure-activity relationships now can predict a human skin permeability coefficient, but the accuracy of the predicted coefficient is not fully validated to in vivo man. Skin absorption amounts combined with toxicity data can predict potential human health hazard. Handbook of Pesticide Toxicology
Volume 1. Principles
The only way to determine the absolute bioavailability of a topically applied compound is to measure the compound by specific assay in blood or urine after topical and intravenous administration. This is extremely difficult to do in plasma because concentrations after topical administration are often low. However, as advances in analytical methodology bring forth more sensitive assays, estimates of absolute topical bioavailability will become more available (Wester and Maibach, 1999).
43.2.2 RADIOACTIVITY IN EXCRETA Percutaneous absorption in vivo is usually determined by the indirect method of measuring radioactivity in excreta after topical application of the labeled compound. In human studies, plasma levels of compound are extremely low after topical application, often below assay detection level, so it is necessary to use trace methodology. The compound, usually labeled with 14C or tritium, is applied and the total amount of radioactivity excreted in urine or urine plus feces determined. The amount of radioactivity retained in the body or excreted by some route not assayed (C02, sweat) is corrected for by determining the amount of radioactivity excreted after parenteral administration. This final
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Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
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CHAPTER 43
Percutaneous Absorption and Decontamination
Table 43.1 Summary of Parathion Percutaneous Absorption Parathion
o ,O-Diethyl O-(4-nitrophenyl) phosphorothioate Other names: ethylparathione, parathion-ethyl
CAS: 56-38-2; mol. wt. 291.26 Molecular formula: ClOH14N05PS Nonsystemic contact and stomach-acting insecticide and acaricide with some fumigant action Nonphytotoxic except to some ornamentals and under certain weather conditions; absorption takes place readily through any portal; fatal human poisoning has followed skin exposure Skin absorption (species) Human (forearm): solvent, acetone;
Mouse (dermal): no solvent; 1.4%, 1 hr (excretion analysis)
10%,5 days (excretion analysis)a
Mouse (dermal): solvent, acetone; 32%, days (patch)d Human: solvent, acetone; forearm 8.6%,
Frog (dermal): solvent, acetone; 33%, 1 hr (patch)e
palm 11.8%, foot 13.5%, abdomen 18.5%, hand, dorsum 21.0%, forehead 36.3%, axilla 64.0%,
Quail (dermal): solvent, acetone; 40%, 1 hr (patch)'
jaw 33.9%, fossal cubitalis 28.4%, scalp 32.1 %, ear canaI46.6%, scrotum 101.6%b Rat (dermal): solvent, acetone; 59%, 1 hr
Roach (dermal): solvent, acetone; 14%, 1 hr (patch),
(patchY
Hornworm (dermal): solvent, acetone; 8%, 1 hr (patch), Skin absorption (mathematical model) kp (crnlhr)
LogP(Kow)
1.59 x 10- 2
3.83
= -2.74 + [0.71 x log P(Kow )]- [0.0061 x MW], where kp = permeability coefficient, = partition coefficient in octanol compared to water, MW = molecular weight!
based on the formula logkp P(Kow )
Toxicity Rat: Oral, male LD 50: 13-15 mg/kg
Skin, male LD 50: 21 mg/kg
Oral, female LD 50: 3-3.6 mg/kg
Skin, female LD 50: 6.8 mg/kg
aFeldmann and Maibach (1974). bMaibach et al. (1974). cKnaak et al. (1984). dMarty (1976) eShah et al. (1983). !Guy and Potts (1992).
amount of radioactivity is then expressed as the percentage of applied dose that was absorbed (Feldmann and Maibach, 1974). The equation used to determine percutaneous absorption is absorption (%) total radioactivity after topical administration --------------------------------- x 100 total radioactivity after parenteral administration
taining the chemical that has been absorbed through skin. The skin flap can be used to study percutaneous absorption in vivo or in vitro. The absorption of chemicals through skin and metabolism within the skin can be determined by assay of the perfusate (Wester and Maibach, 1997). 43.2.4 STRIPPING METHOD
43.2.3 SKIN FLAPS The methodology is to isolate surgically a portion of skin so that a singular blood supply is created to collect blood con-
The stripping method determines the concentration of chemical in the stratum corneum during an application period and predicts the percutaneous absorption of that chemical. The chemi-
43.3 Regional Variation in Percutaneous Absorption Distribution of systemic absorption Compound name: Parathion Dose: 4 \lg/cm' on whole body area (1.8 m') continuous/infinite dose application Exposure: 24 hours
Total systemic absorption: 4.918 grams· Head, Neck and Arms = 1.116 grams· Estimated systemic LD50 of Parathion is 980 mg (human, 70 kg) • indicates 50% lethality dose
Figure 43.1 Simulated parathion human skin exposure to regions of the body. As early as 8 hr following exposure lethality is possible. At 24 hr, lethality is possible if only certain body regions are exposed, for example, head and neck of a field worker.
cal is applied to skin of animals or humans, and at various skin application times the stratum corneum is removed by successive tape application and removal. The tape strippings are assayed for chemical content. Pharmacokinetic Cmax , Tmax, and area-under-curve (AUC) parameters can be calculated for stratum corneum bioavailability.
43.2.5 BIOLOGICAL RESPONSE
Another in vivo method of estimating absorption is to use a biological or pharmacological response. Here, a biological assay is substituted for a chemical assay and absorption is estimated. An obvious disadvantage to the use of a biological response is that it is only good for compounds that will elicit an easily measurable response. An example of a biological response would be the vasoconstrictor assay in which the blanching effect of one compound is compared to that of a known compound. This method is perhaps more qualitative than quantitative. The best known use of this method is in the comparison of various hydrocortisone products for skin dermatitis (Wester and Maibach, 1997).
907
43.2.6 IN VITRO METHODOLOGY In vitro percutaneous absorption is best done with human skin. The skin should be used as soon as possible and stored in the refrigerator no longer than 7 days. In vitro penetration into skin gives results suitable for distinguishing drug formulations, especially in cases where the drug will not partition into reservoir fluid. Material balance in an in vitro study design adds to the overall data presentation. In vivo verification of skin absorption, preferably in humans, adds relevance to the in vitro data. The human skin sample can be kept viable if stored properly in the refrigerator (freezing kills skin viability) and used appropriately (Wester et aI., 1984). Table 43.2 gives the in vitro human skin and in vivo percutaneous absorption of several chemicals from a variety of vehicles. The in vitro absorption is divided into skin content and receptor fluid (either buffered saline or human plasma) accumulation. Generally, receptor fluid accumulation does not agree with in vivo percutaneous absorption. The reason for this is lack of solubility in the receptor fluid. In some cases, skin content (see DDT) was reflective of in vivo absorption because the chemical was able to penetrate skin (and, lacking solubility, failed to partition into receptor fluid). Chemicals with high log P (octanol:water partition coefficient) will not partition into receptor fluid (Wester and Maibach, 1997, 1999).
43.3 REGIONAL VARIATION IN HUMAN AND ANIMAL PESTICIDE PERCUTANEOUS ABSORPTION Feldmann and Maibach (1967) first explored the potential for regional variation in percutaneous absorption. The first absorption studies were done with ventral forearm, because this site is convenient to use. However, skin exposure to chemicals exists over the entire body. They first showed regional variation with the absorption of parathion (Fig. 43.2). The scrotum was the highest-absorbing skin site (scrotal cancer in chimney sweeps was the key to identifying this fact). Skin absorption was lowest for the sole and highest around the head and face. Table 43.3 gives the effect of anatomical region on the percutaneous absorption of pesticides in humans (Maibach et aI., 1971). There are two major points in this study. First, regional variation was confirmed with the different chemicals; note that parathion and malathion are chemically related to some chemical warfare agents. Second, those skin areas that would be exposed to the pesticides, the head and face, were of the higher absorbing sites. Body areas most exposed to environmental contaminants are among the areas with the higher skin absorption. Table 43.4 gives site variability for parathion skin absorption with time. Soap-and-water wash in the first few minutes after exposure is not a perfect decontaminant. Site variation is
908
CHAPTER 43
Percutaneous Absorption and Decontamination
Ta ble 43.2 In vitro versus in vivo Percutaneous Absorptiona Percentage dose In vitro Compound
Logpa
DOT
6.9
Vehicle
Skin
Acetone Soil
Benzo [a ]pyrene
5.97
Acetone
Chlordane
5.58
Acetone
Pentachlorophenol
5. 12
Acetone
Receptor fl uid
In vivo
18.1 ± 13.4
0.08 ± 0.02
18.9 ± 9.4
1.0 ± 0.7 23.7 ± 9.7
0.04 ± 0.01 0.09 ± 0.06
1.4 ± 0.9 10.8 ± 8.2 0.3 ± 0.3
0.0 1 ± 0.06 0.Q7 ±0.06 0.04 ± 0.05
3.3 ± 51.0 ± 13.2 ± 6.0 ± 4.2 ±
0.6 ± 0.09 0.01 ±O.OO
29.2 ± 5.8 24.4 ± 6.4
Soil Soil
3.7 ± 1.7 0.11 ± 0.04
Soil PCBs (1242)
21.4 ± 8.5 18.0 ± 8.3 20.8 ± 8.3
Acetone
High
TeB
PCBs (1254)
M ineral oil
6.4 ± 6.3
Soil
1.6 ± 1.1
0.3 ± 0.6 0.04 ± 0.05
10.0 ± 16.5
0. 1 ± 0.07
2.8 ± 2.8
0.04 ± 0.05
1.6 ± 0.2 1.0 ± 1.0 0.3 ±0.2 6.7 ± 4.8 0.09 ± 0.03 28.5 ± 6.3 7.9 ± 2.2
0.02 ± 0.01 0.9 ± l.l 0.4 ± 0.5 0.4 ± 0.2 0.03 ± 0.02 0.Q7 ± 0.01
14.1 ± 1.0 14.6 ±3.6
Acetone
High
28.0 ± 8.3 20.4 ± 8.5
T CB Mineral oil Soil 2,4 Dichlorophenoxy-
2.81
13.8 ± 8.6 ± 15.9 ± 2.0 ± 3.2 ±
Acetone
acetic acid (2,4-0)
Soil
Arsenic
Water
Cadmium
Water
Soil Soil Water
Mercury
Soil
0.5 22.0 3.4 2.8 1.8
2.7 2.1 4.7 1.2 1.9
0.06 ±0.01
aNote that a log P of 6 means that 106 (1,000,000) molecules will partition into octanol for each (1) molecule which will partition into water.
200
0
w CD a:
•m
0
(J)
et
w (J) 0
FOFEARM
100
w <-' ~ z w u a: w Cl.
~ AlDCJt,.EN 0 HAND
•m
FOSSA CUBIT A1..IS
EJ
JAW
fA 0
POSTAURICULAA
•m
Percentage of dose absorbed
PALM
m RXJT
CD
0
Table 43.3 Effect of Anatomical Region on in vivo Percutaneous Absorption in Humans
SCALP
FOReiEAD
EAR CANAL
[3 AXILLA scroTUM
Anatomical region
Different parts of the body vary in percutaneous absorption. T his is an important consideration in risk assessment.
Malathion
1.0
Palm
8.6 11.6
6.8 5.8
Foot, ball
0.8 0.2
13.5
6.8
Abdomen
1.3
18.5
9.4
7.6
21.0 36.3
12.5 23.2
64.0
28.7 69.9
Hand, dorsum Forehead Axilla
3.1 12.2
Scalp Scrotum
33.9
4.4
28.4 32.1 46.6
36.2
101.6
Fossal cubitalis Ear canal
apparent early in skin exposure (Wester and Maibach, 1985). Decontamination is discussed later.
Parathion
Forearm
Jaw angle
Figure 43.2
Hydrocortisone
Guy and Maibach (1985) took the hydrocortisone and pesticide data and constructed penetration indices for five anatomi-
43.3 Regional Variation in Percutaneous Absorption
cal sites (Table 43.5). The indices might be used with their total surface areas (Table 43.6) when estimating systemic availability relative to body exposure sites. Van Rooy et al. (1993) applied coal-tar ointment to various skin areas of volunteers and determined absorption of polycyclic aromatic hydrocarbons by surface disappearance of (PAH) and the excretion of urinary I-OH pyrene. Using PAH disappearance, skin ranking (highest to lowest) was shoulder> forearm> forehead> groin> hand (palmar) > ankle. Using
Table 43.4 Site Variation and Decontamination Time for Parathion Parathion dose absorbeda (%)
Skin residence time
Palm
before soap-and-water wash
Arm
Forehead
I min
2.8
8.4
6.7
7.1 12.2
13.3
Ihr
8.4
10.5
11.7
4 hr
8.0
27.7
7.7
24hr
8.6
36.3
11.8
6.2
5min 15 min 30min
13.6
aEach time is a mean for four volunteers. The fact that there were different volunteers at each time point accounts for some of the variability with time for each skin site. Table 43.5 Penetration Indices for Five Anatomical Sites Assessed Using Hydrocortisone Skin Penetration Data and Pesticide (Malathion and Palathion) Absorption Results Penetration index based on Site
Hydrocortisone data
Pesticide data
Genitals
40
12
Arms
I
Legs
0.5
I
Trunk
2.5
3
Head
5
4
I-OH pyrene excretion, skin ranking (highest to lowest) was neck> calf> forearm> trunk> hand. Table 43.7 compares their results with Guy and Maibach (1985). Wester et al. (1984) determined the percutaneous absorption of paraquat in humans. Absorption was the same for the leg (0.29 ± 0.02%), hand (0.23 ± 1%), and forearms (0.29 ± 0.1 %). Here, the chemical nature of the low-absorbing paraquat overcame regional variation. Skin absorption in the Rhesus monkey is considered to be relevant to that of humans. Table 43.8 shows the percutaneous absorption of testosterone (Wester et aI., 1980), fenitrothion, aminocarb, and diethyltoluamide (DEET) (Moody and Franklin, 1987; Moody et aI., 1998), in the Rhesus monkey compared with the rat. What is interesting is that, for the Rhesus monkey, there is regional variation between forehead (scalp) and forearm. If one determines the ratio of forehead (scalp) to forearm for the Rhesus monkey and compares the results with those for the human, they are found to be similar (Table 43.9). Therefore, the Rhesus monkey can be a relevant animal model for human skin regional variation.
Table 43.7 Absorption Indices of Hydrocortisone and Pesticides (ParathionlMalathion) Calculated by Guy and Maibach (1985) Compared with Absorption of Pyrene and PAH for Different Anatomical Sites by Van Rooy et al. (1993) Absorption index Anatomical site
Hydrocortisonea
Pesticides b
Genital
40
12
Pyrenec
PAHd
I
Arms
0.8
0.5
Hand
I
Legs, ankle
0.5
I
1.2
0.8,0.5
Trunk, shoulder
2.5
3
1.1
-,2.0
Head, neck
5
4
-,1.3
1.0
aBased on hydrocortisone penetration data (Feldmann and Maibach, 1967). bBased on parathion and malathion absorption data (Maibach et aI., 1971). CBased on excreted amount of I-OH-pyrene in urine after coal-tar ointment application (Van Rooy et al., 1993). dBased on the PAH absorption rate constant (Ka) after coatar ointment application (Van Rooy et al., 1993).
Table 43.6 Body Surface Areas Distributed over Five Anatomical Regions for Adults and Neonate Neonate
Adult Anatomical region
Body area (%)a
Area (cm2 )
Body area (%)
180
I
19
Arms
18
3,240
19
365
Genital Legs
36
6,480
30
576
Trunk
36
6,480
31
595
Head
9
1,620
19
TOTAL aNote the "rule of 9" when trying to remember human body surface areas.
909
18,000
365 1,920
910
CHAPTER 43
Percutaneous Absorption and Decontamination
43.4 PERCUTANEOUS ABSORPTION FROM CHEMICALS IN CLOTHING Chemicals in cloth cause cutaneous effects. For example, Hatch and Maibach (1986) reported that chemicals added to cloth in 10 finish categories (dye, wrinkle resistance, water repellency, soil release, and so on) caused irritation and allergic contact dermatitis, atopic dermatitis exacerbation, and urticarial and phototoxic skin responses. This is qualitative information that chemicals will transfer from cloth to skin in vivo in humans. Quantitative data are lacking. Snodgrass (1992) studied permethrin transfer from treated cloth to rabbit skin in vivo. Transfer was quantitative but less than expected. Interestingly, permethrin remained within the cloth after detergent laundering. In other studies (Wester et at., 1996), in vitro percutaneous absorption of glyphosate and malathion through human skin were decreased when added to cloth (the cloth then placed Table 43.8 Percutaneous Absorption of Fenitrothion, Aminocarb, DEET, and Testosterone in Rhesus Monkey and Rat
on skin) and this absorption decreased as time passed over 48 hr (Table 43.10). It is assumed that, with time, the chemical will sequester into deep empty spaces of the fabric, or some type of bonding will be established between chemical and fabric. When water was added to glyphosate-cloth and water/ethanol to malathion-cloth, the percutaneous absorption increased (malathion to levels from solution). This perhaps reflects clinical situations where dermatitis occurs most frequently in human sweating areas (axilla, crotch). The clothing must not be a collection system for pesticides and it cannot be assumed that laundering will remove the agents.
43.5 SKIN DECONTAMINATION Although decontamination of a chemical from the skin is commonly done by washing with soap and water, as it has been assumed that washing will remove the chemical, recent evidence Table 43.9 Percutaneous Absorption Ratio for Scalp and Forehead to Forearm in Humans and Rhesus Monkeys
Applied dose absorbed (5) skin site Percutaneous absorption ratio Chemical
Species
Forehead
Forearm
Fenitrothion
Rhesus
49
21
Aminocarb
Rhesus
74
37
Rat
84
Rat Testosterone
Rhesus
DEET
Rhesus
Back
88
20.4a
8.8
Rat
47.4 33
14
Rat
36
Chemical
Species
Scalp/forehead
Forehead/forearm
Hydrocortisone
Human
3.5
6.0
Benzoic acid
Human
2.9
Parathion
Human
Malathion
Human
3.7
4.2
Testosterone
Rhesus
Fenitrothion
Rhesus
2.3
Aminocarb
Rhesus
2.0
DEET
Rhesus
2.4
3.4 2.3
a Scalp.
Table 43.10 In Vitro Percutaneous Absorption of Glyphosate and Malathion from Cloth through Human Skina Percentage of dose Chemical Glyphosate
Malathion
Donor conditions
Treatment
absorbed
I % solution (water)
None
1.42 ± 0.25
I % solution on cloth
Ohr
0.74 ±0.26
I % solution on cloth
24hr
0.08 ±0.01
I % solution on cloth
48 hr
0.08 ±0.01
I % solution on cloth
Add water
0.36 ±O.07
I % solution (water/ethanol)
None
8.77 ± 1.43
I % solution on cloth
ohr
3.92± 0.49
I % solution on cloth
24hr
0.62 ± 0.11
I % solution on cloth
48 hr
0.60± 0.14
1% solution on cloth
Add water/ethanol
7.34± 0.61
aBoth glyphosate and malathion in solution (treatment = none) are absorbed through human skin. GIyphosate and malathion on cotton cloth show absorption in skin, depending upon the time the chemical was added to cloth (treatment = 0, 24, 48 hr). When the cloth was wetted (treatment = add water/ethanol), the transfer of glyphosate and malathion from cloth to human skin was increased. This suggests that sweating, skin oil, or even rain may facilitate transfer of chemicals from cloth to skin.
43.6 Discussion 100
• ~
80
w
w
~ Z
~
5% SOAP 50%SOAP
POLYPR:lP'VLENE -DTAM
60
--0--
W
CORN OIl
o
z
W
-
-
Cl
60
Cl
0
~-. -_-=:;~:;;::oc::::::::--
8
Cl)
W
- - WATER-ONLY
80
(/)
0
C
100
SOAP&WATER WATER ONLY
W
911
a:
w
40
Q.
a:
w
40
Il..
20
o 0 3
6
2
4 6 TIME (HOURS)
8
10
24
TIME (HOURS)
Figure 43.3 Alachlor is a lipophilic chemical which is better removed from skin by soap and water than by water only.
suggests that many times the skin and the body are unknowingly subjected to enhanced penetration and systemic absorption or toxicity because the decontamination procedure does not work or may actually enhance absorption. Figure 43.3 illustrates skin decontamination of alachlor with soap and water or with water only over a 24-hr dosing period, using grid methodology. Note that the amount recovered decreases over time, which happens because this is an in vivo system and percutaneous absorption is taking place, decreasing the amount of chemical on the skin surface. There also may be loss due to skin desquamation. A second observation is that alachlor is more readily removed with soap-and-water wash than with water only. The reason is alachlor is lipid soluble and needs the surfactant system for more successful decontamination (Wester et al., 1991, 1992). In the preceding illustration, decrease in alachlor wash recovery over time was thought to be due to ongoing absorption and loss due to skin desquamation. These factors are probably true, but are probably not the main reason, which is soap-andwater wash effectiveness. In the home and workplace, decontamination of a chemical from skin is traditionally done with a soap-and-water wash, although some workplaces may have emergency showers. It has been assumed that these procedures are effective, yet workplace illness and even death occur from chemical contamination. Water, or soap and water, may not be the most effective means of skin decontamination, particularly for fat-soluble materials. A study was undertaken to help determine whether there are more effective means of removing methylene bisphenyl isocyanate (MDI) from the skin. MD! is an industrial chemical for which skin decontamination, using traditional soap and water and nontraditional polypropylene glycol, a polyglycol-based cleanser (DTAM), and corn oil were all tried in vivo on the Rhesus monkey, over 8 hr (Fig. 43.4). Water, alone and with soap (5 and 50% soap), was partially effective in the first hour after exposure, removing 51-69% of the ap-
Figure 43.4 Mean percentage of applied dose of MD! removed with designated decontamination procedure at designated time period. Water, and soap and water are the least effective, especially at 4 and 8 hr.
plied dose. However, decontamination fell to 40-52% at 4 hr and 29-46% by 8 hr. Thus, the majority of MD! was not removed by the traditional soap-and-water wash; skin tape stripping after washing confirmed that MD! was still on the skin. In contrast, polypropylene glycol, DTAM, and corn oil all removed 68-86% of the MD! in the first hour, 74-79% at 4 hr, and 72-86% at 8 hr. Statistically, polypropylene glycol, DTAM, and corn oil were all better (p < 0.05) than soap and water at 4 and 8 hr after dose application. These results indicate that a traditional soap-and-water wash and the emergency water shower are relatively ineffective at removing MDI from the skin. More effective decontamination procedures, as shown here, are available. These procedures are consistent with the partial miscibility of MD! in corn oil and polyglycols (Wester et aI., 1999). Thus, if there is skin contamination with a pesticide and the skin is washed with soap and water, it cannot be assumed that the pesticide has been removed from the skin.
43.6 DISCUSSION The lesson partially learned but still ongoing is that pesticide use can achieve its chemically intended goals, but that continued knowledge in human risk assessment needs to be achieved. Understanding percutaneous absorption as a major route of pesticides entering the body is an integral part of the risk assessment process. Data in humans can be achieved safely using trace-measurement methodology, and with low-risk doses coupled with high-tech analytical methodology. Data from animal and computer models are simplier to use. Safety is debatable if the models are not validated to humans, because the resulting risk assessment may also be wrong.
912
CHAPTER 43
Percutaneous Absorption and Decontamination
REFERENCES DuBois, K. P., Doull, J., Salerno, P. R., Coon, J. M. (1949). Studies on the toxicity and mechanism of action of p-nitrophenyl diethyl thionophosphate (parathion). 1. Pharmacal. Exp. Ther. 95, 79-91. Feldmann, R. J., and Maibach, H.I. (1967). Regional variation in percutaneous penetration of [14C] cortisol in man. 1. Invest. Dermatal. 48, 181-183. Feldmann, R. 1., and Maibach, H. I. (1974). Percutaneous penetration of some pesticides and herbicides in man. Taxieal. Appl. Pharmacal. 28, 126. Gaines, T. B. (1960). The acute toxicity of pesticides to rats. Taxieol. App!. Pharmaea!' 2, 88-99. Guy, R. H., and Maibach, H. I. (1985). Calculations of body exposure from percutaneous absorption data. In "Percutaneous Absorption" (R. Bronaugh and H. Maibach, eds.), pp. 461-466. Dekker, New York. Guy, R. H., and Plotts, R.O. (1992). Structure-permeability relationships in percutaneous penetration. 1. Pharm. Sci. 81, 603-604. Hatch, K. L., and Maibach, H. I. (1986). Textile chemical finish dermatitis. Cantact Dermatitis 14, 1-13. Knaak, J. B., Yee, K., Ackerman, C. R., Zweig, G., Foy, D. M., and Wilson, B. W. (1984). Percutaneous absorption and dermal dose cholinesterase response studies with parathion and carbaryl in the rat. Taxiea!. Appl. Pharmacal. 76, 252-263. Maibach. H. I., Feldmann, R. J., Milby, T. H., and Sert, W. R. (1971). Regional variation in percutaneous penetration in man. Arch. Enviran. Health 23,208-211. Maibach, H. I. (1974). Systemic absorption of pesticides through the skin of man. Occupational Exposure to Pesticides: Federal Working Group Pest Management. 120-127. Marty J. P. (1976). Fixation des substances chimiques dans les structures superficielles de la pesu: Importance dans Ies problemes de decontamination et de biodosponibilIite. Ph.D. Thesis, University of Paris-Sud, Paris. Moody, R. P., and Franklin, C. A. (1987). Percutaneous absorption of the insecticides fenitrothion and aminocarb. 1. Taxieol. Environ. Health 20, 209-219. Moody, R. P., Benoit, F. M., Reedle, D., Retter, L., and Franklin, C. (1998). Dermal absorption of the insect repellent DEET in rats and monkeys: Effect of anatomic site and multiple exposure. Personal communication. Shah, P. v., Montoe, R. J., and Guthrie, F. E. (1983). Comparative penetration of insecticides in target and non-target species. Drug Chem. Taxiea!. 6, 155179.
Snodgrass, H. L. (1992). Permethrin transfer from treated cloth to the skin surface: Potential for exposure in humans. 1. Taxieol. Environ. Health 35, 912915. Van Rooy, T. G. M., De Roos, J. H. c., Bodelier-Bode, M. D., and Jongeneelen, F. J. (1993). Absorption of polycyclic aromatic hydrocarbons through human skin: Differences between anatomic sites and individuals. 1. Taxica!. Environ. Health 38, 355-368. Wester, R. C., and Maibach, H. I. (1985). In viva percutaneous absorption and decontamination of pesticides in human. 1. Taxieol. Enviran. Health 16, 2537. Wester, R. c., and Maibach, H. I. (1997) Toxicokinetics: Dermal exposure and absorption of toxicants. In "Comprehensive Toxicology" (1. Bond, ed.), Vol. 1, pp. 99-114. Elsevier Science, Oxford, U.K. Wester, R. c., and Maibach, H. I. (1999). In vivo methods for percutaneous absorption measurements. In "Percutaneous Absorption" (R. Bronaugh and H. Maibach, eds.), 3rd ed., pp. 215-227. Dekker, New York. Wester, R. C., Noonan, P. K., and Maibach, H. I. (1980). Variation on percutaneous absorption of testosterone in the Rhesus monkey due to anatomic site of application and frequency of application. Arch. Dermatol. Res. 267, 229-235. Wester, R. c., Maibach, H. I., Buchs, D. A. w., and Aufrere, M. B. (1984). In vivo percutaneous absorption of paraquat from hand, leg and forearm of humans. 1. Taxiea!. Environ. Health 14,759-762. Wester, R. c., Melendres, J., Sarason, R., McMaster, J., and Maibach, H. I. (1991). Glyphosate skin binding, absorption, residual tissue distribution, and skin decontamination. Fundam. App!. Taxieo!. 16,725-732. Wester, R. c., Melendres, J., and Maibach, H. I. (1992). In vivo percutaneous absorption of alachlor in rhesus monkey. 1. Taxiea!. Environ. Health 36, 1-12. Wester, R. C., Quan, D., and Maibach, H. I. (1996). In vitro percutaneous absorption of model compounds glyphosate and malathion from cotton fabric into and through human skin. Faod Chem. Toxieol. 34,731-735. Wester, R. c., Christoffel, J., Hartway, T., Poblete, N., Maibach, H. I., and Forsell, J. (1997). Human cadaver skin viability for in vitro percutaneous absorption: Storage and detrimental effects of heat-separation and freezing. Pharm. Res. IS, 82-84. Wester, R. C., Hui, X., Landry, T., and Maibach, H. I. (1999). In vivo skin decontamination of methylene bisphenyl isocyanate (MDI): Soap and water ineffective compared to polypropylene glycol, polyglycol-based cleanser, and corn oil. Taxieol. Sei. 48, 1-4.
CHAPTER
44 Chemistry of Organophosphorus Insecticides Roward W. Chambers, J. Scott Boone, Russell L. Carr, and Janice E. Chambers Mississippi State University
44.1 INTRODUCTION It should be emphasized at the outset that this review is not for
the serious organic chemist, but for toxicologists and others desiring a basic knowledge of this important class of toxicants. Further, it is recognized that organic phosphorus chemicals have many uses other than as pesticides. Only insecticides will be considered here, with passing reference to chemical warfare agents. Finally, organophosphorus chemistry is an exceedingly complex subject and only generalizations will be made in the sections on synthesis and reactions. Other reviews (cited in the following) and numerous research articles are available that provide more detailed information.
44.2 CHEMISTRY OF ORGANOPHOSPHORUS INSECTICIDES 44.2.1 HISTORY Organophosphorus (OP) chemistry apparently began around 1820 with the esterification of alcohols to phosphoric acid. Despite synthesis of a number of OP compounds in the early 1900s, the potential toxicity went unrecognized until the 1930s. By 1940, groups led by B. C. Saunders in England and Gerhard Schrader in Germany had produced several highly toxic compounds for possible use as chemical warfare agents. The most notable of these were sarin and soman, both phosphorofiuoridates, and tabun, a phosphorocyanidate (Fig. 44.1). Schrader's group also produced the first commercial OP insecticides, including tepp in 1937, dimefox in 1940, schradan (OMPA) in 1942, and parathion in 1944. Following World War 11, with the capture of Schrader's research records, interest in OP insecticides grew rapidly. Although all of the early chemicals were effective insecticides, they were also highly toxic to mammals. In 1950, however, American Cyanamid produced malathion, one of the safest OPs ever marketed. By 1959, it was estimated that more than 50,000 Handbook of Pesticide Toxicology Volume 2. Agents
OP compounds had been made. In 1970, more than 200 OP insecticides were marketed worldwide. Though development of resistance in pests and the marketing of new, safer insecticides have greatly decreased the usage of OPs and slowed the development of new products, these remain an important group of pest control agents and will probably do so for another decade or more. 44.2.2 CLASSIFICATION AND ~OMENCLATURE
Because of the vast number and wide variety of organic phosphorus chemicals that can exist, any comprehensive classification system would be too complex to undertake here. Instead, a classification scheme will be presented into which all commercially important OP insecticides will fit, based on the central phosphorus atom and the four atoms immediately surrounding it. Similarly, the nomenclature of OPs will primarily address these five atoms. The general structure of OP insecticides can be represented by
L, the so-called leaving group, is the most reactive and most variable substituent. The term "leaving group" comes from the fact that it is the substituent that is displaced when the OP phosphorylates acetylcholinesterase, the primary target enzyme. The leaving group is also usually the most susceptible to hydrolysis. RI and R2 are less reactive and are most commonly alkoxy groups, but may be alkyl, aryl, alkylthio, or alkylamino. X is either oxygen or sulfur. OP insecticides may be considered to be derivatives of phosphoric acid (H3P04) or phosphonic acid (H3P03) in which all H atoms are replaced by organic moieties. Thus, phosphates are compounds in which the P atom is surrounded by
913
Copyright © 200 1 by Academic Press. An rights of reproduction in any form reserved.
914
CHAPTER 44
Chemistry of Organophosphorus Insecticides
Sarin
Soman
Figure 44.1
Structures of early OP nerve gases.
Table 44.1 Subclasses of OP Insecticides Atoms around P Name
0
S
N
C
Phosphate
4
0
Phosphorothioate*
3
0 0
Phosphorodithioate*
2
2
0
0 0 0
Phosphoramidate
3
0
I
0
Phosphorodiamidate
2
0
2
Phosphoramidothioate*
2
I
0 0
Phosphonate
3
0
Phosphonothioate
2
0 0
2
0
Phosphonodithioate
Tabun
four 0 atoms. In phosphonates, there are three 0 atoms and one phosphorus-carbon bond. Phosphinates, which have two 0 atoms and two P-C bonds, have been investigated and show biological activity, but none have been developed commercially. In many OPs, one or more of the oxygen atoms are replaced by sulfur and/or nitrogen. For phosphoric acid derivatives, the 0, S, and N atoms can be arranged in 20 different configurations. Another 12 different configurations can exist for derivatives of phosphonic acid. Fortunately for the classification and nomenclature to be considered here, most of the 32 possible configurations have not appeared in commercial OP insecticides. Table 44.1 lists the nine subclasses of OPs for which one or more commercial insecticides are known and the numbers and types of atom attached to the phosphorus. Subclasses with names followed by an asterisk are represented by compounds with two different configurations. For example, the S atom in a phosphorothioate may be within an ester bond (formerly called a phosphorothiolate) or may be doubly bonded to the P atom (a phosphorothionate). Similarly, a phosphorodithioate may be either a phosphorodithiolate or a phosphorothiolothionate. Recent literature has largely abandoned the thiolo and thiono terminology, however, and the positions of the atoms are designated using 0- and S- prefixes before substituent names. Thus, (Me-Oh P(:S) is named O,O,O-trimethyl phosphorothioate, whereas (Me-0)2 P(:O)-S-Me is O,O,S-trimethyl phosphorothioate. In the
first case, because the name designates a single S atom but none is within the ester linkages, the S must be doubly bonded to the P. In the second case, because the lone S atom is within an ester linkage, the doubly bonded atom must beO. For phosphonates and phosphonothioates, the name of the substituent attached by a P-C bond is included as a single word with the root name. (Me-Oh P(:O)- Me, therefore, is dimethyl methylphosphonate and (Me-Oh P(:S)-Et would be named 0,0 -dimethy1ethylphosphonothioate. Similarly, for phosphoramides with substituents on the nitrogen atom, the names of those substituents immediately precede the phosphorus term as a single word. In this case, those names are preceded by N - to designate their attachment to the nitrogen atom. For example, (Et-Oh P(:S)-N(Meh would be named 0, O-diethyl N ,N -dimethylphosphoramidothioate. To further clarify the structures and nomenclature of the major subclasses of OPs, representatives of each are shown in Fig. 44.2. As will be noted, for the phosphate, phosphonate, and phosphoramidate subclasses, groups attached to oxygen atoms are not preceded by 0-. In these cases, using such a prefix is unnecessary because no sulfur atoms are present in the molecules.
44.2.3 SYNTHESIS
Because of the wide variety of organophosphorus compounds known, no attempt will be made here to cover the many different pathways involved in their synthesis. Rather, the preparation of the seven most important intermediates will be presented, followed by representative examples of their use in the synthesis of specific insecticides. For additional information, the reader is referred to more comprehensive reviews by Eto (1961) and Fest and Schmidt (1982). Initially, elemental phosphorus is converted into P2 Ss by reaction with sulfur or into PCl3 by direct chlorination. These two materials are then converted into the following seven interme-
44.2 Chemistry of Organophosphorus Insecticides
diIrethyI2,2-dichlorovinyl phosphate
O,O-diIrethyl O-(4-nitrophenyl) phosphorothioate
o
o,,1I
CH3
/ p - S--CH 2-
CH2- S-CH 2CH 3
O,O-diIrethyl S-(2-ethylthio )ethyl phosphorodithioate
CH 30
S
CH 3CH 20
,,11
/ p - S - CH 2 -
S-CH2CH3
O,O-diethyl S-ethylthiorrethyl phosphorodithioate
CH 3CH 20
o
,,11
CH3CH2CH2S
/ P - 0 - C H 2CH 3
O-ethyl S,S-dipropyl phosphorodithioate
CH 3CH 2CH 2S
rrethy12-chloro-4-t-buty lpheny I N-rrethylphosphoramidate
CH3NH
o ,,~-~
~
CH 3NH/
pheny IN,N-diIrethy lphos phorodiamidate
S
o,,1IP - y~ Cl
CH 3
CH 3CHNH
-
/
I
O-rrethyl O-(2,4-dichlorophenyl) N-isopropylphosphoramidothioate
Cl
CH3
o
o,,1I
CH 3
O,S-diIrethyl phosphoramidothioate
/P-NH2 CH 3S
o CH 30 "
OH
11 I
Cl
I
P-CH-C-Cl
I
CH30/
diIrethy 12,2,2-trichloro-l-hydroxyethylphosphonate
Cl
O-ethyl O-(4-nitrophenyl) phenylphosphonothioate
S CH 3CH 20
,,~-S~ /
CH 3CH 2
Figure 44.2
~
O-ethyl S-phenyl ethylphosphonodithioate
Nomenclature of major subclasses of OP insecticides.
915
916
CHAPTER 44
Chemistry of Organophosphorus Insecticides
formation of the thiol ester. Disulfoton is produced in this way:
diates from which most OP insecticides are synthesized: S 11
P2 S S + 4 R-OH
1
2 (R-OhP-SH
S
S
11
11
(R-OhP-SH + Y2CI2
1
(R-OhP-CI
(1)
(2)
o 11
PCI) + Y202 - - _ I PCI3 S
(3)
(4) (5) (6)
o 11
1
(R-OhP-CI
0
I
11
0
CH 30-P + HCCCI3 - - _ I
I
CH 3CH 20
CH 3CH 20
,/'" P-S-CH 2 CH 2 SCH 2 CH 3
The free acid will add across certain carbon-carbon double bonds. Malathion is prepared from diethyl maleate by this reaction: S O S 0 11 11 CH 3 0 11 11 )P-SH+CHC-OCH 2CH 3 )P-S-CHC-OCH 2 CH 3 11 CH 30 CH 30 CHC-OCH 2 CH 3 CH 2 C-OCH,CH 3
CH 30
)
11 0
o
"
Also, the free acid reacts with formaldehyde and a mercaptan to produce compounds such as phorate: S
S
cH 3cH,o,1I CH 3 CH 20 , 11 /P-SH +CH 20 + HS-CH 2 CH 3 ------/P-S-CH 2 -S-CH 2 CH 3 CH 3CH 2 0 CH 3 CH 2 0
For OPs with three different substituents on the P atom, it is necessary to begin with P(:O) Cl3 (from intermediate 3) or P(:S) Cl3 (from intermediate 4). The sequence in which the groups are added varies with the specific compound. Two examples, fenamiphos and sulprofos, respectively, are as follows: ~
~H3
PC]3 + (1) HO~S-CH3
+ (2) CH 3CH 2 -OH
11
CH 3 0
11
(7)
Because there are almost as many procedures for formation of the P-C bond of phosphonates and related OPs as there are insecticides containing this bond, the synthesis of required intermediates will not be discussed. It may be noted, however, that most processes involve PCl3, AICb, and alkyl halides. Trialkyl phosphites (intermediate 5) are particularly useful in the preparation of dialkyl vinyl phosphates from a-chloroaldehydes and ketones. The synthesis of DDVP serves as a good example: CH 30
S CH 3CH 20 , 11
/ ' P- S-Na + CI-CH2 CH 2SCH 2 CH 3 - - + -
I
11
PCI) + S--+-I PCI3 PCI) + 3 R-OH hase 1 (R-OhP PCI3 + 3 R-OH DO ha:lhl (R-OhP-OH
PCI3 + 2 R-OH + Y202
S CH 3CH 20 , 11
P-O-CH=CCI2
CH 3 0
+ (3) CH 3CH-NH 2 - - -
I
CH 3
CH 3 0
Interestingly, the same aldehyde reacts with dimethyl phosphite (intermediate 6) to produce the phosphonate insecticide trichlorfon:
o 11 P-OH + HCCCI3- - -
CH 3 0 CH 3 0
)
Dialkyl phosphates may also be prepared from dialkyl phosphorochloridates (intermediate 7) by reactions analogous to the formation of carboxylic esters from acid chlorides. Paraoxon, the oxygen analog and active metabolite of parathion, is readily prepared by this reaction:
An analogous, and more commonly used, reaction is that of dialkyl phosphorochloridothioates (intermediate 2) to produce phosphorothionates such as parathion: S
CH3CH 20 , 11 ---o-~ N0 . / P-Cl + HO 2 CH 3CH 2 0 -
S
base
~
CH 3CH20 , 11 ---o-~ N0 . / P-O 2 CH 3CH 2 0 -
Dialkyl phosphorodithioates undergo three distinct types of reactions. The salts of these acids react with alkyl halides with
Although the preceding discussion is by no means a comprehensive treatment of OP synthesis, the reactions shown are used in the preparation of more than 90% of commercial OP insecticide compounds. One final type of reaction not previously mentioned is worthy of consideration because of its usefulness in laboratory syntheses. Dialkyl phosphates with phenolic or heterocyc1ic leaving groups are easily prepared from the phenol or heterocyc1ic alcohol and the dialkyl phosphorochloridate. Unfortunately, the only phosphorochloridates readily available commercially are the dimethyl and the diethyl, the former being quite unstable. At least four dialkyl phosphites (dimethyl, diethyl, di-i-propyl, and di-n-butyl) are available at reasonable prices and are more stable in storage. The reaction alluded to in the previous paragraph, then, is the synthesis of dialkyl phosphates from dialkyl phosphites without independent preparation and isolation of the phosphorochloridate. Carbon tetrachloride, in the presence of an organic base (e.g., triethylamine), will chlorinate dialkyl phosphites by the
References following reaction:
o
0
11 (CH 3 CH 2 h N 11 (R-OhP-H + cC4 - - - - _ 1 (R-OhP-CI + CHCl3
This allows a one-step synthesis of many phenyl or heterocyclic dialkyl phosphates. For example, methyl paraoxon (catalog price> $500/g) is prepared easily and inexpensively by: o
CH 3
0,1I -o-~ NO, /p-o
CH 3 0
-
This synthesis has been done several times by the authors, obtaining moderate yield of product of > 98% purity. Although the process is useful on a small scale, it is rarely used in industrial syntheses because of the hazard associated with CCk
44.2.4 REACTIONS OP insecticides, when kept cool, dark, and anhydrous, are usually quite stable. Exposure to heat, light (especially ultraviolet), and/or water, however, may lead to chemical alterations. The three primary reactions involving the phosphorus atom and those immediately surrounding it are hydrolysis, oxidation, and rearrangement. Except at very low pHs, hydrolysis of the P-O-C linkage results primarily by OH- attack on the P atom with cleavage of the P-O bond. Thus, rates of hydrolysis increase with increasing pH. Three additional generalizations may be made concerning this type of hydrolysis: 1. Compounds containing P=O hydrolyze faster than analogous compounds containing P=S. 2. Cleavage occurs between the P atom and the leaving group. 3. The hydrolysis rate decreases with increasing size of the alkyl substituents.
The most notable exception is that alkaline hydrolysis of OPs containing Me-O-P(:S)= often results in cleavage of the methyl rather than the leaving group. Hydrolysis of the P-S-C linkage differs from that described previously in that alkaline hydrolysis results primarily in cleavage of the S-C bond. An exception to this is fonofos (O-ethyl S-phenyl ethylphosphonodithioate), which is apparently hydrolyzed at the P-S bond to yield thiophenol. Acid
917
hydrolysis, on the other hand, consistently leads to cleavage of the P-S bond. The P-N bond of phosphoramides is generally rather resistant to alkaline hydrolysis. Because the N atom is readily protonated, however, these compounds are quite susceptible to acid hydrolysis. Oxidation of OPs by 02 is enhanced by ultraviolet (UV) light. It may also be accomplished by oxidants such as HN03 or organic peroxyacids. These oxidations most commonly occur with P=S-type OPs, but the sulfur of the P-S-C linkage may also be oxidized. Oxidation of P=S presumably results in transient formation of the phosphooxythiirane (a three-membered ring consisting of one each of P, 0, and S). This intermediate spontaneously decomposes to produce the oxon (P=O) with loss of sulfur or cleaves between the P atom and the leaving group. Oxidation of the sulfur of P-S-C produces ester-sulfoxides, which are highly reactive and usually degrade rapidly. Upon exposure to UV or high temperatures, compounds containing C-O-P=S will undergo rearrangement in which C-S-P=O is produced. Most commonly, the C involved is in an alkyl substituent but the leaving group can also be involved. Both types of rearrangements are known for parathion as shown by
o -o-~ cH,cH,o,1I /p-s NO, CH,CH,O -
Whether such re arrangements are intramolecular, intermolecular, or both is unclear. Toxicologically, the chemical reactions mayor may not result in loss of toxicity. Hydrolysis completely detoxifies the OP. Oxidation of P=S to P=O and both rearrangements illustrated lead to an increase in toxicity. Though other reactions of OPs may occur, those presented are the most important in the environment and in long-term storage.
REFERENCES Eto, E. (1961). "Organophosphorus Pesticides: Organic and Biological Chemistry." CRC Press, Cleveland. Fest, c., and Schmidt, K.-J. (1982). "The Chemistry of Organic Pesticides," 2nd ed. Springer-Verlag, BerlinlHeidelberglNew York.
CHAPTER
45 The Metabolism of Organophosphorus Insecticides J anice E. Chambers, Russell L. Carr, J. Scott Boone, and Howard W. Chambers Mississippi State University
45.1 INTRODUCTION As was illustrated in the previous chapter, the class of organophosphorus insecticides contains a diverse array of structures, all united by the presence of a pentavalent phosphorus atom with three singly bonded constituents and a coordinate covalent bond (typically drawn as a double bond) to either a sulfur or an oxygen. These insecticides or their metabolites are potent inhibitors of serine esterases through phosphorylation of the serine hydroxyl moiety within the active site of the esterase. The primary target esterase from a toxicological standpoint is acetylcholinesterase, a widely distributed enzyme within the vertebrate nervous system that mediates hydrolysis of the neurotransmitter acetylcholine throughout the central and peripheral nervous systems. The phosphorylation of acetylcholinesterase is relatively persistent, with spontaneous hydrolysis, and therefore recovery of the enzyme activity, requiring hours to days. The inhibition of acetylcholinesterase results in the accumulation of acetylcholine in cholinergic synapses and neuromuscular/glandular junctions with subsequent hypercholinergic activity. Such activity leads to a variety of signs and symptoms of intoxication, with death in mammals in lethal level poisonings resulting from respiratory failure. Other serine esterases, such as butyrylcholinesterase or carboxylesterases, can also be phosphorylated by the organophosphorus insecticides or their metabolites; however, phosphorylation of these other targets does not appear to result in toxic responses. More detailed descriptions of the neurotoxicity of organophosphorus insecticides can be found in a number of chapters and reviews, including Chambers (1992) and Ecobichon (1996). The potency of the organophosphorus insecticides or their active metabolites as inhibitors of target brain acetylcholinesterase does not correspond to the acute toxicity levels, indicating that metabolism and disposition are of great significance in determining the overall acute Handbook of Pesticide Toxicology Vulume 2. Agents
toxicity level of these insecticides (Chambers et at., 1990). An overview of organophosphorus insecticide chemistry, biochemistry, and toxicology can be found in Chambers and Levi (1992). These insecticides display substantial chemical diversity, including a variety of atoms in addition to the carbon and phosphorus mandated by the compounds being "organophosphorus" compounds, such as su1fur, nitrogen, and oxygen. Therefore, the organophosphorus insecticides are subject to many metabolic pathways mediated by several of the groups of xenobiotic metabolizing enzymes. The group connected through the single bond that is the least thermodynamically stable of the three single bonds is the "leaving group" and is the group that is eliminated from the molecule as it phosphorylates its esterase targets. The leaving group may be subject to some metabolic pathways that the other substituents are not, as will be described later on. The other two substituents may be the same as one another or different and are also subject to metabolism. The organophosphorus insecticides or their metabolites are subject to oxidations, reductions, hydrolyses, and conjugations. Because of their metabolic and chemical lability, they do not readily remain intact either in the environment or in the organism. Their environmentallability was one of the factors that allowed them to replace the highly stable organochlorine insecticides as the dominant class of insecticides. The several types of reactions that occur in the metabolic pathways of organophosphorus insecticides will be discussed later, indicating the types of enzymes involved in the reaction, some examples of these reactions with specific organophosphorus insecticides, and the toxicological outcomes of these metabolic pathways. An overview of the types of biotransformation enzymes involved in the metabolism of organophosphorus insecticides can be found in Parkinson (1996). Specific metabolic pathways for a number of specific insecticides can be found in Aizawa (1982,1989) and Dikshith (1991).
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Copyright © 2001 by Academic Press. All nghts of reproductIon in any form reserved.
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CHAPTER 45 The Metabolism of Organophosphorus Insecticides
45.2 OXIDATIONS 45.2.1 CYTOCHROMES P450
The cytochrome P4S0 enzymes (P4S0s) comprise one of the most important, if not the most important, class of xenobiotic metabolizing enzymes. The P4S0s are a superfamily of related enzymes characterized by the presence of a heme iron in the active site. They have a broad substrate specificity, with different enzymes (isoforms) catalyzing a variety of oxidations on phosphorus, sulfur, nitrogen, and carbon, with different substrate specificities among the isoforms. The P4S0s are monooxygenases and catalyze the oxidations by the addition of one atom of molecular oxygen into the substrate via an electron transport pathway with the electrons supplied by reduced nicotinamide adenine dinucleotide phosphate (NADPH) and sometimes reduced nicotinamide adenine dinucleotide (NADH). The electron transfers are catalyzed by NADPH cytochrome P4S0 reductase. Except for some specialized P4S0s, such as those involved in steroidogenesis, the P4S0 pathway occurs in the endoplasmic reticulum of vertebrate cells, with the highest xenobiotic metabolizing capacity in the mammalian liver and a more limited capacity observed in other mammalian tissues and in sub mammalian vertebrates. More detailed descriptions of P4S0 and associated reactions may be found in Parkinson (1996). Many of the P4S0-mediated reactions on organophosphorus insecticides are obvious oxidations, resulting in a more highly oxidized product with the presence of the oxygen apparent in the products. However, some of the P4S0-mediated reactions are not as obviously oxidations. Because of the addition of polar reactive groups by these P4S0-mediated reactions, some of the resultant products are more biologically reactive and therefore more toxic [such as with greater reactivity toward neural target molecules or toward deoxyribonucleic acid (DNA)] than the parent compounds were, whereas other reactions result in detoxified products; therefore, P4S0s mediate both bioactivations and detoxications. One of the most important of the P4S0-mediated reactions involving the organophosphorus insecticides is the desulfuration reaction occurring with phosphorothionates and other compounds having the phosphorus bonded to sulfur by a coordinate covalent bond (P=S). The desulfuration reaction involves an attack of the phosphorus by oxygen, resulting in a putative phosphooxythiiran intermediate, which rearranges to a P=O group with a loss ofthe sulfur as illustrated in Fig. 4S.1 (Neal, 1980). The sulfur released in the desulfuration reaction is a reactive moiety and has the ability to destroy surrounding biomolecules, such as the P4S0s. A classic example of this desulfuration reaction is the conversion of parathion to its phosphate (oxon) metabolite, paraoxon. The P=S compounds are relatively poor anticholinesterases, whereas the oxons are potent anticholinesterases, with a three order of magnitude difference in potency with at least some of the insecticides (Forsyth and Chambers, 1989). Because so many of the most popular of the organophosphorus insecticides are
phosphorothionates or related P=S compounds, the desulfuration reaction is required for them to display appreciable anticholinesterase activity, and therefore to display classical organophosphorus insecticide neurotoxicity. Because so many of the phosphorothionate insecticides display very high acute toxicity levels (e.g., rat oral LDSOs for parathion, methyl parathion, and azinphosmethyl are 2, SO, and 10 mg/kg, respectively; Meister, 1990), one can infer that the de sulfuration reaction occurs in vivo to an appreciable extent. An example of the desulfuration of parathion is given in Fig.4S.2. A reaction occurring parallel to the desulfuration reaction, and concurrently with it, is the dearylation reaction, occasionally termed oxidative hydrolysis, which occurs from the same putative phosphooxythiiran intermediate described previously for the desulfuration reaction. Instead of rearranging to eliminate the sulfur as occurs during the desulfuration reaction, the rearrangement in the dearylation reaction eliminates the aryl leaving group. The resultant products are the leaving group plus either the dialkyl phosphorothioate or the dialkyl phosphate; therefore, the reaction resembles a hydrolysis, but the occurrence of these reaction products is dependent on the presence of P4S0, oxygen, and NADPH. Additionally, classic hydrolysis reactions do not readily occur with the phosphorothionates. Therefore, the dearylation reaction appears to be an oxidation although the products do not readily suggest that an oxidation has occurred. Because the phosphate/oxon structure is required for anticholinesterase activity, the dearylation reaction is a detoxication reaction. The dearylation reaction with parathion as an example is given in Fig.4S.3. The concurrent and competing reactions of desulfuration (activation) and dearylation (detoxication), again using parathion as an example, are illustrated in Fig. 4S.4, along with the putative phosphooxythiiran intermediate. Studies conducted with purified isoforms of P4S0 have indicated that different isoforms have different desulfuration to dearylation ratios, indicating that substrate specificity and pathway preference among the P4S0 isoforms differ (Levi et aI., 1988). Therefore, it is expected that the activity of different isoforms of P4S0 would have different impacts on toxicity. An additional P4S0-mediated reaction that can occur on the intact phosphorothionate or its oxon is a dealkylation reaction in which one of the carbons in an alkoxy group is oxidized to the aldehyde that is removed, leaving a hydroxyl group associated with the phosphorus (Appleton and Nakatsugawa, 1972). The oxidized product would be formaldehyde in the case of a methoxy group or acetaldehyde in the case of an ethoxy group. Using parathion as an example once again, the dealkylation reaction is illustrated in Fig.4S.S. Oxidations of substituents in the leaving group are also possible, with a wide variety of reactions possible because of the great diversity of the leaving groups within the insecticide class. A few illustrative examples are provided in Fig.4S.6.
45.2 Oxidations
Figure 45.1
Desulfuration reaction of a phosphorothionate, illustrating the phosphooxythiiran intermediate.
Paraoxon
Parathion Figure 45.2
Desulfuration (activation) of parathion.
+ diethyl phosphorothioic acid or diethyl phosphate
Parathion Figure 45.3
4-nitrophenol
Oxidative dearylation (detoxication) of parathion.
Figure 45.4 Oxidation of parathion through both the desulfuration and the dearylation pathways, arising from a common intermediate.
+ Parathion Figure 45.5
acetaldehyde Oxidative deethylation of parathion.
desethyl Parathion
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CHAPTER 45
The Metabolism of Organophosphorus Insecticides
Phosphamidon
Fenitrothion
Diazinon
Dimethoate Figure 45.6
Examples of the oxidation of substituents within organophosphorus compounds.
45.2.2 FLAVIN MONOOXYGENASES
An additional class of monooxygenases capable of oxidizing N, P, or S occurring in xenobiotics are the flavin monooxygenases (FMOs), which are also microsomal, most prevalent in the mammalian liver, insert one atom of molecular oxygen into the substrate molecule, and require NADPH (Levi and Hodgson, 1992). They have a flavin group instead of a heme group to catalyze the substrate oxidations. There are fewer isoforms of the FMOs than the P450s, and they have more limited substrate specificity than the P450s. One of the more important reactions catalyzed by the FMOs is the sulfoxidation of phorate to the sulfoxide then to the sulfone (Fig. 45.7). The sulfoxidaton of the oxon of phorate to its sulfoxide, then to its sulfone, is also possible.
45.3 REDUCTIONS Reductions are possible outcomes of P450-mediated reactions, though these would be considered rare in mammalian systems, which are oxidizing environments with few exceptions (e.g., gut contents). Nevertheless, reductive reaction products have been discovered and might be expected to occur to a limited extent. Reduction reactions are probably of greater significance environmentally because of the greater opportunity to provide reducing environments. Using, once again, parathion as an example, the nitro group on the aromatic ring of the leaving group can be reduced, yielding amino-parathion (Fig. 45.8).
45.4 HYDROLYSIS Catalytic hydrolysis of the phosphates/oxons with elimination of the leaving group is catalyzed by the A-esterases (phos-
45.4 Hydrolysis
923
•
!
Phorate
Figure 45.7
Sulfoxidation of phorate to its sulfoxide and sulfone metabolites .
• Anrinoparathion
Parathion Figure 45.8
Reduction of the nitro group within parathion.
photriesterases), which are hydrolases designated as capable of hydrolyzing organophosphates and not being inhibited by them (Aldridge, 1953). In the mammalian system with the insecticidal compounds or oxons, the A-esterases are calcium dependent. Other A-esterases have greater specificity for diisopropyl fluorophosphate and some of the nerve agent phosphates and have different metal cofactor requirements. The A-esterases are largely microsomal and do not appear to have the diversity of isoforms as the oxidative enzymes. Similar to the oxidation enzymes, they occur in the highest activity levels in the liver of the mammal and at lesser activity levels in extrahepatic tissues. Phosphate/oxon hydrolysis is a detoxication reaction. The A-esterases have a relatively high affinity for some phosphates/oxons, such as chlorpyrifos-oxon and diazoxon, the active metabolites of chlorpyrifos and diazinon, respectively, but only a very low affinity for many, perhaps most, of the phosphates/oxons (Chambers et aI., 1994; Furlong et aI., 1989; Pond et aI., 1996, 1998). Therefore, the in vivo importance of the Aesterases is probably great for a few insecticides, but is difficult to estimate for many of the insecticides. The A-esterases do not appear to hydrolyze P=S compounds. Even though a relatively poor substitute, for the sake of consistency throughout this chapter, the A-esterase-mediated hydrolysis of paraoxon, the active metabolite of parathion, is illustrated in Fig. 45.9. Noncatalytic hydrolysis of the phosphates/oxons also occurs when these compounds phosphorylate serine esterases, such as carboxylesterases, butyrylcholinesterase, and even the target acetylcholinesterase; all of these esterases are classified as
B-esterases, hydrolases that are inhibited by organophosphates and that cannot catalytically hydrolyze them (Aldridge, 1953). These reactions would not be considered metabolism, because the phosphorylation is persistent, leading to a stoichiometric destruction of one phosphate/oxon molecule per serine esterase molecule with enzyme incapacitation for a long period of time. Nevertheless, the phosphorylation event releases the leaving group of the molecule, which is the same product produced in dearylation and catalytic hydrolysis reactions. Therefore, serine esterase phosphorylation contributes the leaving group to the pool of metabolite formed by catalytic reactions, and this amount is sufficiently high in in vitro preparations to be conveniently measured (Tang and Chambers, 1999). A schematic of the phosphorylation of a serine esterase by paraoxon is given in Fig. 45.10. The carboxylesterases also perform a very important catalytic hydrolysis of the carboxylic acid esters in malathion and contribute greatly to the low mammalian toxicity of malathion. Hydrolyses to the (X- and ,B-monoacids and the diacid occur from the parent malathion (Fig. 45.11). In mammals, these detoxifying hydrolyses occur more readily than the P450mediated desulfuration, allowing the malathion to be effectively detoxified prior to appreciable bioactivation and resulting in a very low acute toxicity level (rat oral LD50 of 1200 mglkg; Meister, 1990). Some representative metabolic schemes for a few important organophosphorus insecticides illustrating the major oxidations and hydrolyses are given in Figs. 45.12 and 45.13.
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CHAPTER 45
The Metabolism of Organophosphorus Insecticides
+ diethyl phosphate
Paraoxon Figure 45.9
4-nitrophenol
Hydrolysis of paraoxon by A-esterases.
Protein
I
OH serine esterase
+
+
Paraoxon Figure 45.10
Phosphorylation of serine esterases by paraoxon.
Malathion
Figure 45.11
phosphorylated esterase
Hydrolysis of the carboxylic esters of malathion by carboxylesterases.
4-nitrophenol
45.4 Hydrolysis
Diazinon
1
\
Malathion
o S
11 CH3 0 " 11 / CH2COH P-S-CH CH 0 / "COH 11
3
I
Figure 45.12
Examples of major metabolic pathways of some representative organophosphorus insecticides.
o
925
926
CHAPTER 45
The Metabolism of Organophosphorus Insecticides
Dichlorvos
o /0
11
+
HC-HC
'0
+
Tetrachlorvinphos
1
1
Figure 45.13
Examples of major metabolic pathways of some representative organophosphorus insecticides.
45.5 CONJUGATIONS The Phase 2 (conjugation) reactions can render the insecticides or metabolites even more water soluble. These conjugations frequently occur with the leaving groups produced by organophosphate hydrolysis, such as some of the phenols and heterocyclic alcohols or amines. These hydrolytic metabolites would already be detoxified products, at least with respect to anticholinesterase activity, so further metabolism would not be necessary for additional detoxication. Therefore, the Phase 2 reactions have far less impact on toxicity than the Phase 1 reactions do. However, these conjugation reactions will render the metabolites more water soluble than the parent compound or the intermediate metabolites and, therefore, allow the metabolites to be readily excreted. Sulfate and glucuronide conjugates are
possible, catalyzed by sulfotransferases and glucuronosyl transferases, respectively; both types of conjugates are hydrophilic. Glutathione conjugation is also possible. Glutathione transferases can theoretically mediate the dealkylation, primarily with methoxy compounds, of organophosphorus insecticides, such as the demethylation of methyl parathion. However, the in vivo significance of this reaction is controversial (Sultatos, 1992).
45.6 SUMMARY The organophosphorus insecticides are metabolic ally highly labile, as illustrated by the previous discussion. This metabolic lability, along with their general lack of extreme lipophilic-
References
927
ity, prevent their bioaccumulation. A variety of oxidation, reduction, hydrolysis, and conjugation reactions are possible within the group of organophosphorus insecticides. The mechanism of their acute toxicity is the inhibition of acetylcholinesterase. Some of the organophosphorus insecticides are active anticholinesterases, and any metabolism is therefore a detoxication. Many of the insecticides, however, are not active anticholinesterases in their parent form and require bioactivation in order to be effective anticholinesterases. The P450mediated desulfuration reaction is responsible for the majority of these bioactivations. Most other routes of metabolism would be detoxications. The fact that many of the insecticides or the active metabolites of those insecticides requiring bioactivation are potent anticholinesterases and others are not, as well as the fact that the efficiencies of bioactivations and of detoxications vary substantially among compounds, impart to the organophosphorus insecticides a very wide range of mammalian acute toxicity levels.
Dikshith, T. S. S. (ed.) (1991). "Toxicology of Pesticides in Animals." CRC Press, Boca Raton, PL. Ecobichon, D. J. (1996). Toxic effects of pesticides. In "Casarett and Doull's Toxicology: The Basic Science of Poisons" (c. D. Klaassen, ed.), 5th ed., pp. 643-690. Pergamon, Elmsford, NY. Forsyth, C. S., and Chambers, J. E. (1989). Activation and degradation of the phosphorothionate insecticides parathion and EPN by rat brain. Biachem. Pharmacal. 38,1597-1603.
REFERENCES
NeaI, R. A. (1980). Microsomal metabolism of thiono-sulfur compounds: Mechanisms and toxicological significance. In "Reviews in Biochemical Toxicology" (E. Hodgson, J. R. Bend, and R. M. Philpot, eds.), Vol. 2, pp. 131-172. ElsevierlNorth Holland, New York.
Aizawa, H. (1982). "Metabolic Maps of Pesticides." Academic Press, New York. Aizawa, H. (1989). "Metabolic Maps of Pesticides," Vol. 2. Academic Press, New York. Aldridge, W. N. (1953). Serum esterases: Two types of esterase (A and B) hydrolyzing p-nitrophenyl acetate, propionate and butyrate, and a method for their determination. Biachem. J. 53, 110-117. Appleton, H. T., and Nakatsngawa, T. (1972). Paraoxon deethylation in the metabolism of parathion. Pestic. Biachem. Physial. 2, 286--294. Chambers, H. W. (1992). Organophosphorus compounds: An overview. In "Organophosphates: Chemistry, Fate and Effects" (1. E. Chambers and P. E. Levi, eds.), pp. 3-18. Academic Press, San Diego. Chambers, H. W., Brown, B., and Chambers, J. E. (1990). Non-catalytic detoxication of six organophosphorus compounds by rat liver homogenates. Pestic. Biachem. Physial. 36, 308-315. Chambers, J. E., and Levi, P. E. (eds.) (1992). "Organophosphates: Chemistry, Fate and Effects." Academic Press, San Diego. Chambers, J. E., Ma, T., Boone, J. S., and Chambers, H. W. (1994). Role of detoxication pathways in acute toxicity levels of phosphorothionate insecticides in the rat. Life Sci. 54, 1357-1364.
Furlong, D. E., Richter, R. J., Seidel, S., and Motulsky, A. G. (1989). Spectrophotometric assays for the enzymatic hydrolysis of the active metabolites of chlorpyrifos and parathion by plasma paraoxonaseiarylesterase. Anal. Biachem. 180, 242-247. Levi, P. E., and Hodgson, E. (1992). Metabolism of organophosphorus compounds by the flavin-containing monooxygenase. In "Organophosphates: Chemistry, Fate and Effects" (1. E. Chambers and P. E. Levi, eds.), pp. 141154. Academic Press, San Diego. Levi, P. E., Hollingworth, R. M., and Hodgson, E. (1988). Differences in oxidative dearylation and desulfuration of fenitrothion by cytochrome P450 isozymes and in the subsequent inhibition of monooxygenase activity. Pestic. Biachem. Physial. 32, 224-231. Meister, R. T. (ed.) (1990). "Farm Chemicals Handbook 1990." Meister, WiIIoghby, OH.
Parkinson, A. (1996). Biotransformation of xenobiotics. In "Casarett and Doull's Toxicology: The Basic Science of Poisons" (c. D. Klaassen, ed.), 5th ed., pp. 113-186. Pergamon, Elmsford, NY. Pond, A. L., Chambers, H. w., Coyne, C. P., and Chambers, J. E. (1998). Purification of two rat hepatic proteins with A-esterase activity toward chlorpyrifos-oxon and paraoxon. J. Pharmacal. Exp. Ther. 286, 1404-1411. Pond, A. L., Coyne, C. P., Chambers, H. w., and Chambers, J. E. (1996). Identification and isolation of two rat serum proteins with A-esterase activity toward paraoxon and chlorpyrifos-oxon. Biachem. Pharmacal. 52, 363369. Sultatos, L. G. (1992). Role of glutathione in the mammalian detoxication of organophosphorus insecticides. In "Organophosphates: Chemistry, Fate and Effects" (1. E. Chambers and P. E. Levi, eds.), pp. 155-168. Academic Press, San Diego. Tang, J., and Chambers, J. E. (1999). Detoxication of paraoxon by rat liver homogenate and serum carboxylesterases and A-esterases. J. Biachem. Mol. Taxical. 13,261-268.
CHAPTER
46 Organophosphate Pharmacokinetics Charles Tirnchalk Pacific Northwest National Laboratory
46.1 BACKGROUND In this chapter, an overview will be presented of the pharmacokinetic principles that are of major importance in understanding the toxicology of organophosphate (OP) insecticides in animals and humans. The approach will not entail a comprehensive review of the extensive literature, but rather a focused presentation highlighting important principles by utilizing specific examples for this class of insecticide. Organophosphates constitute a large family of insecticides that are structurally related, pentavalent phosphorus acid esters. Their insecticidal as well as toxicological mode of action is primarily associated with their ability to target and inhibit the enzyme acetylcholinesterase (AChE) (Sultatos, 1994). In this regard, the acute toxic effects of OP insecticides are associated with the capacity of the parent chemical or an active metabolite to inhibit AChE enzyme activity within nerve tissue (Murphy, 1986; Sultatos, 1994). The three major classes of OP insecticides are the phosphorothionates, the phosphorodithioates, and the phosphoroamidothiolates (Chambers, 1992; Mileson et al., 1988). As an example, phosphorothionate insecticides such as chlorpyrifos, parathion, and diazinon are weak inhibitors of AChE, but once they undergo metabolic activation (desulfuration) to their corresponding oxygen analogs (oxon), they become extremely potent inhibitors. This enhanced toxicity is due to the oxon having a high affinity and potency for phosphorylating the serine hydroxyl group within the active site of AChE (Mileson et al., 1988; Sultatos, 1994). The toxic potency is dependent on the balance between a delivered dose to the target site and the rates of bioactivation and/or detoxification as illustrated in Fig. 46.1 (Calabrese, 1991). The pharmacokinetics and biochemical interactions between OPs and AChE and the toxicological implications of AChE inhibition are well understood. To further illustrate this point, a diagram relating OP toxicity with pharmacokinetic disposition and the formation of key OP metabolites is presented in Figs. 46.2 and 46.3. The thionophosphate pesticide diazinon [O,O-diethyl-O-(2isopropyl-4-methyl-6-pyrimidinyl) phosphorothioate] is being utilized for illustration purposes; however, based on a common mode of action, this scheme is readily extended to other OPs. Handbook of Pesticide Toxicology Volume 2. Agents
Organophosphate insecticides, like all chemical contaminants, can gain entry into the body and, based on the detection of low levels of OP metabolites in urine within human populations, there is good evidence for widespread although low level, exposures (Aprea et al., 1999; Hill et al., 1995). These exposures can come from numerous sources. For example, ingestion of pesticide residues on foods may account for some of the lowlevel body burdens detected, whereas accidental or intentional ingestion of OP insecticides is associated with acute poisoning, resulting in significantly higher blood, tissue, and urine concentrations of relevant OP metabolites (Drevenkar et al., 1993). Dermal exposure represents a potential exposure route during the mixing, loading, and application of OP insecticides or from skin contact with contaminated surfaces (Knaak et al., 1993). Likewise, inhalation of airborne insecticide is feasible either during an application or as the result of exposure associated with chemical drift (Vale and Scott, 1974). Once the OP arrives at a portal of entry, it is available for absorption and, based on the bioavailability of a given OP and the exposure route, a systemic dose of the parent compound (Fig. 46.3, #1) will enter the systemic circulation. Although localized portal of entry metabolism (i.e., lung, intestines, skin) is feasible, the bulk of the metabolic activation as well as detoxification reactions occur within the liver (Sultatos, 1988; Sultatos et al., 1984). As previously mentioned, phosphorothionates like diazinon do not directly inhibit AChE, but must first be metabolized to the corresponding oxygen analog (oxon; Fig. 46.2, #2) (Iverson et al., 1975; Miicke et al., 1970; Murphy, 1986; Sultatos, 1994). Activation to the oxon metabolite (#2) is mediated by cytochrome P450 mixed-function oxidases (CYP450) primarily within the liver, although extrahepatic metabolism has been reported in other tissues, including the brain (Chambers and Chambers, 1989; Guengerich, 1977). In addition, oxidative dearylation of the parent compound, forming both 2-isopropyl4-methyl-6-hydroxypyrimidine (IMHP, #3) and diethylthiophosphate (DETP, #4), represents a competing detoxification pathway that is likewise mediated by hepatic CYP450 (Ma and Chambers, 1994). These initial activation/detoxification reactions are believed to share a common phosphooxythiran intermediate and represent the critical biotransformation steps required for toxicity (Neal, 1980). Differences in the ratio
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Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
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CHAPTER 46
Organophosphate Pharmacokinetics
OP Pharmacokinetics BaJance
Figure 46.1 Parameters impacting the organophosphate (OP) insecticide toxicity balance.
of activation to detoxification are associated with chemical-, species-, gender-, and age-dependent sensitivity to OPs (see Fig. 46.1) (Ma and Chambers, 1994). Hepatic and extrahepatic (i.e., blood and tissue) A-esterase can effectively metabolize the oxon metabolite (#2), forming IMHP (#3) and diethylphosphate (DEP, #4) metabolites. Likewise, B-esterases such as carboxylesterase (CaE) and butyry1cholinesterase (BChE) that are also well distributed across tissues can metabolize the oxon; however, these B-esterases become irreversibly bound (1: 1 ratio) to the oxon and thereby become inactivated (Chanda et al., 1997; Clement, 1984). It is likewise clear from both tissue distribution and partitioning studies that phosphothionate OPs are generally well distributed in tissue throughout the body (Tomokuni et al., 1985; Wu et al., 1996). Finally, due to the extensive metabolism, little, if any, parent phosphothionate or oxon is available for excretion; however, more stable metabolites such as DEP, DETP, and IMHP are readily excreted in the urine (lverson et al., 1975; Milcke et al., 1970). Numerous pharmacokinetic approaches have been applied to OP insecticides, including: 1. Application of pharmacokinetics to understand the overall disposition and clearance of OPs 2. Development and application of pharmacokinetic models for quantitative biological monitoring to assess OP insecticide exposure in humans 3. Studies that facilitate extrapolation of dosimetry and biological response from animals to humans and the assessment of human health risk
To illustrate the utility of pharmacokinetics in addressing the health concerns associated with OP insecticides, several examples of these types of pharmacokinetic studies with OP insecticides will be used to illustrate both their utility and their limitations.
46.2 PHARMACOKINETIC PRINCIPLES OF IMPORTANCE TO ORGANOPHOSPHATE INSECTICIDES Pharmacokinetics are concerned with the quantitative integration of those processes associated with the absorption, distribution, metabolism, and excretion (ADME) of drugs and xenobiotics within the body (Renwick, 1994). Studies on the pharmacokinetics ofaxenobiotic provide critically useful insights into the toxicological response associated with a given agent. In this regard, pharmacokinetics provides quantitative data on the amount of toxicant delivered to a target site as well as species-, age-, and gender-specific and dose-dependent differences in biological response. An important application of pharmacokinetics within toxicology has been to provide a realistic estimate of risk by providing a means to quantitatively estimate the absorbed dose of a chemical under realistic exposure conditions (Clewell, 1995). Toxicology studies are designed to provide a quantitative assessment of toxicity based on what the chemical agent does to the test animals. In contrast, pharmacokinetics focuses on what the animal does to the chemical. Clearly, toxicity and pharmacokinetics are integrally related because the extent of absorption, retention, metabolic activation, or detoxification is ultimately responsible for delivering a dose to a target tissue, resulting in observed effects. Pharmacokinetics represent a critically important tool that, if used correctly, can quantitatively establish a unifying model that describes both dosimetry and biological response across exposure routes, species, and chemical agents. This approach is particularly useful for OP insecticides because they share a common mode of action through their capability to inhibit AChE activity (Mileson et al., 1988). Pharmacokinetic strategies for quantitating dosimetry can be developed to measure the parent compound and its active (i.e., oxon) or inactive metabolites. It is also feasible to link dosimetry with biologically based pharmacodynamic (PD) response models based on a common mode of action (i.e., AChE inhibition). In general, pharmacokinetic modeling approaches can be characterized as empirical or physiologically based and both types of models have been applied to understand the toxicological response to OP chemicals in multiple species (Brimer et al., 1994; Gearhart et al., 1990; Pena-Egido et al. , 1988; Sultatos, 1990; Tomokuni et al., 1985; Wu et al., 1996).
46.2.1 COMPARTMENTAL PHARMACOKINETIC MODELS Compartmental models have formed the cornerstone of pharmacokinetic analysis and as such have been extensively utilized to assess bioavailability, tissue burden, and elimination kinetics in various species, including humans. All pharmacokinetics are concerned with the time course by which a chemical is absorbed into the systemic circulation, distributed throughout the body, altered through metabolic transformation, and eliminated. Compartmental models are empirical and as such
46.2 Pharmacokinetic Principles of Importance to Organophosphate Insecticides
} El Parent Compound Oxon @ Metabolite
Diazinon-oxon
Diethylphosphate (DEP) 2.isopropyl.4.methyl.6.
@
hydroxypyrimidine
Major Compound Specific Metabolites
(IMHP)
t
Non-Specific Metabolite
CYP450
C;H,OH
H~N')-CH' ....
NCH
3
CH,
Figure 46.2
Metabolic scheme for the metabolism of the organophosphate (OP) insecticide diazinon.
Stomach
Liver
CD
CD
~2 3 4
o CYP450~ - Activation -Detoxification -A-EST (detoxification) -CaE (detoxification)
L-Lun~ g ----'
'---------l~ Systemic Circulation ~ oA-EST (detoxification) ~
Other Tissuesr1"\l o
Fat
\..Y
- Muscle
J /
.t ? /
-r--- -- - - ,
f3YA'4
, - - - - - --.t
Skin
oCaE (detoxification) oBlood AChE inhibition \2;\j:)
CD Target Tissoe ACbE Inhibition
Figure 46.3 Compartmental flow diagram illustrating the critical tissue compartments associated with absorption, distribution, metabolism and excretion of organophosphate (OP) insecticides. The circled numbers (1-4) correspond to the parent compound and major metabolic products associated with metabolism of diazinon (see Fig. 46.2) that are most likely found within each compartment.
931
932
CHAPTER 46
Organophosphate Phannacokinetics
Oral Dose (mglkg)
ka (hr-l)
- - - - - 1...~Cb (fJg/mL) Ill"'"
Cb (J.lglmL) = Ka x dose x F Vd x (ka-ke) Urinary Excretion Rate
Absorbed Dose (l!g)
=
Urine
Volume Distribution Vd (mL)
x
exp (ke x t-lme - ka x t-lme)
(J.lglhr)
= Cb x ke
x
(1)
Vd x Body wt_ (2)
Figure 46.4 Single compartment model used to describe the blood and urine time-course of 3,5,6trichloropyridinol (TCP) a major metabolite of the organophosphate (OP) insecticide chlorpyrifos (CPF). Equations adapted from Nolan et al. (1984).
consider the organism as a single- or multicompartment homogeneous system. The number and behavior of the compartments are primarily determined by the equations chosen to describe the time-course data and not the physiological characteristics of the organism (Krishnan and Andersen, 1994). In these models, the net transfer between compartments is directly proportional to the difference in chemical concentration between compartments. However, the rate constants associated with the transfer between compartments cannot be experimentally determined (Srinivasan et aI., 1994). Compartmental models range from simple well-mixed single-compartment models to more complicated multicompartment models that are used to describe the blood and/or plasma time course of a chemical or drug. These simple compartmental approaches have been broadly utilized to model the pharmacokinetics of OP insecticides and their major metabolites (Braeckman et aI., 1983; Drevenkar et aI., 1993; Nolan et aI., 1984; Wu et aI., 1996). For example, Nolan et al. (1984) developed a one-compartment pharmacokinetic model that accurately describes the blood and urine time course of 3,5,6trichloropyridinol (TCP), a major metabolite of the OP insecticide chlorpyrifos, in human volunteers. A diagram of this single-compartment model is illustrated in Fig. 46.4. In this model, the blood TCP concentration and urinary excretion data were simultaneously fit to a single-compartment model using the equations shown in the figure. Absorption (k a ) and elimination (k e ) are handled as first-order processes, the blood TCP concentration is represented by Cb, and F and Vd represent the fractional absorption and the volume of distribution, respectively. To develop this model, male volunteers were orally administered a dose of 0.5 mg chlorpyrifoslkg of body weight. Then blood and urine specimens were collected at specified in-
1000
100
~.....-urine Blood
0.1
0.01 1 - - - - r - - - , - - - - - , - - - , - - - - r - - - - - - ,
20
40
60
80
100
120
Time (hrs)
Figure 46.5 Time-course of 3,5,6-trichloropyridinol (TCP) in the blood and urine of male volunteers orally administered a 0.5 mg chlorpyrifos (CPF)/kg of body weight dose. Figure adapted from Nolan et al. (1984).
tervals and analyzed for TCP. The model parameters used to describe the time course of TCP and the model fit of the experimental data are presented in Table 46.1 and Fig. 46.5. The model provides an excellent fit of the experimental data and, based on the model parameters, it was determined that approximately 72% of the ingested dose was absorbed and eliminated in the urine with a half-life of 27 h. Based on this model, Nolan et al. (1984) suggested that the blood TCP concentration and/or urinary excretion rate could be utilized to quantify the amount of chlorpyrifos absorbed under actual use conditions. Although compartment modeling is extremely useful for interpolation within the confines of the test species and experimental conditions (i.e., exposure routes and dose levels), these
46.2 Pharmacokinetic Principles of Importance to Organophosphate Insecticides
933
Table 46.1 Selected Model Parameters Describing Blood Concentrations and Urinary Excretion of 3.5,6-Trichloropyridinol by Individual Volunteers Following Oral Administration of the Organophosphate Insecticide Chlorpyrifos Model
Percentage
Absorption
Absorption
Absorption
Volume
Elimination
Elimination
predicted
dose
Body weight
lag time
half-life
distribution
percentage dose
recovered
(h)
(h)
Vd (mllkg)
rate constant (h- I )
half-life
(kg)
rate constant ka (h- I )
ke (h)
absorbed
in urine
Range
72-102
0.9-1.9
0.1-2.7
0.4-6.9
160-204
0.02-0.03
21-32
52-84
49-81
Mean± SD
83.3 ± 10.3
1.3 ± 0.4
1.5 ± 1.2
0.5
181 ± 18
0.026 ± 0.005
26.9
72± 11
70± 11
Parameter
Data obtained from six male volunteers. Data adapted from Nolan et al. (1984).
models are limited in their capability to extrapolate across dose levels, species, and exposure routes (Krishnan and Andersen, 1994). To enable extrapolation, physiologically based pharmacokinetic (PBPK) models have emerged as an important tool that has seen broad applications in toxicology and more specifically in human health risk assessment (Andersen, 1995; Clewell and Andersen, 1996; Krishnan and Andersen, 1994; Leung and Paustenbach, 1995; Mason and Wilson, 1999). 46.2.2 PHYSIOLOGICALLY BASED PHARMACOKINETIC MODELS
Unlike compartment modeling approaches, PBPK models utilize biologically meaningful compartments that represent individual organs such as liver and kidney or groups of organ systems (i.e., well perfused/poorly perfused) (Mason and Wilson, 1999). The general model structure is based on an understanding of comparative physiology and xenobiotic metabolism, a chemical's physical properties that define tissue partitioning, the rates of biochemical reactions determined from both in vivo and in vitro experimentation, and the physiological characteristics of the species of interest (Krishnan and Andersen, 1994). PBPK models have been developed to describe target tissue dosimetry for a broad range of environmental contaminants such as solvents, heavy metals, and pesticides, including OP insecticides (Andersen et aI., 1987a; Corley et aI., 1990; Gearhart et aI., 1990; O'Flaherty, 1995; Sultatos, 1990). A number of reviews have been published on the development, validation, application, and limitations of PBPK models in human health risk assessment (Andersen, 1995; Clewell, 1995; Clewell and Andersen, 1996; Frederick, 1995; Krishnan and Andersen, 1994; Leung and Paustenbach, 1995; Mason and Wilson, 1999; Slob et aI., 1997). To illustrate the application of this modeling approach to OP insecticides, a PBPK model that also incorporates a pharmacodynamic (PD) component to describe AChE inhibition following diisopropylfluorophosphate exposure in rodents will be described (Gearhart et aI., 1990). Gearhart et al. (1990) developed a basic PBPKlPD model structure that described target tissue dosimetry and AChE inhibition following an acute exposure to diisopropylfluorophosphate in mice and rats. In developing this model, the authors were primarily interested in
building a structure that could readily be extended to describe the acute effects for a broad range of commercially important OP insecticides. A diagram of the PBPKlPD model for diisopropylfluorophosphate in rats is illustrated in Figs. 46.6 and 46.7. The conceptual representation of the PBPK model for diisopropyl fluorophosphate is based on the anatomical and physiological characteristics of the rat and the major determinants of diisopropyl fluorophosphate disposition, which include esterase binding and hydrolysis, tissue partitioning, and diisopropylfluorophosphate volatility (Gearhart et aI., 1990; Krishnan and Andersen, 1994). Because this OP ester does not require metabolic activation, like thionophosphate OPs, the hydrolysis of diisopropylfluorophosphate by blood and tissue Aesterase is a major factor in determining the protection against AChE inhibition. Diisopropylfluorophosphate binds to and inhibits B-esterases, including AChE, BChE, and CaE. Although binding to AChE is associated with acute neurotoxicity, the binding to BChE and CaE is without adverse physiological effect and as such represents a detoxification pathway (Clement, 1984; Fonnum et aI., 1985; Pond et aI., 1995). The PBPKlPD model compartments included those tissues associated with toxicological response (i.e., brain, lung, and diaphragm), those containing high A-esterase activity (i.e., liver, kidney, and blood), a fat compartment having the highest tissuelblood partitioning, and the remaining tissues being collectively lumped (Gearhart et aI., 1990). To develop this model, tissue partitioning coefficients (PCs) were determined by the vial equilibration technique (Gargas et aI., 1989; Sato and Nakajima, 1979). In general, the tissue : blood PCs ranged from 0.77 to 1.63; however, the fat : blood partitioning was the highest with a coefficient of 17.6. The generalized mass balance differential equation for calculating diisopropylfluorophosphate tissue concentration and AChE tissue inhibition are also presented in Figs. 46.6 and 46.7. Within each tissue compartment, the net concentration of diisopropylfluorophosphate (mg/l) is a function of blood flow to the tissue, chemical partitioning from the blood into the tissue, and the loss of diisopropylfluorophosphate due to hydrolysis by A-esterase and inhibition of B-esterases (AChE, BChE, and CaE).
934
CHAPTER 46
Organophosphate Pharmacokinetics
iv & s.c. dose
~r.
a,.J
Mass Balance Differential Equations for DFP Dosimetry
~+ Lungs
"'1 ,
..... ....
r
.... ....
r
L L
1
Qc ,---.,
...,..
J
.... Qbr
Brain
J ....
Liver
1.... QJ J ....
0 0
;:t5
..... ....
I L
.... ....
r,.
en :;
0
5 >
r
.....
1
.....
J
.....
J'"
,
.1 .... Qr LRapId PertusedJ....
.....
,
Kidney
1..... Qk
1.... Qf
Fat
r
.....
1 Diaphragm
(DFP hydrolysis by A-EST) (4) (5)
- KCDE • CCE ... C
(OI<' P inhibition of Ca E)
(6)
- KBChE • CUE'" C
(OFI' inhibition of RChE)
(7)
~
Q
~
-5
CA Cv Vrnnx Km KACbE CAE KDoE CcE KnChE CBE
~
..... Qs
1.... Qd J.....
*C
(OFP inhibition of AChE)
where, V = C =
L Slow PcrfusedJ .....
.....
- (Vm..-C.)lKm + Cv)
0 0
J.....
.
(OFI' tissue concentration) (3)
- KACbE - CAE
"0
"'0
V • dC/dt = Q • (CA - Cv)
= = =
= =
= = = =
= =
Volumc of tissuc (L) DFP conccntration in tissue (mgIL) Blood flow to tissue (Llh r) DFP concentration in arterial blood entering tissue (mglL) OFP concentration in venous blood Icaving tissue (mglL) Maximum rate of A-EST hydrolysis of DFP (mglhr) Michaelis constant for A-EST in tissue (mglL) Bimolecular rate constant for DFP reaction with AChE (J.lM hrrl AChE tissue concentration (J.lM) Bimolecular rate constant for OFI' reaction with CaE(jlM hl'r' CaE tissue concentration (J.1M) Bimolecular rate constant for OFP reaction with BChE(jlM hrr' BChE tissue concentration (jlM)
Figure 46.6 Physiologically based phannacokinetic (PBPK) model structure and O1as balance differential equation describing the distribution of diisopropylfluoropho phate (DFP) in the rat. Figure adapted from Gearhart et al. (1990),
Synthesis of new AChE
....
DFP ...
,..
,
Free AChE
....
,
Bimolecular rale of AChE Inhibition
...,..
..... Regeneration of Bound AChE
Rate of
Inhibited AChE
AChE Aging ...
,..
"Aged" AChE
Differential Equations for AChE Inhibiti on
Basal degradation of AChE
V • dAl':fdt = (KAChE •
where V C KAChE CAE
KRA AE
KM
CA~~
• C)
(inhibition of AChE)
(8)
- (KM ' AE)
(Regeneration of AChE)
(9)
-(KM - AE)
(Aging of AChE)
(10)
=
Volume of tis ue (L) DFP concentration in tissue (mgIL) Bimolecular rate constant for DFP reaction with AChE Free AChE tissue concentration (J.lM) Rate of regeneration of inhibited AChE (hr'l) Inhibited AChE (J.lM) Rate of aging of inhibited AChE (hr")
Figure 46.7 Phannacodynamic (PD) model structure and mass balance differential equations describing the inhibition of acetylcholinesterase (AChE) by diisopropylfIuorophosphate (DFP) in the rat. Figure adapted from Gearhart et al. (1990).
(J.lM hrrl
46.2 Phannacokinetic Principles of Importance to Organophosphate Insecticides
Gearhart et at. (1990) calculated basal AChE activity (~mol) based on a zero-order enzyme synthesis rate (~mol/h) and a first-order rate of enzyme degradation (h- I ). A balance between the bimolecular rate of inhibition and the rate of AChE regeneration and aging determined the amount of free AChE. Similar equations were utilized to quantify the impact of diisopropylftuorophosphate on tissue CaE and BChE activity. The capability of the diisopropylftuorophosphate model to simulate both diisopropylfluorophosphate tissue dosimetry and AChE inhibition is illustrated in Figs. 46.8 and 46.9 in mice that were administered a single intravenous (iv) dose of 1 mg diisopropylftuorophosphate/kg of body weight. The model does a reasonably good job of describing brain tissue dosimetry and AChE inhibition. In brain, the diisopropylftuorophosphate concentration rapidly falls to a fraction of its peak concentration within about 1 min , whereas AChE was rapidly inhibited to 20% of control activity. In both cases, the model 1200
:;
300
000
:;
J..
Cl E
-...
(1J
(1J
(1J
a..
.5 a.. u..
400 I·
0
200 000 000
10
L
025
-
E
I
200 150
075
100
125
00 050 000 000
j
050
1 50
'0°
:;
10°
---C'l
(1J
,.
E (fj (1J
Q: C
g .(6
--
CXl
0- 2 ;:
.5 n. u..
n. u..
0
0
o-l
'I
0
5
0
15
20
25
, 30
1.00
1.50
200
~
'0-
c
...
0.50
r
(c)
I
:; -.... C'l
250
CXl
u.. 0
• (b )
c
E (fj Q: c
simulations were consistent with the experimentally derived data. The development and application of PBPK modeling for human health risk assessment are not without their challenges and limitations. Before a model can be used to assess risk, a determination must be made concerning the model's capability to accurately predict dosimetry and biological response (Frederick, 1995). Furthermore, PBPKlPD models are data intensive, so to adequately develop and validate a model generally requires extensive experimentation to support model parameterization and validation (Clewell, 1995). Nonetheless, a consensus opinion of an expert scientist panel concluded that biologically based risk assessments that include well-validated PBPKlPD models can provide the most accurate quantitative assessment of human health risk from exposure to environmental chemicals (Frederick, 1995).
(a)
Cl E
935
250
300
0
'0 - ~ 3
0
0
5
0
15
20
Time (minutes) Figure 46.8 Time-course of free diisopropyl fluorophosphate (DFP) concentration (mglL) in plasma and brain in male mice after tail vein injection of I mg DFPlkg. Each datum represents the mean ± SD of 5 animals. Solid line depicts PBPKlPD model simulation. Figure used with permission from Gearhart et at. (1990).
30
936
CHAPTER 46
Organophosphate Pharmacokinetics 120
-
'0
100
~
~
C
c: 0
U
0
'0
c:
t)
c:
t) (\I
~
~
LL
I
I
0.40
w
or.
U
c:
·co
060
LL
or.
<
0.80
.2
060
(\I
W
1.00
U
0.80
'0 .2
(b)
'0
(8 )
040
}
U
<
020
c:
·co...
~
ID
020
ID
0.00 000
500
10 00
I
I
15.00
20.00
2500 3000
000 000
I
400
Time (minutes)
'0
800
12.00
16.00 2000
2400
Time (hours)
125 (c l
~
C 0
u '0
00
.2
075
c:
t)
...
(\I
LL
I
050
W
or.
U
<
.=
025
(\I
ID
0.00
0
40
80
120
160
200
Time (hours) Figure 46.9 (a-c) Time-course of brain acetylcholinesterase (AChE) activity in male mice after tail vein injection of I mg diisopropylfluorophosphate/kg. Each datum represents the mean ± SD of 5 animals. Solid line depicts PBPKJPD model simulation. Figure used with permission from Gearhart et at. (1990).
46.3 PHARMACOKINETIC APPROACHES APPLIED TO ORGANOPHOSPHATE INSECTICIDES 46.3.1 APPLICATION OF
PHARMACOKINETICS TO UNDERSTAND THE OVERALL DISPOSITION AND CLEARANCE OF ORGANOPHOSPHATE INSECTICIDES Pharmacokinetic studies conducted in multiple species, at various dose levels, and across different routes of exposure can provide important insight into the in vivo behavior of a chemical agent and how it contributes to the observed toxicological response in a given species. To illustrate this point, a comparison is made of selected pharmacokinetic parameters obtained from a diverse group of studies conducted in animals exposed to either parathion or diazinon. As noted in Tables 46.2 and 46.3, no single study provides all the pertinent information; yet,
collectively, they provide a consistent qualitative picture of the overall pharmacokinetics of these OP insecticides. The bioavailability of OP insecticides, defined as the amount of systemically available dose, is a function of the extent of absorption and first-pass metabolism. Braeckrnan et al. (1983) conducted a pharmacokinetic study in the dog following both oral and iv administration of parathion. Comparisons of plasma parathion areas under the curve (AV Cs) indicated that 1-29% of the orally administered parathion was bioavailable. The authors suggest that the low systemic oral bioavailability of parathion is primarily associated with a rapid hepatic first-pass metabolism based on the high hepatic extraction (82-97%) that was determined after iv administration. Wu et al. (1996) conducted similar bioavailability studies in the rat with diazinon. The blood time course of diazinon in the rat following iv and oral doses of 10 and 80 mg/kg, respectively, is presented in Fig. 46.10. The results suggest that, following oral administration, absorption is rapid (absorption tl/2 = 2.6 h) with peak plasma concentrations of diazinon being attained within 2 h postdosing; yet a comparison of AVCs, when corrected for administered dose, indicate
46.3 Pharmacokinetic Approaches Applied to Organophosphate Insecticides
937
Table 46.2 Selected Model Parameters Describing Blood Concentration Pharmacokinetics of Parent Compounds in Various Species Following Exposure to the Organophosphate Insecticides Parathion and Diazinon Elimination kinetics Absorptionlbioavailability kinetics
Species Rabbita
Two-compartment model
Distribution kinetics
Absorption
Hepatic
Volume
Protein
distribution
binding
tl/2 a
percentage
tl/2 (h)
extraction percentage
Vd ss (lJkg)
percentage
(h)
100
N/A
14.24 ± 6.34
0.021 ±0.04
7.58 ± 6.45
100
N/A
2.6±0.9
100
N/A
9.76 ± 5.65
Dose
Bioavailability
(route) 1.5 mglkg
tl/2f3 (h)
Elimination
Clearance
ke t l/2 (h)
Cl (l/h/kg)
5.08 ± 3.06
3.99 ± 1.13
1.08 ± 0.27
2.54 ± 1.67 6.59 ± 3.36
(iv) Rabbit"
68 b
3 mg/kg
0.13 ± 0.29
(oral) PigletC
0.5 mg/kg
97 ± 1
3.0 ± 1.5
(iv) Pigd
1 mg/kg
3.6 ± 1.08
4.42 ± 1.20
-
4.69 ±0.8
(iv) Pigd
50mg/kg
9.93 ± 5.28
(dermal)
Doge
5 mg/kg
N/A
82-97
99
(iv)
Doge
10 mglkg
1-29%
99
(oral) Rat!
5-10 mg/kg
100
N/A
48-55
20.01 ± 11.27
89.1
22.93 ±4.82
89.1
0.33 ± 0.10
4.70 ± 1.84
(iv) Rat!
35.5
80 mglkg
2.55
2.86±0.58 4.60 ± 1.05
(oral) apena-Egido et al. (1988). bEstimated by comparing oral and iv AUC after adjusting for dose. cNielsen et al. (1991). dBrimer et al. (1994). fWuet al. (1996). Table 46.3 Tissue Concentration, Tissue Plasma Ratio and Partition Coefficients Following Exposure to the Organophosphate Insecticides Parathion and Diazinon
Diazinonc
l4C-Parathiona
Parathionb
Paraoxonb
0.5 mg/kg; iv;
partition coefficient (PC)
partition coefficient (PC)
Tissuelblood
Tissuelblood
nglg
piglet 3 h postdosing
10 mglkg; iv;
20 mg/kg; iv;
rat 4 h postdosing
mouse 5 h postdosing
Tissue/plasma Tissue Blood/plasma
ng/g
ratio
Diazinond
Tissue/plasma
262 ± 145
Tissue/plasma
ratio
nglg
~130
35
Liver
1254 ± 638
4.78
5.21
6.62
325 ± 25
2.50
120
Kidney
1360 ± 546
5.19
5.21
790 ± 60
6.08
3000
280 ± 10
2.15
160
Lung
421 ± 92
1.60
5.21'
6.62 6.62e
Muscle
484± 92
1.85
2.55!
3.62!
Heart
302 ± 85
1.15
215 ± 76
0.82
Fat Brain
101.2
aNielsen et al. (1991). b Gearhart et al. (1994). cWu et al. (1996). dTomokuui et al. (1985). eWell-perfused tissue. f Poorly perfused tissue.
4.56
ratio
3.42 85.7
10.22 2.31
4.57
938
CHAPTER 46
Organophosphate Pharmacokinetics
10
.,
~ E
§
Peak Plasma Concentration (2 hr)
/
..1<~«............-....... •
-
ir..
•
.; ~'" ,'* '.~... •..-- ~~a~ ~~4~~~~1 rn. I,,' L'
C
~
(J
c
,
0
(J
Z
N 0.1 Q
'.
--- --+--
~
-----.____
E 'Jl
.. _ _ _iv 10mg/kg - - - _+ ~ AUC 2.27 ± 0.47 mg hr-I Vl
s:;= 0.01~-_-_-_--_-__- _ -_ _- _ - _
o
2
3
4
5
6
7
8
9
Time (hrs) Figure 46.10 Plasma time-course of diazinon (DZN) in rats following intravenous (iv) and oral administration of 10 and 80 mg DZN/kg of body weight, respectively. Data extracted from Wu et al. (1996).
that only about 35% of the oral dose was systemically bioavailable. The hepatic extraction ratio for diazinon ranged from 48 to 55% and was qualitatively consistent with the findings of Braeckman et al. (1983) for parathion in the dog as well as chlorpyrifos in mice (hepatic extraction ratio ~46%) (Sultatos, 1988). Rapid oral absorption (t1/2 = 0.02 h) and lower oral bioavailability (~68%) were also demonstrated in a study in which rabbits were administered iv and oral doses of parathion (Pena-Egido et aI., 1988). Likewise, in vivo animal models also suggest that dermal absorption and systemic bioavailability of OP insecticides will be quite low (Brimer et aI., 1994). Once these OP compounds have been absorbed, systemic distribution throughout the body tissues is rapid (Vale, 1998). For example, a high volume of distribution was observed ranging from 3 to 14 and from 20 to 23 l/kg in several different species administered parathion or diazinon, respectively (see Table 46.2). A cross-species comparison of the tissue distribution data following parathion or diazinon exposure is consistent with the large volume of distribution, suggesting that the OP tissue concentration follows the order of kidney > liver > lung/muscle/heart> brain (see Table 46.3). Phosphorothioates such as diazinon and parathion are more lipophilic than their respective oxon metabolites and therefore can be sequestered in the fat compartment, which may account for prolonged intoxication and observed clinical relapses (Vale, 1998). Gearhart et al. (1994) determined the PCs for both parathion and the toxic metabolite paraoxon (see Table 46.3). In general, the PCs for parathion and paraoxon are comparable; however, parathion has an order of magnitude (101 vs. 10.11) greater affinity than paraoxon for fat. The systemic distribution, elimination kinetics, metabolic transformation, and target site availability of a drug or chemical
are often dependent on the extent of reversible plasma/serum protein binding (Renwick, 1994). For example, as shown in Table 46.2, parathion and diazinon are extensively bound to plasma protein (ranging from 89 to 99%) and the extent of binding is concentration independent. This response is likewise consistent with the findings of Sultatos et al. (1984), who reported that chlorpyrifos is approximately 97% bound to mouse plasma proteins over a broad concentration range. This high degree of protein binding in conjunction with the high volume of distribution also suggests that tissue binding may in fact be more important than plasma binding in determining the overall disposition and clearance of OPs (Braeckman et aI., 1983). Although the OP insecticides parathion and diazinon are well distributed throughout the body and extensively bind to both plasma and tissue proteins, they are both rapidly cleared from the body primarily in the urine as degradation metabolites of the parent OPs [i.e., p-nitrophenol, 2-isopropyl-4-methyl-6hydroxyprimidine (lMHP)] (Iverson et aI., 1975; Mticke et aI., 1970; Nielsen et aI., 1991; Vale, 1998). The overall systemic clearances for both parathion and diazinon are quite fast and comparable, ranging from 4 to 6.6 l/h/kg, and are consistent with the rapid blood/plasma terminal phase half-life (2.5-5 h; see Table 46.2). As previously indicated, comparative species pharmacokinetic analysis is useful for understanding the in vivo behavior of OP insecticides. Although generalization to all OP agents is unwise, these types of comparative analyses do provide important insights. In summary, the oral absorption of both parathion and diazinon is rapid with peak plasma concentrations being obtaincd within a few hours of exposure. However, oral bioavailability is low and appears to be at least partially associated with a high rate of hepatic first-pass metabolism. Although these OP
46.3 Pharmacokinetic Approaches Applied to Organophosphate Insecticides insecticides are extensively bound to plasma proteins, they are equally well distributed throughout the body's tissues and the parent phosphorothioates can sequester within the fat compartment. Nonetheless, the overall clearance is quite fast and is most likely associated with the rapid metabolism and elimination of the OP metabolites. 46.3.2 DEVELOPMENT OF PHARMACOKINETIC MODELS FOR QUANTITATIVE BIOLOGICAL MONITORING TO ASSESS ORGANOPHOSPHATE INSECTICIDE EXPOSURE IN HUMANS
In assessing human exposure to chemical agents, biological monitoring (biomonitoring) is an important quantitative measure of the amount of chemical agent that is systemically absorbed. This approach entails the quantitation of the chemical or its metabolites in biological fluids (i.e., blood, urine, exhaled breath) and offers the best means of accurately assessing exposure because it measures actual, rather than potential, exposure (Woollen, 1993). However, to accurately predict human dosimetry from occupational and/or environmental exposure to xenobiotics, human volunteer pharmacokinetic studies conducted under controlled conditions are of vital importance (Wilks and Woollen, 1994; Woollen, 1993). Both occupational and environmental exposures to OPs are primarily associated with dermal exposure; accounting for more than 90% of the absorbed dose (Aprea et aI., 1994). Therefore, an understanding of the percutaneous absorption of OPs is critical for quantitatively determining a systemic dose. The extent of dermal bioavailability for a number of 14C-Iabeled OP insecticides has been determined in humans utilizing both in vivo studies in volunteers and in vitro dermal penetration with skin obtained from cadavers (Wester et aI., 1983, 1992, 1993). A summary of the percentage absorption following in vivo and in vitro dermal exposure to the OP insecticides diazinon, isofenphos, and malathion is illustrated in Fig. 46.11. The general experimental design of these studies entailed three major components. First, human volunteers were administered a topical dose of a known concentration of 14C-Iabeled OP for a specified exposure period. The extent of absorption was determined by quantitating the amount of 14C excreted in the urine and remaining on the skin surface. Second, in vitro percutaneous absorption was determined using a glass flow-through penetration cell in which the percentage absorption through human cadaver skin was determined by the amount of radiotracer that transferred into the receptor fluid. Finally, to calculate the in vivo percentage absorption, rhesus monkeys were given a 14C-Iabeled OP as an iv dose. The percentage dose absorbed in humans was calculated from the ratio of 14C excreted in the urine after topical (humans) and iv (monkey) dosing. The in vivo absorption for the three OP insecticides diazinon, isofenphos, and malathion in human volunteers following a topical application is very low, ranging from 2.5 to 3.9% of the applied dose. The percentage absorption as determined in vitro
939
was likewise comparable for isofenphos (3.64% ± 0.48), but slightly higher and considerably more variable for diazinon (14.1% ± 9.2). Percutaneous absorption studies conducted in humans are of particular importance because it is known that dermal absorption in animals, such as the rat, is often greater than in humans (Wester and Maibach, 1983). For example, Knaak et al. (1990) conducted a dermal absorption study in rats with isofenphos and reported that 47% of the applied dose was absorbed, which is 12-fold higher than the results seen in human volunteers. The major limitation associated with the experimental design of Wester et al. (1983, 1992, 1993) is that the quantitation of only 14C provides no information on the specific form of the compound (i.e., parent or metabolite) that is systemically available. Nonetheless, these studies provide important quantitative information on the extent of dermal absorption. To better understand the systemic pharmacokinetics of OPs and to develop pharmacokinetic models that can be utilized for biomonitoring, controlled human studies that quantitate the time course of parent chemical or metabolites in blood and urine are key. Nolan et al. (1984) conducted a controlled human pharmacokinetic study to follow the fate of a major metabolite, 3,5,6-trichloropyridinol (TCP), which is excreted in the urine following both oral and dermal administration of chlorpyrifos. Griffin et al. (1999) also conducted a controlled human study with chlorpyrifos in human volunteers, but quantitated the urinary excretion of the dialkylphosphate metabolite. A selection of comparative pharmacokinetic parameters from the controlled human chlorpyrifos studies is presented in Table 46.4. Overall, the pharmacokinetic results obtained using TCP or dialkylphosphate in human volunteers are entirely consistent with each other. For example, following oral administration, chlorpyrifos is rapidly absorbed with maximum plasma concentration and excretion being obtained by 6 and 7 h postdosing for TCP and dialkylphosphate, respectively. The extent of absorption was quite good, based on the amount of metabolite (70-93%) recovered in the urine. In comparison, the dermal absorption was consistently slower with peak concentrations of metabolite being achieved by 17-24 h postdosing for both studies. Also, the amount recovered based on TCP and dialkylphosphate metabolites in the urine was 1.35 and 1%, respectively, suggesting limited dermal absorption of chlorpyrifos. Nolan et al. (1984) reported an elimination half-life of 26.9 h following oral administration, whereas Griffin et al. (1999) reported half-lives of 15.5 and 30 h for dialkylphosphate following oral and dermal exposure to chlorpyrifos, respectively. The increase in the urinary elimination half-life following dermal exposure is most likely associated with a delay in chlorpyrifos absorption through the skin. However, differences in the rates of TCP and dialkylphosphate kinetics are also a possible explanation (Griffin et al., 1999). Nonetheless, the elimination half-life for chlorpyrifos based on either TCP or dialkylphosphate clearance is consistent. These types of pharmacokinetic data are being used to develop models to biomonitor for OP exposure. Nolan et al. (1984) developed a one-compartment pharmacokinetic model having the same volume of distribution and elimination rate
940
CHAPTER 46
Organophosphate Pharmacokinetics
Figure 46.11 Summary of human dermal penetration (in vivolin vitro) for the organophosphate (OP) insecticides diazinon, isofenphos and malathion.
Table 46.4 Comparison of Oral and Dermal Pharmacokinetic Parameters Describing the Blood Concentration and Urinary Excretion of 3,5,6-Trichloropyridinol (TCP) and Dialkylphosphate (DAP) by Volunteers Following Exposure to the Organophosphate Insecticide Chlorpyrifos Exposure
Absorption
Absorption
Absorption
Elimination rate
Elimination
Model predicted
route/
Dose
rate
half-life
percentage dose
recovered in
(mg/kg)
(ng/cm2/h)
(h)
constant ke (h- I )
half-life
metabolite
rate constant ka (h- I )
(h)
absorbed
urine
1.5 ± 1.2
0.5
0.0258 ± 0.0051
26.9
72± 11
Percentage dose
Oral Tcpa
0.5
DApb
0.014c
Dermal Tcpa
5
DApb
0.41
15.5 0.0308 ± 0.01 456
aNolan et al. (1984). bGriffin et al. (1999). CEstimated based on average body weight (71 kg).
22.5
1.35 ± 1.02 30
70 ± 11 93 (range 55-115) 1.28 ±0.83 1.00
46.3 Pharmacokinetic Approaches Applied to Organophosphate Insecticides constant to describe blood and urinary TCP kinetics following oral and dermal exposure to chlorpyrifos (see Fig. 46.4). Similarly, the quantitative measurement of urinary dialkylphosphate is increasingly being used for biomonitoring for OP exposures (Gargas et al., 1989). The development of pharmacokinetic models that are capable of describing the uptake, distribution, and elimination of OP insecticides based on the quantitation of major degradation metabolites represents an extremely useful and simple approach for exposure biomonitoring. 46.3.3 THE APPLICATION OF PHARMACOKINETICS FOR QUANTIFYING EXPOSURE TO ORGANOPHOSPHATE INSECTICIDES
The ability to more accurately quantitate human exposure to OP insecticides has been enhanced by the use of biomonitoring approaches linked to pharmacokinetic analysis. This has successfully been used to estimate agricultural worker exposures during and after the application of insecticides, to evaluate secondary exposures within cross-sectional epidemiology studies, and to assess dosimetry in persons who have been acutely poisoned either accidentally or through intentional self-administration (Drevenkar et aI., 1993; Lavy et aI., 1993; Loewenherz et aI., 1997). Historically, workplace exposure to chemicals has been controlled through environmental monitoring that has primarily focused on the measurement of the chemical contaminant in the ambient air. However, because airborne concentrations may not be linearly correlated with absorption, this approach does not provide an accurate assessment of the internal dose (Franklin et al., 1986). In agricultural settings, worker exposure studies have incorporated personal external monitoring to estimate the amount of chemical available from inhalation (i.e., breathing-zone sampling pumps) and dermal absorption (i.e., patch method and hand washes). Where feasible, these studies have also incorporated biomonitoring (i.e., urinary metabolites) to quantitate the amount of absorbed dose (Chester, 1993; Franklin et aI., 1981, 1986). Franklin et al. (1981, 1986) estimated exposure of workers to the OP insecticide azinphos-methyl (guthion) utilizing both external personal monitoring and urinary biomonitoring of alkylphosphate metabolite. When patch data were utilized to calculate exposure and plotted against total urinary metabolite excretion, no correlation was observed (Franklin et aI., 1981). However, the authors did report a much better correlation when the amount of alkylphosphate metabolite excreted in the urine was compared against the amount of active ingredient sprayed. This relationship is illustrated in Fig. 46.12, where the amount of alkylphosphate metabolite excreted in the urine increases with increasing amounts of active ingredient. Because agricultural workers routinely apply numerous pesticides and are often sequentially exposed to OP insecticides within a relatively short time span, a number of exposure studies have been conducted to evaluate mixed OP insecticide exposures. Hayes et al. (1980) evaluated the occupational exposure
941
of pest control operators in which the bulk of the pesticide applications (~ 80%) involved the combined use of the OP insecticides vaponite, diazinon, and chlorpyrifos. Worker biomonitoring was based on blood AChE determination and the quantitation of dimethyl- and diethylphosphate and dimethyland diethylphosphothioate metabolites in the urine. The authors reported that external air monitoring did provide information regarding the levels and types of OP exposures, but did not provide adequate information on the degree to which these OPs were absorbed. The urinary alkylphosphate levels provided sensitive quantitative information on absorption and excretion of these pesticides. However, because the alkylphosphate metabolites are not specific to anyone OP, this approach is indicative only of general OP exposures to these mixtures and cannot be used to quantitatively assess individual OP dosimetry. More recently, Lavy et al. (1993) conducted a comprehensive year-long biomonitoring study of tree nursery workers, who are routinely exposed to multiple pesticides. In this study, it was recognized that as many as 28 pesticides are regularly used and 17 of the most common pesticides were selected for monitoring, including a number of OPs. Evaluation of the human and animal pharmacokinetic data suggested that adequate metabolism information was available on 8 of the selected pesticides to support biomonitoring. In this year-long study, 3134 urine specimens were analyzed, but only 42 of these contained measurable pesticide metabolites (1.3%) and were composed of only three pesticides (benomyl, bifenox, and carbaryl) (Lavy et at., 1993). In addition, based on a calculated margin of safety, the exposure levels were clearly below a level that would be of concern to human health. Biomonitoring strategies have also been successfully applied to quantitatively assess secondary exposures to OP insecticides resulting from both acute and chronic exposures. Richter et al. (1992) quantitated diethylphosphate in the urine of individuals who were symptomatic for OP exposure and resided in a house that had been sprayed with diazinon approximately 4.5 months earlier. In this particular study, very high levels of urinary diethylphosphate were observed in family members, whereas cholinesterase activity, although slightly depressed, was well within the range of "normal." The quantitation of urinary diethylphosphate was used to establish a persistent household exposure to diazinon residues as the most likely explanation. This study clearly illustrates the utility of urinary OP metabolites for quantitative biomonitoring of exposure. Biomonitoring based on the measurement of OP metabolites has also been used to compare pesticide exposure in children who live in proximity to high spray areas (i.e., orchards) and whose parents/guardians are pesticide applicators (Loewenherz et aI., 1997). Based on known pesticide use patterns, it was determined that OP insecticide exposure would be primarily associated with azinphos-methyl, chlorpyrifos, and phosmet. Therefore, the study focused on the quantitation of the alkylphosphate metabolites (dimethylthiophosphate, dimethyldithiophosphate, dimethylphosphate) in the children's urine. Loewenherz et al. (1997) collected and evaluated 160 spot urine specimens from 88 children and reported detectable
942
CHAPTER 46 Organophosphate Pharmacokinetics 700
600 Q,j
:::
·c
0
500
.S ~
~ ~
400
f-;
~
Q
-; .... 300 0
f-;
200
100
0 1
2
3
4
5
6
Amount Sprayed (kg a.i.) Figure 46.12 Relationship between the amount of alkyl phosphate (dimethylthiophosphate; DMTP) metabolite in urine of workers and the amount of active ingredient (a.i.) sprayed. Data obtained from FrankJin et al. (1981).
levels of these metabolites in 27 and 47 % of the reference children and applicator children, respectively. In addition, the biomonitoring data suggest that the children of applicators had a significantly higher dose than the reference children (0.021 vs. 0.005 J.lg/I, respectively). Biomonitoring of parent OPs and metabolites in blood and urine has also been used to provide a quantitative assessment of dosimetry in human poisoning victims following acute highdose exposures (Drevenkar et aI., 1993; Vasilic et aI., 1992). Although acute AChE depression (i.e., 50% of baseline) is used to substantiate OP poisoning, the analysis of intact pesticides or specific metabolites in body fluids (blood/urine) can be used to identify the specific causative chemical agent(s) (Ellenhom and Barceloux, 1988; Lotti et at. , 1986). The utilization of pharmacokinetic models such as the one developed for chlorpyrifos (Nolan et aI., 1984) can be extremely useful for the estimation of dosimetry under these acute exposure scenarios. To illustrate this point, a two-compartment pharmacokinetic model was used to fit pharmacokinetic data obtained from a poison victim who ingested a commercial insecticide formulation containing chlorpyrifos (Drevenkar et aI. , 1993). These same data have been modeled utilizing a PBPKJPD model developed for the quantitation of chlorpyrifos, chlorpyrifos-oxon, and TCP in the rat and human (Timchalk and Nolan, 1998). The time course and PBPKJPD model-predicted TCP and chlorpyrifos concentration in the blood and serum of human volunteers and following oral ingestion for a single poison victim are pre-
sented in Fig. 46.13. The model adequately reflects the data from these limited human samples; more important, these examples illustrate the strength of using pharmacokinetic models for quantitating dosimetry under both controlled and noncontrolled conditions. In summary, these examples have been presented to illustrate the practical application of pharmacokinetics to assess exposure to chemicals. Biomonitoring is clearly an integral component of the agricultural pesticide exposure assessment strategy. However, the successful application of biomonitoring for quantitative dosimetry is primarily limited by a lack of chemicalspecific pharmacokinetic data in humans.
46.3.4 STUDIES THAT FACILITATE EXTRAPOLATION OF DOSIMETRY AND BIOLOGICAL RESPONSE FROM ANIMALS TO HUMANS AND THE ASSESSMENT OF HUMAN HEALTH RISK Organophosphate insecticides constitute a large class of chemical pesticides that are widely used in the agricultural industry and in home applications. This suggests that there is significant potential for exposure and the health consequences of these exposures may be impacted by both interindividual and extrinsic variability (see Fig. 46.14). For example, extrinsic factors such as multiple exposure routes, chemical/drug interactions, and
46.3 Pharmacokinetic Approaches Applied to Organophosphate Insecticides
A.
10
0.1
-I-~-~-~-~-~~-~-~-~~
o
10
20
30
40
50
60
70
80
90
100
Time (h)
5
B.
10
~
~ ·c...
oS
..
c..
•
Q
:a
~ 0.1
::l.
0.01-1-_~_~_~_~_~_ _ _ _ _~
o
50
100
150
200
250
300
350
400
Time (h) Figure 46.13 (A) Mean blood time-course of 3,S,6-trichloropyridinol (TCP) from six human volunteers administered a single oral dose of O.S mg chlorpyrifos (CPF)lkg of body weight. Data obtained from Nolan et al. (1984). (B) Time-course of CPF in the serum of a single poison victim that orally ingested a commercial insecticide product containing CPF. The symbols represent observed data while the line represents the model prediction. Data obtained from Drevenkar et al. (1993).
variable exposure rates may significantly modify the toxicological response to OPs. In addition, person-to-person differences in metabolism, genetic predisposition, physical environment, and age (infant, children, and elderly) are important determinants of pharmacokinetic and/or pharmacodynamic response. As previously discussed, Gearhart et al. (1990) developed a PBPKlPD model for quantitative OP dosimetry and AChE inhibition utilizing diisopropylfluorophosphate as a representative highly toxic OP (see Figs. 46.6 and 46.7). This PBPKlPD model was developed as a prototype that could readily be extended to the commercially important OPs. In this respect, the development and application of PBPKlPD mode ling represents a logical approach for assessing risk and understanding the toxicological implications of known or suspected exposures to various OP insecticides. To extend this initial modeling effort, Gearhart et al. (1994) modified their original model to incorporate the phosphorothionate insecticide parathion and its oxon, paraoxon. In addition, Timchalk and Nolan (1998) built a model to incorporate chemical-specific parameters for the OP insecticide chlorpyrifos that are based on the models of Gearhart et al. (1990, 1994). This model was developed to describe the time course of chlor-
943
pyrifos, the oxon metabolite, the elimination of TCP, and the inhibition of target esterases by the oxon. The chlorpyrifos PBPKlPD model incorporates CYP450-mediated activation of chlorpyrifos to the oxon and detoxification to TCP. In addition, hydrolysis of the oxon by B-esterases (AChE, BChE, and CaE) is modeled in the liver, blood, diaphragm, and brain. A diagram outlining the critical features of this model is presented in Fig. 46.15. Although this is a preliminary PBPKlPD model requiring further validation and refinement, it qualitatively behaves consistent with the general understanding of OP toxicity, pharmacokinetics, and pharmacodynamic responses. To illustrate this point, the model has been used to simulate the serum time course of chlorpyrifos and TCP in poisoned humans (see Fig. 46.13). In addition, a simulation of the dynamics of tissue esterase inhibition following a single acute exposure to two different doses of chlorpyrifos in the rat is qualitatively consistent with observed AChE inhibition kinetics and is illustrated in Fig. 46.16. It is anticipated that this basic model structure can be readily modified to accommodate other important phosphorothioates. Likewise, once validated, these models can be used as a foundation for understanding complex-mixture interactions, sensitive subpopulations, and the role of metabolic polymorphisms in altering dosimetry and biological response.
46.3.4.1 Organophosphate Mixtures Both occupational and residential exposures to OP insecticides often entail simultaneous or sequential contact with OP mixtures (Hayes et aI., 1980; Lavy et aI., 1993; Loewenherz et aI., 1997). The potential for OP interactions has been well understood for some time. Early studies demonstrated the acute, synergistic, and toxicological interactions between the OPs malathion and EPN (ethyl-p-nitrophenyl phenylphosphonothionate) (Frawley et aI., 1957). In addition, non-OPs have been reported to influence the pharmacokinetic and toxicological response of OPs. For example, phenobarbital or alcohol pretreatment of mice protects against the acute toxicity of chlorpyrifos and parathion, respectively (O'Shaughnessy and Sultatos, 1995; Sultatos, 1988). Wu et al. (1996) reported that pretreatment of rats with cimetidine potentiated the acute toxicity of diazinon as a result of reducing diazinon total body clearance. Likewise, co-administration of diazinon with cocaine significantly increased the concentration of cocaine and norcocaine in the blood and tissues of mice apparently due to competition for esterase enzyme detoxification (Benuck et aI., 1989; Kump et al., 1994). A combination of malathion and the carbamate pesticide carbaryl alters the fundamental pharmacokinetic properties of the individual compounds and it has been suggested that this may explain some of the observed toxicity seen from exposure to this chemical mixture (WaldronLechner and Abdel-Rahman, 1986). OP pesticides as a class of compounds share common metabolic processes for activation and detoxification as well as a common mechanism for toxicological response through the inhibition of AChE (Murphy, 1986; Sultatos, 1994). Based on similar pharmacokinetic and mode-of-aetion properties,
944
CHAPTER 46
Organophosphate Pharmacokinetics
Extensive toxicological & pharmacokinetic database for OP pesticides
D Focusedin vivo & in vitro studies
t
Integrated PBPKlPD model
~
Y-----J
Biomonitoring Studies
D Health risk from OP exposure
Figure 46.14 A conceptual model for evaluating the impact of intrinsic/extrinsic factors on human health risk from exposure to organophosphate (OP) insecticides.
the potential for interactions between mixtures of OPs has been hypothesized. Organophosphates can interact at a number of important metabolic steps (see Table 46.5), including (1) CYP450-mediated activation/detoxification, (2) plasma protein binding, (3) A-esterase detoxification, and (4) AChE binding/inhibition. The net effect of these interactions (additivity, synergy, or antagonism) will be dependent on the specific OP mixture, dose ranges of exposures, and sensitivity of the individual. Several integrated approaches have been proposed to investigate the potential for toxicological interactions of chemical mixtures (EI-Masri et aI., 1997). The proposed strategies all emphasize the utilization of PBPK modeling. Until recently, the majority of these models have focused on a single chemical exposure. However, exposure to low-dose chemical mixtures represents a more realistic exposure scenario. Although several PBPK models have been developed for binary and ternary mixtures of organic solvents, little is known about the potential metabolic and toxicological interactions of these mixtures on biological systems (Andersen et aI., 1987b; Pelekis and Krishnan, 1997; Purcell et aI., 1990; Tardif et aI., 1993, 1995, 1997). The evaluation of mixtures is complicated by the myriad of chemicals, doses, exposure routes, and dynamic responses observed, making it impractical to effectively test all possible permutations. In this regard, the application of PBPK models for mixtures provides a limited means to quantitatively describe the disposition of chemicals as a result of exposure to various combinations, doses, and routes of administration.
46.3.4.2 Sensitive Subpopulations (Children
and Polymorphisms) There is currently a significant focus and concern over the potential increased sensitivity of infants and children to the toxic effects of chemicals. The importance of this issue is highlighted by the National Research Council's report, Pesticides in the Diets of Infants and Children, and the passage of the Food Quality Protection Act. It is recognized that children are not just "small adults," but rather a unique subpopulation that may be particularly vulnerable to chemical insult. Age-dependent changes in a child's physiology (i.e., body size, blood flow, organ functions) and metabolic capacity (i.e., phase I and 11 metabolism) may significantly impact their response to a chemical insult, resulting in either beneficial or detrimental effects (Miller et aI., 1997). Clear variability in the capacity to detoxify environmental chemicals has been established in both animals and humans. However, the current risk assessment paradigms do not adequately consider the implications of these differences on the risk to infants and children. Numerous studies have demonstrated that juvenile animals are more susceptible to the acute effects of OP insecticides than adults (Benke and Murphy, 1975; Brodeur and DuBois, 1963; Gaines and Linder, 1986; Harbison, 1975; Moser and Padilla, 1998; Pope and Liu, 1997; Pope et aI., 1991). This greater sensitivity has primarily been attributed to the lack of complete metabolic competence during neonatal and postnatal development (Benke and Murphy, 1975).
Chlorpyrifos Qc
Chlorpyrifos (CPF) Specific Parameters
'.Cl ~
Ut
B-Esterase Inhibition (shaded compartments)
"Free" O xon +
Ko = First-order oral absorption rate (hr-1) CalCao = CPF or CPF -oxon arterial concentration (JlmolelL) Cv/Cvo= CPF or CPF -oxon venous concentration (JlmolelL) Q = Blood flow to tissues (Llhr) Vmax = Maximum rate of CYP450 hydrolysis to TCP and formation of oxon (Jlmolelhr/kg) Km = Michaelis constant for CYP450 in tissues (JlmolelL) Compt A = Compartment model for urinary elimination of TCP Ke = First-order elimination rate (hr-1) Other model parameters for mass balance and esterase inhibition as described in Figs. 49.6 and 49.7
Free Esterase hr"
Degradation of Esterase
Released MetaboJite
F igure 46.15 Physiologically based pharmacokinetic and pharmacodynamic (PBPKlPD) model structure describing the distribution of chlorpyrifos (CPF), CPF-oxon and 3,5,6-trichloropyridinol (TCP) in the rat.
B.
A.
. - - Brain
100
100
90
90
80
80
_
70
'0 ...
0
.,l:j
60
o U
50
=
~
.a;;..
0\
~
...... o
=
U
40
0 ~
Brain
70
~
60 50
~
40
~
3D
Diaphragm
0
30 20
20
10
10
o
3
6
9
12 15 18 21 24 27 30 33 36 39 42 45 48 51 54 57 60 63 66 69 72
Time (hrs)
Hepatic
o
3
6
9 12 15 18 21 24 27 3D 33 36 39 42 45 48 51
Time (hrs)
Figure 46.16 Simulated AChE inhibition in selected tissues (brain, diaphragm, blood and liver) following a single oral dose of (A) 0.5 rug and (B) 5 mg chlorpyrifos (CPF)lkg of body weight.
54 57 60 63 66 69 72
46.3 Pharmacokinetic Approaches Applied to Organophosphate Insecticides
947
Table 46.5 Important Metabolic and Response Interactions for Mixtures of Organophosphate Insecticides
Parameter
Importance
Type of chemical interaction
Implication
CYP4S0 mixed-function
Metabolic activation!
Substrate (parent compound)
Changes in oxon
oxidase metabolism
detoxification of parent
competition for enzyme
concentrations
compound Reversible plasma protein binding A-esterase metabolism
Systemic transport of
Substrate (parent compound)
parent compound
competition for available
parent chemical available
protein binding sites
for metabolism
Important metabolic step responsible for
Substrate (oxon) competition for enzyme
Increased levels of "free"
Changes in oxon concentrations
detoxification AChE binding/inhibition
Toxicological response
Substrates (oxon) combine to increase inhibition of AChE
Several recent studies provide important perspective on this age-dependent sensitivity and the selected results are illustrated in Fig. 46.17. Atterberry et al. (1997) evaluated the developmental changes in brain AChE levels and hepatic CaE- and CYP450-mediated metabolism of chlorpyrifos in the juvenile rat. They correlated age-dependent chlorpyrifos toxicity (Fig. 46.17a) with the capacity to activate and detoxify chlorpyrifos (Fig. 46.17b and c). Their results suggest that the age-dependent sensitivity of young animals is associated with a decreased CaE-mediated hydrolysis and CYP450-mediated dearylation capacity in young animals relative to adults. In addition, it has likewise been suggested that the sensitivity of young animals is associated with a lower A-esterase activity than is found in adults (Mortensen et aI., 1996). The importance of A-esterase in protecting against OP toxicity has been demonstrated in several studies in which exogenous administration of A-esterase can protect against OP poisoning in rodents (Costa et aI., 1990; Li et aI., 1993, 1995; Main, 1956). Mortensen et al. (1996) compared both plasma and hepatic A-esterase enzyme activity in adult and neonatal animals and reported that neonatal plasma and liver A-esterase activity were 9 and 50% of adult activity, respectively. This is consistent with the results of Li et al. (1997), who reported that serum A-esterase activity toward paraoxon, chlorpyrifos-oxon, and diazoxon were very low at birth, but reached adult levels of activity by about 25 days of age in rodents. This time span in rodents corresponds to in utero humans to approximately in 6 months old. These findings in animals are in agreement with observations in which newboms and children less than 6 months old have lower plasma A-esterase activity than adults (Augustinsson and Barr, 1963; Mueller et aI., 1983). In addition to age-dependent differences in A-esterase activity, a human and animal genetic polymorphism has been well established (Eckerson et aI., 1983; Furlong et aI., 1988; Geldmacher-von Mallinckrodt et al., 1983). This polymorphism is known to result in the expression of a broad range of A-esterase en-
Increased toxicity due to additive response
zyme activity within a large segment of the human population. Although young rats appear to be more sensitive (based on LDso and AChE inhibition) to the acute effects of OP insecticides relative to adults, AChE activity is reported to recover faster in young animals (Moser and Padilla, 1998; Pope et aI., 1991). This more rapid recovery of AChE (Lajtha and Dunlop, 1981; Moser and Padilla, 1998) is associated with a faster synthesis rate as well as higher steady-state enzyme levels. The capacity of young rats to recover AChE activity faster than adults may be of greater importance in dealing with intermittent low-dose exposure to OPs. In summary, these data suggest that the sensitivity of juvenile animals and humans to the toxic effects of OP insecticides may be a function of the maturational stage of development for a number of critical metabolic steps. These include CYP450-mediated activation/detoxification, hepatic CaE hydrolysis binding, and plasma and hepatic A-esterase detoxification. The application of PBPKJPD mode ling offers a unique opportunity to integrate age-dependent changes in OP metabolic activation and detoxification pathways into a comprehensive model that is capable of quantitating dose and response across all ages. In this context, PBPK models are being extended to the mode ling of chemical exposure in developing neonatal animals. A number of these models have focused on the incorporation of xenobiotic lactational transfer to nursing pups (Byczkowski et aI., 1994; Fisher et aI., 1990; Sundberg et aI., 1998). Based on the potential sensitivity of children to OP insecticides, there is a need to develop quantitative models that can be used to assess the risk associated with OP exposure in infants and children. However, there are currently no published models available for OPs that are readily applicable for quantitating age-dependent changes in dose and response. For toxicants that have long residence times within the body, there is a need to develop PBPK models that appropriately incorporate growth and maturational development of physiological and metabolic function (O'Flaherty, 1991a).
948
CHAPTER 46
Organophosphate Pharmacokinetics 0.4
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Figure 46.17 (A) Inhibitory concentration 50% (ICso) of chlorpyrifos-oxon (CPF-oxon) (nM) to CaB activity in the liver of rats with increasing age, (B) CPF (50 J.lM) CYP450 desulfation and dearylation by rat hepatic microsomes with increasing age, (C) Developmental pattern of liver CaB activity with increasing age in the rat. All data was adapted from Atterberry et at. (1997).
O'Flaherty (1991a, b, 1993) has developed a series of PBPK models that begin to describe lead kinetics from birth to adulthood. Ultimately these models will enable quantitation of lead dosimetry over an entire lifetime. It is envisioned that the framework of the PBPKlPD model that has been developed for diisopropylfiuorophosphate, parathion and chlorpyrifos (Gearhart et aI., 1990, 1994; Sultatos, 1990; Timchalk and Nolan, 1998) can readily be extended to incorporate agedependent changes in CYP450, CaE, A-esterase and AchE activity. In summary, pharmacokinetics and in particular the application of PBPKlPD mode ling have been shown to be extremely useful approaches for dosimetry and biological response extrapolation for the assessment of human health risk from chemical exposures. The utilization of PBPKlPD modeling to address OP insecticide toxicity issues is particularly intriguing since these models can be used to assess the health consequences of both inter-individual (i.e., age, gender) and extrinsic factors (i.e., multiple exposure routes, chemicaIJdrug interactions and variable exposure rates) that may significantly modify the toxicological response to OPs.
46.4 SUMMARY AND CONCLUSIONS This chapter has illustrated a number of current and future applications of pharmacokinetics to assess OP dosimetry, biological response and risk in humans exposed to these insecticides. Pharmacokinetics is concerned with the quantitative integration of absorption, distribution, metabolism and excretion and can be used to provide useful insight into the toxicological responses associated with OP insecticides. Since OP insecticides share a common mode of action through their capability to inhibited AChE activity, it is feasible to develop pharmacokinetic strategies that link quantitative dosimetry with biologically based pharmacodynamic (PD) response modeling. Pharmacokinetic studies that have been conducted with OP insecticides in mUltiple species, at various dose levels, and across different routes of exposure have provided important insight into in vivo behavior of these OPs. The development and application of pharmacokinetic models capable of describing uptake, distribution, metabolism, and elimination of OP insecticides in humans represent a crucial research element needed for quantitative biomonitoring. In this regard, the successful application of biomonitoring for quantitating OP dose is primarily
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Sundberg, J., Jonsson, S., Karlsson, M. 0., Palminger Hallen, 1., and Oskarson, A. (1998). Kinetics of methylmercury and inorganic mercury in lactating and nonlactating mice. Toxicol. Appl. Pharmacol. 151, 319-329. Tardif, R., Charest-Tardif, G., Brodeur, J. and Krishnan, K (1997). Physiologically based pharmacokinetic modeling of a ternary mixture of alkyl benzenes in rats and humans. Toxicol. Appl. Pharmacal. 144, 120-134. Tardif, R., Lapare, S., Charest-Tardif, G., Brodeur, J., and Krishnan, K (1995). Physiologically-based pharmacokinetic modeling of a mixture of toluene and xylene in humans. Risk Anal. 15, 335-342. Tardif, R., Lapare, S., Krishnan, K, and Brodeur, J. (1993). Physiologically based modeling of the toxicokinetic interaction between toluene and mxylene in the rat. Toxicol. Appl. Pharmacol. 120, 266-273. Timchalk, c., and Nolan, R. J. (1998). Physiologically based pharmacokinetic/pharmacodynamic (PBPKJPD) modeling of chlorpyrifos and its oxon metabolite in the rat. Toxicol. Sci. 42, 693. Tomokuni, K, Hasegawa, T., Hirai, Y., and Koga, N. (1985). The tissue distribution of diazinon and the inhibition of blood cholinesterase activities in rats and mice receiving a single intraperitoneal dose of diazinon. Toxicology 37,91-98. Vale, J. A. (1998). Toxicokinetic and toxicodynamic aspects of organophosphate (OP) insecticide poisoning. Toxicol. Lett. 102-103,649-652. Vale, J. A., and Scott, G. W. (1974). Organophosphate poisoning. Guy's Hosp. Gazette 123, 12-25. Vasilic, Z., Drevenkar, Rumenjak, Stengl, B., and Frobe, Z. (1992). Urinary elimination of diethylphosphorus metabolites in persons poisoned by quinalphos or chlorpyrifos. Arch. Environ. Contam. Toxicol. 22, 351-357. Waldron-Lechner, D., and Abdel-Rahman, M. S. (1986). Kinetics of carbaryl and malathion in combination in the rat. 1. Toxicol. Environ. Health 18, 241-256. Wester, R. c., and Maibach, H. 1. (1983). Cutaneous pharmacokinetics: 10 steps to percutaneous absorption. Drug Metab. Rev. 14, 169-205. Wester, R. c., Maibach, H. 1., Bucks, D. A. w., and Guy, R. H. (1983). Malathion percutaneous absorption after repeated administration to man. Toxicol. Appl. Pharmacol. 68, 116-119. Wester, R. c., Maibach, H. 1., Melendres, J., Sedik, L., Knaak, J., and Wang, R. (1992). In vivo and in vitro percutaneous absorption and skin evaporation of isofenphos in man. Fundam. Appl. Pharmacol. 19,521-526. Wester, R. c., Sedik, L., Melendres, J., Logan, E, Maibach, H. 1., and Russell, 1. (1993). Percutaneous absorption of diazinon in humans. Food Chem. Toxical. 31,569-572. Wilks, M. E, and Woollen, B. H. (1994). Human volunteer studies with nonpharmaceutical chemicals: Metabolism and pharmacokinetic studies. Hum. Exp. Toxicol. 13,383-392. Woollen, B. H. (1993). Biological monitoring for pesticide absorption. Occup. Hyg. 37, 525-540. Wu, H. X., Evrenx-Gros, C., Descottes, J. (1996). Diazinon toxicokinetics, tissne distribution and anticholinesterase activity in the rat. Biomed. Environ. Sci. 9, 359-369.
v.,
v.,
CHAPTER
47 Neuropathy Target Esterase Martin K. Johnson and Paul Glynn Medical Research Council Toxicology, UK
47.1 INTRODUCTION Neuropathy target esterase (NTE) is a biochemical mystery. This protein, present in neural tissue, has the capacity to catalyze rapid hydrolysis of certain (unphysiological) carboxylate ester substrates, and, similar to acetylcholinesterase (AChE), its catalytic activity is inhibited by covalent reaction with a variety of progressive inhibitors, including organophosphorus esters (OPs). However, mere loss of the catalytic activity of NTE seems not to be deleterious in adult animals. Rather, the nature of the chemical group covalently bound at the catalytic centre determines whether toxicological effects will follow. Thirty years after its discovery, there is now overwhelming evidence that initiation of OPIDN (organophosphorus ester-induced delayed neuropathy) starts with NTE: Initiation may occur within hours of ingestion of a single dose of some OPs although clinical expression is deferred for 1-4 weeks. The clinical and morphological features of OPIDN are described in Chapter 49 as are some of the biochemical and physiological changes reported to accompany (consequentially or causally) the development of the syndrome after initiation. In this chapter, we review briefly the evidence for NTE as the target and the mechanism of initiation and the possible involvement of related esterases to promotion of OPIDN. This is followed by a discussion of the application of NTE studies to human risk evaluation. Finally, we consider the nature and properties of this remarkable protein, its function in neurons, and its role in OPIDN.
47.2 BRIEF REVIEW OF THE EVIDENCE 47.2.1 ORGANOPHOSPHORUS ESTER PESTICIDES: GENERAL REACTIONS WITH SERINE ESTERASES Acetylcholinesterase, the intended target for OP pesticides, is a member of a large family of serine esterases (Krejci et aI., 1991; Taylor, 1992). A particular serine residue at the active site of these enzymes is rendered reactive by the presence of a histidine and a glutamate or aspartate, the components of the Handbook of Pesticide Toxicology Volume 2. Agents
catalytic triad. Serine esterases cata1yze hydrolysis of carboxylate esters by the mechanism shown, in highly simplified form, in Fig. 47.1a, which involves formation of a covalent acyl enzyme intermediate. This mechanism is essentially the same in serine proteases such as chymotrypsin, which cleave peptide bonds with intermediate formation of a covalent acyl enzyme (Aldridge and Reiner, 1972). OP pesticides have been made from a variety of organophosphates, -phosphonates, -phosphinates, or -phosphoramidates. As a class, these compounds are hydrolyzab1e esters and act as pseudo-substrates for a variety of serine esterases and proteases. As an example, the reaction of an organophosphate with a serine esterase is shown in simplified form in Fig. 47.1 b. The rate of hydrolysis of the phosphorylated enzyme is greatly (6-10 orders of magnitude) reduced compared to that of the acyl enzyme. Thus, the enzyme becomes virtually permanently inhibited, although certain nucleophilic agents such as oximes or fluoride anion can catalyze a speedier dephosphorylationhence the therapeutic use of oximes in the treatment of acetylcholinesterase poisoning. However, the phosphorylated enzyme can subsequently undergo a second reaction, known as aging, which, in the case shown of an organophosphate, results in the liberation of one of the bound R groups into solution (Fig. 47.1b). This leaves the active-site serine covalently attached to a negatively charged monoorganophosphoryl moiety, which is significantly more resistant to removal by therapeutic nucleophiles. Organophosphinates, in which both R groups are directly attached to the phosphorus atom, also covalently react with the active-site serine but cannot undergo the aging reaction (Fig. 47.1c). 47.2.2 NEUROPATHY TARGET ESTERASE AS THE TARGET FOR INITIATION OF ORGANOPHOSPHORUS ESTER-INDUCED DELAYED NEUROPATHY Elucidation of the events that initiate OPIDN involved a sequence of observations in vitro and ex vivo on the interaction of radiolabeled diisopropyl phosphorofluoridate (DFP) and other
953
Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
954
CHAPTER 47
Neuropathy Target Esterase (a) Substrate
E OH
+ Substrate :;::::::
[E
+ Alcohol Q!: Phenol
OH··· SUbstr.]- Eo -Acyl
tL.-_ _ _ _ _ _-=-__oe-c_ _:o..-_-----'I'-Acid (b) Organophosphate
E
to-R ~ OH + X-P 'O-R -
~E OH ... X-Pt
o- R] 'O-R
---l.~
E
to-R O-P 'O-R
~
E
0
O-~/O'OR
Aged inhibited enzyme
(c) Organophosphinate
o
E
"_ _
0II/R ]
R
OH + X-P'R ~
[
EOH ... X-P'R
No aging possible Figure 47.1 Reaction of a serine esterase with a carboxylate ester substrate and with organophosphate and organophosphinate inhibitors. The reactive serine residue at the enzyme active site is represented by -OH. (a) Following reversible formation of an enzyme-substrate Michaelis complex (in square brackets), the serine residue makes a nucleophilic attack on the acyl carbon of the ester and forms a tetrahedral hemiacetal intermediate (not shown). The alcohol moiety is rapidly expelled from this intermediate to produce a covalent acyl enzyme. Rapid aqueous hydrolysis of the acyl enzyme liberates the carboxylic acid and regenerates free enzyme. (b) Part of the efficacy of OP esters as serine esterase inhibitors results from their structural resemblance to the tetrahedral hemiacetal intermediate formed between the enzyme and the carboxylate ester substrate. The rate of hydrolysis of the organophosphorylated or organophosphinylated enzyme is much slower than that of the acyl enzyme, resulting in essentially irreversible inhibition. In addition, organophosphates, but not organophosphinates, are able to undergo a second reaction termed aging. This entails loss of one of the R groups from the organophosphorylated enzyme leaving a negatively charged species attached to the active site.
(unlabeled) esterase inhibitors (OPs, carbamates, and sulfonyl fluorides) with homogenized whole brain of adult chickens: The evidence, doubts, and arguments are detailed by Johnson (1990). In brief, because preliminary study showed that DFP bound only to proteins, criteria were adopted for a putative target site which would be a protein able to undergo the same set of general reactions shown in Fig. 47.1. It was found that about 5% of total DFP-labeling sites in hen brain were not covalently blocked by a variety of OPs known not to cause OPIDN but were inhibited at toxicologically relevant doses by neuropathic compounds. Extensive screening then showed that only two out of more than 60 hydrolyzable esters, lipids, and or peptides competed with labeling of that subset, suggesting that only these two had significant affinity for the relevant DFP-reactive sites. Finally, only one of several esterases that could hydrolyze these competitors (phenyl esters of phenylacetic acid and valeric acid) shared the same inhibitor responses as the apparently homogeneous target site (John son, 1969a, b, 1970). This esterase was dubbed NTE for neurotoxic esterase and, later, neuropathy target esterase; its ability to hydrolyze phenyl valerate (PV) in a reaction sensitive to neuropathic OPs forms the basis for a widely used in vitro screening test for such OPs. A threshold level of about 70-90% inhibition of NTE in the brain and spinal cord of test adult chickens has been found to be the
norm for precipitation of clinically visible OPIDN, and this criterion has been incorporated into current regulatory guidelines (see Section 47.3). Progressive inhibitors of NTE were found to fall into two classes: Some were neuropathic but others, which were unable because of their chemical structure to undergo an aging reaction analogous to that shown in Fig. 47.1b, were not only not neuropathic, but they were actually specifically prophylactic against OPIDN (not against acute anticholinesterase effects) if given to chickens prior to a neuropathic ~P. These prophylactic compounds included carbamates (Johnson and Lauwerys, 1969), sulfonyl fluorides (Johnson, 1970), and phosphinates (as in Fig. 47.1c; Johnson, 1974). The most striking evidence for the validity of NTE as the true target has been the correlation in time of the degree of inhibition of NTE (and of the radiolabeled target) by the nonaging compounds with their prophylactic effect: both short-term (hours) and long-term (up to 5 days) prophylaxis is possible according to the structure of the agent and the "window" correlates with the persistence of inhibition until about 70-80% of NTE is again available to neuropathic challenge by virtue of reactivation or turnover of the inhibited enzyme [summarized by Johnson (1990)]. OPIDN prophylaxis by nonaging NTE inhibitors led to the proposal that generation of a negative charge at the active site
47.3 Toxological Applications
of NTE was critical for initiation of OPIDN (John son, 1974). A neat confirmation of the importance of the aging reaction in initiation is the demonstration that, in the case of chiral phosphonates such as the oxon of EPN (O-ethyl 0-4-nitrophenyl phenylphosphonothioate), both enantiomers can inhibit NTE at tolerable doses in vivo, but the one that engages in the subsequent aging reaction causes OPIDN whereas the other is prophylactic (Johnson and Read, 1987). This raises special questions about the interpretation of regulatory tests for chiral compounds (see Section 47.3.3.3). 47.2.3 POSSIBLE INVOLVEMENT OF OTHER ESTERASES IN ORGANOPHOSPHORUS ESTER-INDUCED DELAYED NEUROPATHY
Accumulated evidence has supported the early identification of NTE as the site for initiation of OPIDN. However, although all detectable binding sites for [32P]-labeled DFP in brain were dissected, it has been proposed that further sites were actually present at such low concentration that they were undetectable but might have equal claim to be the initiating site. No convincing evidence for this idea has emerged over many years, but the role of the recently identified "soluble NTE" is worthy of further consideration (Vilanova et aI., 1999). NTE is only one of at least six distinct phenyl valerate hydrolases easily demonstrated to be present in hen brain (Johnson, 1982a) and, until recently, no association of any of these with normal or abnormal processes had been demonstrated. However, Vilanova and colleagues have made extensive studies of NTE and related enzymes in sciatic nerve of the hen because that tissue undergoes more obvious degenerative changes in OPIDN than does brain [reviewed by Vilanova et al. (1999)]. They have identified and studied what appears to be a freely soluble enzyme behaving rather like NTE according to inhibition characteristics. Apart from having a lower apparent molecular size than the NTE subunit from brain particles (Escudero and Vilanova, 1997), the most striking characteristic of this and related soluble sciatic nerve PV hydrolases is the sensitivity of one or more of these esterases to paraoxon with which they react rapidly and progressively to form covalently inhibited enzymes, which then spontaneously reactivate within a few hours at body temperature. Thus, in the standard in vitro laboratory assay for NTE, such enzymes are excluded by the paraoxon preincubation but, in vivo, they would, for practical purposes, appear to be insensitive to that OP and could therefore be added to the list of putative targets for the site of initiation of OPIDN (Barril et aI., 1999). The uncovering of these enzymes is a tribute to more careful kinetic analysis of the OP/enzyme interaction than has been possible in some general screening operations. Furthermore, these researchers have shown that paraoxon competes strongly with substrate in the assay of soluble NTE so that the activity in sciatic nerve has routinely been underestimated by 20-fold or more (Barril and Vilanova, 1997). Fortunately, neither of these confusing factors seems to exist with regard to
955
particulate NTE. The possible relevance of any of these soluble enzymes to either initiation of OPIDN or to its promotion (a phenomenon discussed in Chapter 49) requires evaluation by structure-activity studies using autopsy tissue from hens dosed with a battery of neuropathic and nonneuropathic OPs. Thus far, studies in vitro with the soluble enzymes (Vilanova et aI., 1999) discourage an association because sensitivity to mipafox in vitro is so considerable (150 less than 0.1 f.lM compared with 7 f.lM for NTE) that one might expect that such activities would be fully inhibited in vivo at doses that would barely affect NTE and would be well below the neuropathic or promoting dose. Lotti and colleagues have produced some correlative structure-activity evidence to suggest that the target for promotion of OPIDN by certain OPs is not one of the two paraoxon-sensitive hydrolases first identified by Poulsen and Aldridge (1964) but may be a component of the particulate PV hydrolases, which they have dubbed "M-200" on the basis of its comparatively low sensitivity to mipafox. M-200 is a portion of the C activity (see Section 47.4.2), which can be segregated in assays where mipafox concentration is raised (from the 50 f.lM used in standard assays) to detect an enzyme with an 150 of 200 f.lM (Lotti and Moretto, 1999; Milatovic et aI., 1997). Further evaluation of these encouraging observations is in progress (M. Lotti, private communication). Also, Vilanova and colleagues have identified a component of soluble PV hydrolases that fits the characteristics of the target for promoters rather than initiators of OPIDN (Cespedes et aI., 1997; Vilanova et aI., 1999).
47.3 TOXOLOGICAL APPLICATIONS 47.3.1 HEN TEST
As noted in Chapter 49, OP pesticides are screened for their relative abilities to inhibit acetylcholinesterase and to cause OPIDN; consequently, humans and susceptible animals are unlikely to develop OPIDN without acute cholinergic toxicity following OP pesticide exposure. The hen test has developed in sophistication from its origins in studies in the early 1930s, which showed that mature adult chickens were the most satisfactory animals to detect the OPIDN potential of various OPs found in the toxic "Ginger Jake" drink that paralyzed thousands of people in the southern United States (Smith et aI., 1930). OPIDN in the adult chicken is, at present, the best available model for the human syndrome. Historically, it was found that hens responded positively and uniformly in tests of compounds suspected of causing OPIDN in humans. The best known examples were tri-cresyl phosphates (not insecticides) and mipafox and leptophos which were pesticides responsible for OPIDN incidents in the United Kingdom and the United States. The spinal tracts, which are known to suffer selective damage in human OPIDN, are well-developed in hens and damage to these regions is easily detected by standard techniques (see Chapter 49). Furthermore, slight abnormalities of gait are easily detected in these bipeds. Biochemically, the sensitivity of hen and
956
CHAPTER 47
Neuropathy Target Esterase
human NTE to inhibition by OP inhibitors is similar (Lotti and Johnson, 1978). Since the 1950s, the UK Ministry of Agriculture Fisheries and Food has required information about the OPIDN potential of OP pesticides submitted for registration. Briefly, batches of hens are required to be dosed with about the maximum tolerable dose accompanied by therapeutic measures to carry them through the inevitable cholinergic crisis. After observation over a period of 21 days for signs of ataxia, positively affected birds are autopsied and histopathological signs are sought: Unaffected birds are redosed and the observation is continued for an additional 21 days with sample birds being autopsied at the conclusion regardless of the presence or absence of clinical signs. Since the 1970s, the U.S. Environmental Protection Agency (U.S. EPA) has required clinical and histopathological tests after long-term (often 90 days) feeding of tolerable low levels of OPs. However, in spite of such testing procedures becoming mandatory, reports continue to appear of occasional cases of full-blown OPIDN resulting from occupational exposure or suicide attempts with pesticides, including methamidophos, leptophos, dichlorvos, trichlorphon, chlorpyrifos, and EPN (Lotti, 1992). The possibility of less severe clinical cases being overlooked has caused concern, and improvements in the discriminatory power of regulatory tests for OPIDN are clearly desirable. Assay of NTE activity in appropriate autopsy samples taken soon after dosing became first an optional extra, then recommended, and finally a required component of OPIDN toxicity tests over a period of about 20 years following the first reports of the enzyme in 1969, although some manufacturers adopted the procedure voluntarily very early (U.S. EPA, 1991; OECD, 1995). The biochemical test does not replace clinical and histopathological observation but, rather, complements them (ECETOC, 1998; Johnson, 1984). It has the advantage that, unlike those subjective and qualitative tests, it is quantitative so that the degree of risk of OPIDN arising from a defined dose can be assessed: hitherto the conclusion from tests could only be 'Yes/No" or occasionally "Marginal." Furthermore, sufficient data on the relationship of chemical structure to neuropathic response have been accumulated in experimental studies in vivo (John son, 1975a) and in vitro (the latter using both human and animal tissues; Johnson, 1975b, 1988; Lotti and Johnson, 1978) that fairly confident predictions of the OPIDN potential of an untested compound can be made: Monitoring of NTE activity in accessible tissue samples taken from a few patients who have deliberately ingested known OPs confirms the relationship (Moretto and Lotti, 1998). Although the early work studied brain tissue only, it is accepted that clinical signs of OPIDN reflect the fact that neuropathic lesions are scarce in brain and more marked in spinal cord and peripheral nerve (see Chapter 49). In some early experiments with dichlorvos (John son, 1978), the dose (although many times the LDso) appeared not to have reached the spinal cord sufficiently to inhibit NTE and no clinical signs developed in pair-dosed birds although brain NTE was in-
hibited. A further dose was necessary to increase inhibition in spinal cord and to precipitate clinical neuropathy. For this reason, it is now customary for ex vivo assays from dosed birds to study both tissues and it is not uncommon to find slightly less inhibition in cord than in brain: The threshold figure of inhibition is accordingly set a little lower for spinal cord. 47.3.2 STRUCTURE-ACTIVITY RELATIONSHIPS AND PREDICTION OF ORGANOPHOSPHORUS ESTER-INDUCED DELAYED NEUROPATHY POTENTIAL IN HENS
The ability to analyze both positive and negative clinical responses in terms of the degree of effect on NTE during OPIDN tests made it possible to review a large amount of test data and to define guidelines for predicting neuropathic potential of both the plasticizer and the pesticide types of OP esters (Johnson, 1975a, 1982a). Guidelines (to be taken in concert) concerning the likelihood that a pesticide-type OP ester with general structure R I R2p(O or S)X will cause neuropathy at less than lethal doses are as follows:
1. Factors that increase OPIDN potential more than acute toxicity include: a. Choice of phosphonates or phosphoramidates rather than analogous phosphates h. Increase in chain length or hydrophobicity of R I and R2 c. A leaving group X that does not sterically hinder approach to the active site of NTE 2. Factors that decrease the comparative potential include:
a. The converse of factors 1a-c h. Choice of R or X groups that are very bulky (naphthyloxy) or nonplanar c. Choice of a nitrophenyl group at X d. Choice of comparatively more hydrophilic X groups (oximes or heterocyclics) e. Choice of thioether linkages at X
Although inclusion of NTE assays in toxicological tests is now mandatory for all OP pesticides submitted for registration or re-registration, most data do not reach the open literature. Table 47 .1 lists the results for compounds reported for 1975-1981 (John son, 1982a). Considering the structure-activity factors, it is clear why malathion, parathion, fenitrothion, and diazinon among the compounds listed are all far below the hazard line for OPIDN and why EPN, a phosphonothioate with a hydrophobic phenyl group at R, is neuropathic even with a 4-nitrophenyl leaving group. Also, it is not surprising that other phenylphosphonothioates such as desbromo-Ieptophos or cyanofenphos are also neuropathic (Johnson, 1975a; Soliman et aI., 1986) and
47.3 Toxological Applications
957
Table 47.1 NTE and Clinical OPIDN Responses of Pesticides Reported 1975-1981 a Brain enzyme Dose (mglkg)3
Pesticide Diazinon
20
(EtOh . P(S)·0-[6-Me-2-(l-methylethyl)]-4-pyrimidinyl
response
AChE
NTE
55
o
97
15
(+ or-)
50** 1000*
Malathion
Clinical
inhibition (%)
(MeO)2 . P(S)·S . CH . CO ·OEt CH2' CO· OEt Chlorpyrifos
100*
51
(EtOh . P(S) ·0-(3.5.6-CI3)-2-pyrimidinyl 50***
Methamidophos
50
MeO . P . (0) . (NH2) . SMe Methyl parathion
100**
85
12
500*
78
8
20*
75
0
100*
76
72
100
74
(MeOh . P(S)· 0-(4-N02 . Ph) Fenitrothion (MeOh . P(S) . 0-(3-Me-4-N02 . Ph) Cyanophos (MeOh . P(S)· 0-(4-CN· Ph) Salithion (MeO)· P(S)· 0-C6H4-2-CH20 Cyanofenphos EtO . (Ph) . P(S) . 0-(4-CN . Ph) EPN (racemic) EtO· (Ph)· P(S)· 0-(4-N02 . Ph) EPN (+)
303
75 83-90
50*
+
60
±
53-69* 50**
45
69-89*** 10-20 x 5 EPN(-)
50*
75
+
69* Omethoate (MeOh . P(O) . S . CH2 . C(O) . NH . CH3 Carbophenothion (Trithion)
± ±
75* 150-300*** 90 x 50***
>90
0
>90
0 0
(EtOh . P(S) . S-(4-CI . Ph) Parathion (EtOh . P(S)· 0-(4-N02 . Ph)
Range + 55 x 6***
Low
aFrom lohnson (1982a). All doses were administered orally in corn oil (or glycerol formal). NTE reponse was measured 1-2 days after dosing; for several
compounds. inhibition peaked at day 2. Where clinical response was negative. the dose is marked according to whether it was about LDso. *; 1-2 x LDSO. **; or more than twice the LDso. ***.
that, in its homologous series, only dichlorvos is not neuropathic at the LD50 dose (Table 47.2). Apart from the obvious correlation with clinical effects at both ends of the range of inhibition, there was a valuable warning of risk with chlorpyriphos dosed to hens at about LDso: At that time, this pesticide was regarded as nonneuropathic on the basis of approved tests involving only clinical and histopathological measures. The risk indication was vindicated by positive OPIDN seen in both hens given higher doses and in a failed human suicide attempt (Capodicasa et at., 1991; Moretto and Lotti, 1998).
47.3.3 APPLICATION OF NEUROPATHY TARGET ESTERASE STUDIES TO HUMAN RISK ASSESSMENT
47.3.3.1 In Vitro Comparison of Enzyme Targets When OP esters that exert biological effects in vivo are administered to animals, they are subjected to a variety of processes (absorption, metabolic activation and/or deactivation, distribution and/or excretion, etc.) before a certain amount of a proximal toxin is delivered to the ultimate targets-acetylcholinesterase
958
CHAPTER 47
Neuropathy Target Esterase Table 47.2 Relative Effects of Some OP Esters against Two Toxicity Targets of the Hen Assessed in vitro and in vivoa
Compound
In vitro
In vivo
AChe IsOINTE Iso
LDSOINTD
Dimethyl 2,2-dichlorovinyl phosphate
0.02
0.05
Diethyl 2,2-dichlorovinyl phosphate
0.16
0.17
Di-n-propyl 2,2-dichlorovinyl phosphate Di-n-pentyl 2,2-dichlorovinyl phosphate
2.6 32
Leptophos oxon
5 13
0.2
nd*
0.1
nd*
Mipafox
5.9
>1
DFP
0.9
Leptophos
b
Trichloronate oxonC Trichloronate
0.8 0.15
b
a From lohnson (I982a) who compiled in vivo data from several sources. NTD is that dose that caused severe ataxia in the majority of hens tested. bThionates free of oxons have negligible antiesterase activity in vitro. cTrichloronate oxon is ethyl 2,4,5-trichlorophenyl ethylphosphonate. *nd: not determined.
in the case of acute toxicity and NTE for delayed neuropathy. However, because the targets for both acute and delayed toxicity are associated with nervous tissue, it seems reasonable to propose that whatever percentage of a dose ultimately reached the nervous system in appropriately reactive form, then that amount would prefer to react with AChE or NTE according to the relative potencies demonstrable with these enzymes in vitro. Table 47.2 demonstrates a fair correlation between the relative potencies in vitro and in vivo of a variety of compounds not restricted to one homologous series. Use of the in vitro ratios, which can be determined easily and very early in a development program, may serve to guide synthetic chemists away from structures carrying neuropathic hazard. Although such in vitro/in vivo correlations have been obtained, the following assumptions built into the system need to be recognized:
1. For compounds that require metabolic activation in vivo, the actual inhibitor species must be identified and tested in vitro. 2. The inhibitor has equal access to both enzymes in the brain. 3. The rates of synthesis of fresh NTE and AChE are sufficiently slow not to affect the prognosis after an intoxication that causes massive inhibition of enzymes. 4. The rate of aging of NTE is rapid compared with the rates of synthesis of new enzymes. 5. The extent of spontaneous reactivation of inhibited enzymes in vivo is small. 6. The hen is an adequate model for humans. Assumption 1 is often possible for OP pesticides. There is good evidence to support assumptions 2-4. Assumption 5 appears true for NTE in our experience; inhibited NTE ages with a half-life of less than 1 h to a nonreactivatable form in most
cases tested and there is little spontaneous reactivation of NTE. However, assumption 5 is not true in all cases of AChE inhibition. For example, after poisoning with haloxon, the di-2chloroethylphosphorylated acetylcholinesterase has a half-life of approximately 22 min. Consequently, the acute toxicity of haloxon is less than might have been predicted and although the ratio of 150S for this compound is 0.01, it is, in fact, neuropathic at less than the LD50 dose (John son, 1982a). For a similar reason, methamidophos has caused OPIDN in humans, although in hens neuropathy is not caused by doses less than 6-8 times the unprotected LD50 given with massive treatment to prevent cholinergic death (John son, 1981; Lotti, 1992; Senanayake and Johnson, 1982aj. Assumption 6 is considered in more detail later in the chapter. The previous considerations led to the suggestion that any compound for which the ratio is more than 0.05 should be viewed with strong suspicion and that neuropathy may be caused in atropinized birds with single doses of compounds where the ratio is as low as 0.01. In spite of such a large difference in species type, it has been shown that the target enzymes in humans and hens are similar in their response to OP inhibitors; the 150S for both AChE and NTE differed between species by no more than a factor of 4 and often by less (Lotti and Johnson, 1978). However, these variations were not identical for each enzyme so that ratios of 150S diverged up to eightfold in some cases. For cases where the ratio is lower for humans, one might predict it would be comparatively more difficult to produce neuropathy in humans than in hens. However, in the case of trichlorphon (dichlorvos being the active inhibitory species), the in vitro ratio is 3 times higher for humans than for hens. There is increasing evidence that trichlorphon in single massive doses can produce neuropathy in humans (Johnson, 1981), whereas it requires more than one dose in hens. This may reflect the greater relative sensitivity
47.3 Toxological Applications
of the human target or it may indicate that a lower proportion of NTE in the human nervous system is required to be phosphorylated and aged to reach the initiation threshold.
47.3.3.2 Neuropathy Target Esterase and the Evaluation of Organophosphorus Ester-Induced Delayed Neuropathy in Humans Lotti (1992) listed 10 different pesticides that have been reported to cause OPIDN in humans, and isofenphos has been added to the list subsequently (Moretto and Lotti, 1998). According to the structure-activity relationships (SARs) listed previously, nine of these have molecular structures indicative of significant neuropathic potential, which has also been confirmed in laboratory studies involving assay of NTE in hen autopsy samples: Lotti noted that a case report naming omethoate (not significantly anti-NTE) as a causative agent had no sound evidence to identify the poison. A case of OPIDN believed to be due to a suicide attempt with a massive dose of parathion (De Jager et aI., 1981) was highly unusual in that it appeared to be a solitary event although acute poisoning with this pesticide is probably the most common of all reported OP intoxications. Parathion and paraoxon have very low anti-NTE potential, but the oxon derived from the impurity ethyl bis-(4-nitrophenyl) phosphorothioate is a potent inhibitor (John son, 1982b) and this could well have been the actual causative agent. The threshold of NTE inhibition that precipitates OPIDN in humans might be established if the following steps were taken more often following severe poisonings due to OPs: 1. The actual agent involved should be identified by chemical analysis. 2. The sample should be analyzed for major and minor OP constituents. 3. Lymphocyte NTE as well as erythrocyte AChE should be assayed in (serial) blood samples during treatment. 4. In the event of fatal poisoning, immediate autopsy samples should be obtained from brain and spinal cord and deep-frozen until AChE and NTE assays can be performed. A few such investigations have been perfonned with patients who did and others who did not suffer a neuropathic consequence of severe OP poisoning (Lotti et aI., 1981; Moretto and Lotti, 1998). The authors concluded that only substantial peak inhibition of NTE was associated with expression of frank clinical OPIDN. On the basis of the preceding limited infonnation, the application of the in vitro ratios of enzyme sensitivity appears to be an acceptable procedure in predictive toxicology and in clinical prognosis.
47.3.3.3 Neuropathy Target Esterase and the Assessment of Chiral Compounds One currently unresolved toxicological issue is the problem of chiral compounds (Johnson, 1987). The relative sensitivities of AChE and NTE in humans and hens to most tested OPs are
959
not greatly different, so that the measured relative susceptibility of hens to cholinergic or OPIDN effects can be transposed to humans. The actual dose effective in these species may differ because, in general, mammals have greater capacity for both bioactivation and detoxification of chemicals, but this does not alter the relative anti-esterase activities of these products. However, chiral OPs are a 50/50 mixture of two distinct chemicals that may be metabolized differently and to different extents in hens Johnson et al. (1991) and differently yet again in humans. Furthennore, the relative anti-AChE/anti-NTE potencies of the various metabolic products are unlikely to be all the same. Thus, in a worst case scenario, the predominant form of the anti-esterase compound(s) circulating in a dosed hen may dominantly affect AChE, whereas a different (antiNTE) compound(s) might predominate in humans. Although the problems of doing full toxicological evaluations of resolved isomers would be immense, the following comparatively easy investigations could be useful:
1. Perfonn assays of AChE and NTE in autopsy samples from a few mammals (rats) dosed with racemic compound and run in parallel with the full OPIDN evaluation in hens: Neither clinical nor histopathological examination is needed because the object is to decide whether the mammal produces markedly different relative effects on the enzyme targets than does the hen. 2. Run assays in vitro for the potency against AChE and NTE of resolved isomers of whatever anti-esterase compounds have been identified during routine metabolism studies of the unresolved compound: These assays are straightforward and require only a very few milligrams of material, whereas whole-animal studies might require many grams. Taken together, these two limited studies should indicate whether or not the hen study with racemic compound is indicative of the human situation.
47.3.3.4 Neuropathy Target Esterase and the Assessment of Effects of Long-Term, Low-Level Exposure A significant concern is whether short-term, high-dose experiments in hens are appropriate to assess the possible neuropathic hazard of long-tenn human exposure to relatively lower levels of OP pesticides, with the possibility of cumulative effects. The quantitative data on inhibition of NTE that emerge after shorttenn tests, even when no other effects are seen, go some way in providing a useful assessment. Thus, whereas a single dose of 50 mg/kg of mono-o-cresyl diphenyl phosphate (MOCP) caused OPIDN in hens, a total of 175 mg/kg dosed daily over 10 weeks did not; monitoring brain and spinal cord NTE activity over this period showed that inhibition reached a stable equilibrium after 1-2 weeks at a value (45-60%) below the threshold (70-90%) required to initiate OPIDN (Lotti and Johnson, 1980a). Current Organization for Economic Cooperation and Development (OECD) Guideline 419 restricts multidose
960
CHAPTER 47
Neuropathy Target Esterase
do not have the appropriate inhibition characteristics to qualify as a target for OPIDN. Thus, when total hydrolytic activity in the absence of inhibitors is dubbed A, the paraoxon-resistant activity is dubbed B, and the residual activity resistant to both paraoxon and mipafox (either together or in sequence) is C, then the activity of NTE is determined as B - C and specificity 47.4 NATURE AND PROPERTIES OF as the ratio (B - C) / B. In an assay of hen brain PV hydrolases, NEUROPATHY TARGET ESTERASE after preincubation of the tissue with paraoxon, the substrate specificity is about 65%, which allows quite accurate determi47.4.1 BIOCHEMICAL STUDIES nation of NTE by the B - C calculation even when overall activity is low as in some autopsy samples from birds dosed Until quite recently, all studies on NTE relied on detection of with neuropathic compounds. its esterase activity or labeling by eH]DFP. Using these methExtensive structure-activity studies for both substrates and ods, it was shown that, in the adult chicken, the highest specific inhibitors in vitro were reported by Johnson (1975b, 1988), activities of NTE were found in brain whereas spinal cord and Thomas et al. (1990), Wu and Casida (1992), and Borhan et at. sciatic nerve contained substantially less (Johnson, 1982a). In (1995); these identified substrates more sensitive than PV (catdissected areas of human brain, NTE varied by a factor of 2 with alytic center activity up to two- to threefold greater) but all cerebral cortex the highest and cerebellum the least (Lotti and were less than 50% specific. For routine investigations, PV is Johnson, 1980b). Relatively high levels ofNTE were present in accepted as widely tested and approved but some of these alterseveral nonneural chicken tissues, including intestine, spleen, natives may be useful for specific studies, such as for a partly and thymus (Johnson, 1982a), and extremely high levels have purified enzyme free of interfering esterases or when tissue been found in bovine adrenal medulla (Sogorb et aI., 1994). activity is very low or for kinetic investigations that require subCultured bovine chromaffin cells and human neuroblastoma strate to have sufficiently low Km to ensure complete quench of cell lines have been shown to have substantial NTE activity and any progessive inhibition at the instant of the addition of subhave been suggested as in vitro systems for assessment of neustrate. It is a fact that the Km of PV for NTE is high (about ropathic OPs (Sogorb et aI., 1997; Veronesi et aI., 1997). 10 mM compared with about 0.1 mM for phenyl phenylacetate, Biochemical fractionation of chicken brain homogenates which was used in early studies but which lacked both sensitivshowed that NTE is enriched in microsomal membrane fractions (Richardson et aI., 1979). [3H]DFP-Iabeling indicated that ity and specificity; Johnson, 1982a). The aging reaction (cf. Fig. 47.1b) ofDFP-inhibited NTE is NTE comprised less than 0.1 % of total microsomal protein very rapid with a half-life of a few minutes and the aged iso(Williams and Johnson, 1981). NTE is an integral membrane propyl group is quantitatively transferred to a covalent acceptor protein as indicated by its requirement for detergent for solubisite, dubbed site Z, within NTE itself; this contrasts with a much lization. The type and concentration of detergent were shown to be important for the maintenance of NTE activity (Davis slower rate of aging for DFP-inhibited cholinesterases in which and Richardson, 1987); fractionation of the solubilized mate- the aged isopropyl group is liberated into free solution (Clothier rial generally led to substantial loss in NTE activity, which and Johnson, 1979). Yoshida et al. (1995) have investigated the could be partially ameliorated by the addition of phospho- reaction of chicken brain NTE with tritiated octyl cyclic salilipids (Pope and Padilla, 1989a). On sodium dodecyl sulfate- genin phosphonate; these authors report that only about 15% of polyacrylamide gel electrophoresis (SDS-PAGE), eH]DFP- the aged saligenin group is transferred to site Z. The identity of labeled NTE runs as a 155-kDa polypeptide (Williams and site Z in NTE is unknown but clearly it is a residue that is in Johnson, 1981) whereas, on gel filtration, detergent-solubilized close proximity to the active site in the native folded protein; NTE migrates as a complex with an apparent molecular weight there is evidence from proteolysis of SDS-solubilized preparagreater than 850 kDa (Pope and Padilla, 1989b; Thomas et aI., tions of eH]DFP-labeled chicken brain microsomes that site Z 1990). lies within 150 residues of the active site serine residue (Glynn et aI., 1993). Although the rapidity of aging of covalently bound OP and 47.4.2 EN ZYMOLOGY OF NEUROPATHY the intramolecular transfer of alkyl groups appears to be a TARGET ESTERASE unique feature of NTE, it has been concluded that the generEnzymically, NTE behaves as a typical B-esterase (i.e., it is sen- ation of a negatively charged species at the active site, rather sitive to OP). No physiological substrate has been identified, but than the modification of site Z, is the critical event in initiation it rapidly hydrolyzes certain hydrophobic artificial substrates of of OPIDN (John son, 1990). In the case of phosphoramidates which phenyl valerate (PV) has the best combination of sensi- such as mipafox, it has been proposed that aging involves loss tivity and specificity under appropriate assay conditions. Even of a proton, rather than an alkyl/aryl group, to leave an elecwith PV, a differential assay is necessary to distinguish NTE tronegative species attached to the active-site serine (Richardfrom other esterases that can hydrolyze the same substrate but son, 1995).
tests to 28 days unless special exposure conditions pertain and suggests that negative results on the biochemical, histopathological, and behavioral endpoints indicate that further testing of the compound is not required (OECD, 1995).
47.4 Nature and Properties of Neuropathy Target Esterase
47.4.3 ISOLATION AND IMMUNOHISTOCHEMICAL LOCALIZATION The low abundance and apparent requirement for membrane lipid to maintain NTE activity impeded its isolation for several years. A fraction substantially enriched in eHJDFP-Iabeled NTE, but far from homogeneous, was isolated from chicken brain (Rueffer-Turner et aI., 1992). An apparently homogeneous NTE preparation from phospholipase A2-solubilized embryonic chicken brain had a specific activity about half that in the initial crude solubilized extract (Mackay et aI., 1996). A breakthrough was finally achieved by the synthesis of a novel reagent, S9B [l-(saligenin cyclic phosphoro)9-biotinyldiaminononaneJ, for affinity purification of NTE (Glynn et al., 1994). S9B reacted rapidly and specifically with NTE in brain microsomes and resulted in the covalent attachment, via a long alkyl spacer, of a biotin molecule to the active-site serine residue. Microsomal proteins, quantitatively solubilized by boiling in dilute SDS, were then subjected to affinity chromatography with avidin-Sepharose, which binds biotinylated polypeptides. S9B-Iabeled NTE was eluted from the avidin by boiling in SDS. Two polypeptides (carboxylases) with endogenous covalent biotin prosthetic groups that co-eluted from avidin-Sepharose with NTE were removed by subsequent preparative electrophoresis (Glynn et aI., 1994). Isolated chicken NTE was digested with endoproteinase Glu-C and the resulting peptide fragments resolved by SDSPAGE. The N-terminal amino acid sequence of one of these fragments provided sufficient information to synthesize an 11residue peptide, which was used to raise a rabbit antiserum to NTE (Glynn et al., 1998). An immunohistochemical survey of the chicken nervous system using this antiserum showed that NTE was present in essentially all neurons but was absent from glia. NTE immunostaining could not be detected in normal sciatic nerve but accumulated at the constriction site 8 h after nerve ligation, indicating that NTE undergoes fast axonal transport. NTE immunostaining filled neuronal cell bodies (except the nucleus) and sometimes extended into the proximal axon; this pattern, taken together with the biochemical data on NTE in microsomal fractions, indicated that NTE is probably associated with the endoplasmic reticulum. These immunostaining characteristics were not detectably altered in chickens 1 or 3 days after treatment with a neuropathic OP, suggesting that OPmodified NTE was neither grossly redistributed nor degraded faster than native NTE (Glynn et aI., 1998).
47.4.4 MOLECULAR CLONING OF HUMAN NEUROPATHY TARGET ESTERASE: IMPLICATIONS FOR STRUCTURE AND FUNCTION The N-terminal sequence of an endoproteinase Glu-C fragment of S9B-Iabeled pig brain NTE was found to be very similar to a human-expressed sequence tag cDNA; the latter was used to
961
3 4
NTE (1327 aa) Regulatory domain
NE T (489 aa + tags)
I E ffector domain
T7
His 6
Figure 47.2 Predicted secondary structure of NTE and NEST. NTE is shown as a linear polypeptide of 1327 amino acids with two major functional domains: an N-tenninal regulatory domain, which contains regions with some similarity to cyclic AMP-binding proteins, and a C-tenninal effector domain (shown in gray), which contains the esterase activity. Four transmembrane segments predicted by TM-pred analysis are shown as thick vertical bars. The active-site serine (Ser 966) lies at the center of putative transmembrane segment 4. NTE residues 727-1216 have been cloned into a pET vector and expressed in E. coli with a short N-tenninal (T7) tag sequence and a C-tenninal His-6 tag; this construct, dubbed NEST, has all the OP-sensitive phenyl valerate hydrolase activity of full-length NTE (Atkins and Glynn, 2000).
initiate screening of human brain cDNA libraries from which a full-length NTE cDNA clone was finally isolated (Lush et aI., 1998). The NTE cDNA clone D16 encoded a polypeptide of 1327 amino acids, and analysis of this sequence with the transmembrane prediction (TM-pred) program indicated the presence offour potential transmembrane segments (see Fig. 47.2). Biochemical experiments indicated that the active-site serine residue labeled by S9B lay between residues 955 and 1033, and attention was drawn to Ser 966, which lay in the motif Gly-Xxx-Ser-Xxx-Gly, common to all serine hydrolases (Lush et aI., 1998). Ser 966 has subsequently been confirmed as the active-site residue by eHJDFP-Iabeling and protease digestion of a recombinant form of NTE (Atkins and Glynn, 2000). Interestingly, Ser 966 is at the center of the fourth predicted transmembrane segment in NTE (Lush et aI., 1998). Whether the segment of NTE containing the active-site serine is actually located within a membrane lipid bilayer is currently under investigation in this laboratory. If this proves to be the case, then, in order to achieve hydrolysis of the acyl enzyme intermediate (Fig. 47.la), the active-site serine of NTE would have to line an aqueous transmembrane pore. In turn, this structure would suggest NTE's physiological function may not be due simply to its esterase activity. Alternatively, this secondary structure prediction may simply indicate that the active-site serine lies in a hydrophobic helical segment, akin to the arrangement in some lipases (Derewenda and Sharp, 1993). Whichever possibility is correct, it is clear that placing a negatively charged group in this location-the result of OP-mediated aging-would be expected to have a drastic effect on the structure of NTE. Human NTE is highly homologous (41% identical) to a Drosophila neuronal protein called swiss cheese (SWS; Lush et aI., 1998). The sws mutation results in glial hyperwrapping of neurons, which, in turn, leads to apoptotic death of both cell types; the name "swiss cheese" derives from the vacuolated appearance of the mutant brains (Kret-
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Neuropathy Target Esterase
959 978 966 1399 816 47
NTE human sws insect YOL4 caeel YMF9-yeast mtcy20bl1 YCHK ecoli
NTE human sws insect YOL4 caeel YMF9- yeast mtcy20bll YCHK ecoli
NTE human sws insect YOL4 caeel YMF9-yeast mtcy20b11 YCHK ecoli
*
*
*, *
*
*
ERSASRTKQREEWAKSMTSVLEPVLI~ ERNITTVTQK' EWSKKMTKWFLQI~ ~I TPDDVVVETR' SWFNGMSSLWRK AH YDLVPIYGRVKKFAGRISSIWRM T ~ :i!CGt-mJ~ATADAY!i1YEYFIRHN . ... P SDl!.AF~ ~I~'9~1~~,V~~1~g~D . RLSALEDWVTSFSYW . .. DVLR b~SWQR
;~~~~IE
;~~~~~~I~i~~:~~I'
TSYTT I SAMFT HE Q RTLTLLE RGLVR RVFNQYREIMPETE GGLLR
*
~IF~i;l·;~I~:~I~lsl
Q CNS • S FCIS
KEFRCVSV~
SRRFAAVA
TD QEI SF ST MRV S PA LARRPVV R LSTGRELWFTE H
1009 1028 1016 1449 862 93
1059 1078 1 066 1499 912 143
1109 1128 1116 1547 959 191
NTE human sws-insect YOL4 caeel YMF9-yeast mtcy20bl1 YCHK ecoli Figure 47.3 Alignment of amino acid sequences in the highly conserved C-terminal region of NTE with homologous proteins from various species. Amino acids (aa) 910-1109 ofNTE (complete sequence = 1327 aa) are aligned with homologous regions from SWS (Drosophila; 1425 aa), YOL4 (Caenorhabditis elegans; 1351 aa), YMF9 (Saccharomyces cerevisiae; 1679 aa), MTCY20B 11.l4c (Mycobacterium tuberculosis; 1048 aa), and YCHK (E. coli; 314 aa). Residues identical in at least four of the proteins are shown white on black. The positions of the active-site serine (Ser 966) and of conserved His, Asp, and Glu residues are shown black on grey and are indicated by an arrow.
zschmar et aI., 1997). It has been suggested that the SWS protein is involved in a cell-signaling pathway, and attention has been drawn to the similarity between an N-terrninal domain of SWS (also present in NTE) that resembles the cyclic AMP-binding regulatory subunit of protein kinase A (Kretzschmar et aI., 1997). In addition to sharing sequence homology, biochemical assays have shown that NTE-like phenyl valerate hydrolase activity is present in wild-type Drosophila but absent from sws mutants (Moser et aI., 2000). is the Drosophila homolog of NTE and, by analogy, NTE may mediate cell signaling in the developing vertebrate brain. In situ hybridization experiments on mouse embryos show that NTE mRNA is expressed in neurons from their earliest appearance in the developing nervous system (Moser et aI., 2000). Sequence database searches revealed that NTE is not related to any known serine esterases or proteases but, in addition to its close similarity with Drosophila SWS, it shares homology with a number of polypeptides predicted from the sequencing of genomes of bacteria, yeast, and nematodes (Lush et aI., 1998). In particular, a 200-amino-acid domain, cor-
responding to NTE residues 910-1109, is highly conserved (29% identity between human NTE and the Escherichia coli homolog YCHK) and, notably, all the homologous proteins contain a serine residue in the same position as Ser 966 of NTE. In addition, a completely conserved His and several acidic (Asp and Glu) residues are found within this domain, which, together with the serine, could comprise a catalytic triad as found in conventional serine hydrolases (Fig. 47.3; Lush et aI., 1998). Thus, NTE is a member of a novel protein family that appears to comprise potential serine hydrolases. However, although recombinant fragments of both NTE (residues 727-1216) and SWS (residues 746-1235) show substantial PV hydrolase activity when expressed in the pET vectorf£. coli system, we have been unable to detect PV hydrolase activity for analogous portions of the bacterial or yeast homologues expressed in the same system (Fig. 47.3; J. Atkins, Y. Li, and P. Glynn, unpublished). Whether this result simply reflects the differing substrate specificity of these potential esterases or, alternatively, indicates that this domain has been conserved for a nonesteratic function is currently under investigation.
47.5 Summary Table 47.3 Species Susceptibility to Single-Dose OPIDN and Brain NTE Levels Brain NTE Species
Susceptibility
(nmoIlminlg)
Human
+ + + + +
2400
Chicken Cat Pig Sheep
2400 2200 2000 1850
Marmoset
2000
Quail
2000
Guinea pig
1000
Rat
950
Mouse
700
Rabbit
600
Gerbil
400
Data from 10hnson (l982a) and Read and Glynn (unpublished).
47.4.5 ROLE OF NEUROPATHY TARGET ESTERASE IN ORGANOPHOSPHORUS ESTER-INDUCED DELAYED NEUROPATHY: A TOXIC GAIN OF FUNCTION? The close similarity between NTE and Drosophila SWS suggests that NTE may have an important function during brain development through involvement in a cell-signaling pathway; a number of experimental approaches are currently investigating this possibility. The role of NTE in OPIDN is a rather different issue. It has long been clear that prolonged inhibition ofNTE's esterase activity has no obvious adverse effects in the adult chicken (John son, 1990). Thus, it may be that NTE no longer has a vital role in adult animals but rather it acquires a novel toxic function on modification by a neuropathic OP. Data on species variation in susceptibility to OPIDN and in brain levels of NTE are shown in Table 47.3. Susceptible species are generally larger and hence have longer axons than resistant species. This is particularly apparent, for instance, when adult chickens are compared with quailsboth have relatively high levels of brain NTE activity and yet the latter birds are resistant. However, an additional consistent observation is that animals with relatively low levels of NTE «1000 nmollminlg) are relatively resistant to OPIDN. Furthermore, we have found that in mice (a particularly resistant species) NTE appears to turn over faster (tl/2 = 2 days; D. J. Read and P. Glynn, unpublished) than in chickens (tl/2 = 4-5 days; Johnson, 1974; Meredith and Johnson, 1988). This would be consistent with a mechanism whereby a certain threshold level of OP-modified NTE must be achieved and then maintained for a finite period in order to initiate OPIDN. Experiments are now underway attempting to generate transgenic mice expressing very high levels of NTE to determine whether these animals show a heightened susceptibility to OPIDN.
963
47.5 SUMMARY Neuropathy target esterase is a high-molecular-weight integral membrane protein present in neurons. It was defined originally by its uniquely selective reactivity with those organophosphorus esters that induce delayed onset neuropathy in humans and various test animals, of which the adult chicken is the most convenient and reproducible in showing the effect. Initiation of OPIDN by a pesticide requires that the compound be in its oxon (P=O) form, which can react covalently with NTE to modify the protein in two steps: (1) organophosphorylation of thc scrinc residue at the catalytic centre of the enzyme (which inhibits the esterase activity) and (2) a rapid intramolecular rearrangement of the bound OP to generate a negatively charged group attached to the serine. Certain classes of NTE inhibitors (carbamates, sulfonyl fluorides, and phosphinates) can inhibit catalytic activity by progressive covalent reaction at the serine but are structurally incapable of undergoing step 2. Doses of these compounds do not cause neuropathy but act as prophylactic compounds against OPIDN by virtue of their ability to block the target site in an apparently innocuous fashion. It appears that the ability of NTE to catalyze hydrolysis of esters is redundant in the adult and it is possible that the initiation of OPIDN involves a toxic gain of function initiated by generation of the negative charge on the NTE molecule very soon after ingestion of the pesticide. Assay of the degree of inhibition of NTE in autopsy samples of brain and spinal cord from dosed animals provides quantifiable biochemical data as a valuable adjunct to clinical and morphological assessment of the effects in tests of the neuropathic potential of OPs. The NTEs of hen and human brain are similar in sensitivity to many OPs. Studies in vitro have provided structure-activity relationships, and comparison of the sensitivities of NTE and acetylcholinesterase enable predictions of the relative acute/neuropathic potential of candidate pesticides, including chiral compounds, which present peculiar problems to regulatory toxicologists. Assays of NTE in accessible lymphocytes from a few poisoned patients suggest that there is predictive value in the procedure and that OPIDN in humans, as in hens, requires a high level of inhibition to precipitate frank clinical neuropathy. Assessment of long-term, low-level exposure to OPs indicates that inhibition of NTE is not totally cumulative and may reach an equilibrium level below that needed to initiate OPIDN. As an enzyme, NTE behaves as a typical B-esterase that rapidly hydrolyzes phenyl valerate and related esters, but no physiological substrate has been identified; related enzymes may be involved in the mechanism of promotion of OPIDN by some inhibitors that are unable to initiate the process. Although attempts to isolate catalytically active enzyme failed, an affinity-Iabeled form has been purified and antibodies raised: Immunohistochemical study shows NTE confined largely to the endoplasmic reticulum of neuronal cell bodies and to be transported rapidly down axons. The primary sequence of NTE has been determined from its cDNA and shown to be unrelated to any known serine es-
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Neuropathy Target Esterase
terases or proteases. NTE is the vertebrate homolog of the swiss cheese protein (SWS) of Drosophila: SWS resides in neurons and appears to be involved in signaling between these cells and glia in the developing fly brain. In embryonic mice, NTE is expressed in neurons from their earliest appearance in the nervous system where it could function in analogous fashion to SWS. Conceivably, the esterase function of NTE is required during neural development but not in later life. The active-site serine of NTE is located in a C-terminal domain that is conserved in proteins expressed by organisms from bacteria to humans; however, unlike NTE and SWS, homologs from unicellular organisms appear to lack phenyl valerate hydrolase activity. The active-site serine ofNTE lies at the center of a hydrophobic segment of the polypeptide chain; generation of a negative charge by a bound neuropathic OP in this location would probably have drastic effects on the structure and function of NTE.
REFERENCES Aldridge, W. N., and Reiner, E. (1972). "Enzyme Inhibitors as Substrates." ElsevierlNorth-Holland, Amsterdam. Atkins, J., and Glynn, P. (2000). Membrane association and critical residues in the catalytic domain of human neuropathy target esterase. 1. BioI. Chem. 275,24477-24483. Barril, J., and Vilanova, E. (1997). Reversible inhibition can profoundly mislead studies on progressive inhibition of enzymes: The interaction of paraoxon with soluble neuropathy target esterase. Chem.-Biol. Interact. 108, 19-25. Barril, J., Estevez, J., Escudero, M. A., Cespedes, M. v., Niguez, N., Sorgob, A., Monroy, A., and Vilanova, E. (1999). Peripheral nerve soluble esterases are spontaneously reactivated after inhibition by paraoxon: Implications for a new definition ofNTE. Chem.-Biol. Interact. 119-120,541-550. Borhan, B., Ko, Y., Wilson, B. W., Kurth, M. J., and Hammock, B. D. (1995). Development of surrogate substrates for neuropathy target esterase. Biochim. Biophys. Acta 1250,171-182. Capodicasa, E., ScapeIIato, M. L., Moretto, A., Caroldi, S., and Lotti, M. (1991). Chlorpyrifos-induced delayed polyneuropathy. Arch. Toxicol. 65, 150-155. Cespedes, M. v., Escudero, M. A., Barril, J., Sogorb, M. A., Vicedo, J. L., and Vilanova, E. (1997). Discrimination of carboxylesterases of chicken neural tissue by inhibition with neuropathic and non-neuropathic organophosphorus compounds and a neuropathy promoter. Chem.-Biol. Interact. 106, 191-200. Clothier, B., and Johnson, M. K. (1979). Rapid aging of neurotoxic esterase after inhibition by di-isopropyl phosphorofluoridate. Biochem. 1. 177,549558. Davis, C. S., and Richardson, R. J. (1987). Neurotoxic esterase: Characterisation of the solubilised enzyme and the conditions for its solubilisation from chicken brain microsomal membranes with ionic, zwitterionic or non-ionic detergents. Biochem. Pharmacol. 36, 1393-1399. De Jager, A. E. J., Van Weerden, T. W., Houthhoff, H. J., and de Monchy, J. G. R. (1981). Polyneuropathy after massive exposure to parathion. Neurology 31, 603-605. Derewenda, Z. S., and Sharp, A. M. (1993). News from the interface: The molecular structures of triacylglceride lipases. Trends Biochem. Sci. 18,2025. Escudero, M. A., and Vilanova, E. (1997). Purification and characterisation of naturally soluble neuropathy target esterase from chicken sciatic nerve by HPLC and Western blot. 1. Neurochem. 69,1975-1982. European Centre for Ecotoxicology and Toxicology of Chemicals (ECETOC) (1998). "Organophosphorus Pesticides and Long-Term Effects on the Nervous System." Technical Report 75, European Centre for Ecotoxicology and Toxicology of Chemicals, Brussels.
Glynn, P., Holton, J. L., Nolan, C. c., Read, D. J., Brown, L., Hubbard, A., and Cavanagh, J. B. (1998). Neuropathy target esterase: Immunolocalisation to neuronal cell bodies and axons. Neuroscience 83, 295-302. Glynn, P., Read, D. J., Guo, R., Wylie, S., and Johnson, M. K. (1994). Synthesis and characterisation of a biotinylated organophosphorus ester for detection and affinity purification of a brain serine esterase: Neuropathy target esterase. Biochem. 1. 301, 551-556. Glynn, P., Rueffer-Turner, M., Read, D. J., Wylie, S., and Johnson, M. K. (1993). Molecular characterisation of neuropathy target esterase: Proteolysis of the [3HJDFP-Iabelled polypeptide. Chem. BioI. Interact. 87,361-367. Johnson, M. K. (l969a). A phosphorylation site in brain and the delayed neurotoxic effect of some organophosphorus compounds. Biochem. 1. 111, 487-495. Johnson, M. K. (I 969b). The delayed neurotoxic effect of some organophosphorus compounds: Identification of the phosphorylation site as an esterase. Biochem.l. 114,711-714. Johnson, M. K. (1970). Organophosphorus and other inhibitors of "neurotoxic esterase" and the development of delayed neurotoxicity in hens. Biochem. 1. 120, 523-531. Johnson, M. K. (1974). The primary biochemical lesion leading to the delayed neurotoxic effects of some organophosphorus esters. 1. Neurochem. 23,785-789. Johnson, M. K. (1975a). Organophosphorus esters causing delayed neurotoxic effects: Mechanism of action and structure/activity relationships. Arch. Toxicol. 34, 259-288. Johnson, M. K. (l975b). Structure-activity relationships for substrates and inhibitors of hen brain neurotoxic esterase. Biochem. Pharmacol. 24,797805. Johnson, M. K. (1978). The anomalous behaviour of some dimethyl-phosphates in the biochemical test for delayed neurotoxicity potential. Arch. Toxicol. 41, 107-110. Johnson, M. K. (1981). Delayed neurotoxicity-do trichlorphon and/or dichlorvos cause delayed neuropathy in man or in test animals? Acta Pharmacal. Toxicol. 49, 87-98. Johnson, M. K. (l982a). The target for initiation of delayed neurotoxicity by organophosphorus esters: Biochemical studies and toxicological applications. In "Reviews in Biochemical Toxicology" (E. Hodgson, J. R. Bend, and R. M. Philpot, eds.), Vol. 4, pp. 141-212. Elsevier, New York. Johnson, M. K. (l982b). Check your paraoxon and parathion for neurotoxic impurities. Veterin. and Hum. Toxicol. 24, 220a. Johnson, M. K. (1984). Delayed neurotoxicity tests of organophosphorus esters: A proposed protocol integrating neuropathy target esterase (NTE) assays with behaviour and histopathology tests to obtain more information more quickly from fewer animals. In "Proceedings of the International Conference on Environmental Hazards of Agrochemicals in Developing Countries, Alexandria, Egypt, November 8-12, 1983" (A. H. EI-Sebae, ed.), pp. 474493. Univ. of Alexandria, Alexandria, Egypt. Johnson, M. K. (1987). The importance of chirality in influencing both acute and delayed neuropathic toxicity of organophosphorus esters. Toxicol. Environ. Chem. 14,321-335. Johnson, M. K. (1988). Sensitivity and selectivity of compounds interacting with neuropathy target esterase: Further structure/activity studies. Biochem. Pharmacol. 37,4095-4104. Johnson, M. K. (1990). Organophosphates and delayed neuropathy-Is NTE alive and well? Toxicol. Appl. Pharmacol. 102, 385-399. Johnson, M. K., and Lauwerys, R. R. (1969). Protection by some carbamates against the delayed neurotoxic effect of di-isopropyl phosphorofluoridate. Nature (London) 222, 1066-1067. Johnson, M. K., and Read, D. J. (1987). The influence of chirality on the delayed neuropathic potential of some organophosphorus esters: Neuropathic and prophylactic effects of stereoisomers of ethyl phenylphosphonic acid (EPN oxon and EPN) correlate with quantities of aged and unaged neuropathy target esterase in vivo. Toxicol. Appl. Pharmacol. 90, 103-115. Johnson, M. K., Vilanova, E., and Read, D. J. (1991). Anomalous biochemical responses in tests of the delayed neuropathic potential of methamidophos (O,S-dimethyl phosphorothioamidate), its resolved isomers and of some higher O-alkyl homologues. Arch. Toxicol. 65, 618-624.
References
Krejci, E., Duval, N., Chatonnet, A., Vincens, P., and Massoulie, 1. (1991). Cholinesterase-like domains in enzymes and structural proteins: Functional and evolutionary relationships and identification of a catalytically essential aspartic acid. Proc. Nat!. Acad. Sci. U.S.A. 88, 6647-6651. Kretzschmar, D., Hasan, G., Sharma, S., Heisenberg, M., and Benzer, S. (1997). The swiss cheese mutant causes glial hyperwrapping of and brain degeneration in Drosophila. 1. Neuroscience 17,7425-7432. Lotti, M. (1992). The pathogenesis of organophosphate polyneuropathy. Crit. Rev. Toxico!. 21,465-487. Lotti, M., and lohnson, M. K. (1978). Neurotoxicity of organophosphorus pesticides: Predictions can be based on in vitro studies with hen and human enzymes. Arch. Toxico!. 41,215-221. Lotti, M., and lohnson, M. K. (1980a). Repeated small doses of a neurotoxic organophosphate: Monitoring of neurotoxic esterase in brain and spinal cord. Arch. Toxicol. 45, 263-271. Lotti, M., and lohnson, M. K. (1980b). Neurotoxic esterase in human nervous tissue. 1. Neurochem. 34,747-749. Lotti, M., and Moretto, A. (1999). Promotion of organophosphate induced polyneuropathy by certain esterase inhibitors. Chem.-Biol. Interact. To appear. Lotti, M., Ferrara, S. D., Caroldi, S., and Sinigaglia, F. (1981). Enzyme studies with human and hen autopsy tissue suggest omethoate does not cause delayed neuropathy in man. Arch. Toxicol. 48, 265-270. Lush, M. 1., Li, Y., Read, D. 1., Willis, A. C., and Glynn, P. (1998). Neuropathy target esterase and a homologous Drosophila neurodegeneration mutant protein contain a domain conserved from bacteria to man. Biochem. 1. 332, 1-4. Mackay, C. E., Hammock, B. D., and Wilson, B. W. (1996). Identification and isolation of a 155kDa protein with neuropathy target esterase activity. Fundam. App!. Toxico!. 30,23-30. Meredith, C., and lohnson, M. K. (1988). Neuropathy target esterase: Rates of turnover in vivo following covalent inhibition with phenyl di-npentylphosphinate. 1. Neurochem. 51, 1097-1101. Milatovic, D., Moretto, A., Osman, K. A., and Lotti, M. (1997). Phenyl valerate esterases other than neuropathy target esterase and the promotion of organophosphate polyneuropathy. Chem. Res. Toxico!. 10, 1045-1048. Moretto, A., and Lotti, M. (1998). Poisoning by organophosphorus insecticides and sensory neuropathy. 1. Neuro!' Neurosurg. Psychiat. 64, 463-468. Moser, M., Stempl, T., Li, Y., Glynn, P., Buttner, R., and Kretzschmar, D. (2000). Molecular cloning of the mouse NTE/SWS gene. Mech. Dev. 90, 279-282. Organization for Economic Co-operation and Development (OECD) (1995). "Delayed Neurotoxicity of Organophosphorus Substances Following Acute Exposure" and "Delayed Neurotoxicity of Organophosphorus Substances: 28-Day Repeated Dose Study." Guidelines for Testing of Chemicals 418 and 419, Environmental Health and Safety Division, Organization for Economic Co-operation and Development, Paris. Pope, C. N., and Padilla, S. S. (1989a). Modulation of neurotoxic esterase activity in vitro by phospholipids. Toxico!. App!. Pharmaca!' 97,272-278. Pope, C. N., and Padilla, S. S. (1989b). Chromatographic characterisation of neurotoxic esterase. Biochem. Pharmaco!' 38, 181-188. Poulsen, E., and Aldridge, W. N. (1964). Studies on esterases in the chicken nervous system. Biochem. 1. 90, 182-189.
965
Richardson, R. 1. (1995). Assessment of the neurotoxic potential of chlorpyrifos relative to other organophosphorus compounds: A critical review of the literature. 1. Toxicol. Environ. Health 44, 135-165. Richardson, R. 1., Davis, C. S., and lohnson, M. K. (1979). Subcellular distribution of marker enzymes and of neurotoxic esterase in adult hen brain. 1. Neurochem. 32,607-615. Rueffer-Turner, M. E., Read, D. 1., and lohnson, M. K. (1992). Purification of neuropathy target esterase from avian brain after prelabelling with [3Hjdiisopropyl phosphorofluoridate. 1. Neurochem. 58, 135-141. Senanayake, N., and lohnson, M. K. (1982). Acute polyneuropathy after poisoning by a new organophosphorus insecticide. N. Engl. 1. Med. 306, 155156. Smith, M. 1., Elvove, E., and Frazier, W. H. (1930). The pharmacological action of certain phenol esters with special reference to the etiology of the socalled ginger paralysis. Public Health Rep. 45, 2509-2524. Sogorb, M. A., Bas, S., Gutierrez, L. M., Vilanova, E., and Viniegra, S. (1997). Bovine chromaffin cells as an in vitro model for the study of non-cholinergic toxic effects of organophosphorus compounds. Arch. Toxico!. Supp!. 19, 347-355. Sogorb, M. A., Viniegra, S., Reig, 1. A., and Vilanova, E. (1994). Partial characterisation of neuropathy target esterase and related phenyl valerate esterases from bovine adrenal medulla. 1. Biochem. Toxico!. 9, 145-152. Soliman, S. A., Curley, A., Farmer, 1., and Novak, R. (1986). In vivo inhibition of chicken brain acetylcholinesterase and neurotoxic esterase in relation to the delayed neurotoxicity of leptophos and cyanophenphos. 1. Environ. Patho!. Toxico!. Oncol. 7, 2[[-224. Taylor, P. (1992). Impact of recombinant DNA technology and protein structure determination on past and future studies on acetylcholinesterase. In "Multidisciplinary Approaches to Cholinesterase Functions" (A. Shafferman and B. Velan, eds.), pp. 1-15. Plenum, New York. Thomas, T. c., Szekacs, A., Rojas, S., Hammock, B. D., Wilson, B. W., and MacNamee, M. G. (1990). Characterisation of neuropathy target esterase using trifluoromethyl ketones. Biochem. Pharmaco!' 40, 2587-2596. D.S. Environmental Protection Agency (EPA) (1991). "Pesticide Assessment Guidelines, Subdivision F; Hazard Evaluation: Human and Domestic Animals: Addendum 10, Neurotoxicity," Series 81, 82, and 83, pp. 3-12. EPA 540/09-91-123, PB 91-154617, Health Effects Division, Office of Pesticide Programs, D.S. Environmental Protection Agency, Washington, DC. Veronesi, B., Ehrich, M., Blusztajn, 1. K., Oortgiesen, M., and Durham, H. (1997). Cell culture models of interspecies selectivity to organophosphorus insecticides. Neurotoxicology 18,283-297. Vilanova, E., Escudero, M. A., and Barril, 1. (1999). NTE soluble isoforms: New perspectives for targets of neuropathy inducers and promoters. Chem.Bio!. Interact. 119-120, 525-540. Williams, D. G., and lohnson, M. K. (1981). Gel electrophoretic identification of hen brain neurotoxic esterase labelled with tritiated di-isopropyl phosphorofluoridate. Biochem. 1. 199,323-333. Wu, S.- Y., and Casida, 1. E. (1992). Neuropathy target esterase inhibitors: 2-Alkyl-, 2-alkoxy-, and 2-(aryloxy)-4H -1,3,2-benzodioxaphosphorin 2-oxides. Chem. Res. Toxico!. 5, 680--684. Yoshida, M., Tomizawa, M., Wu, S.-Y., Quistad, G. B., and Casida, 1. E. (1995). Neuropathy target esterase of hen brain: Active site reactions with 2-[octyl3Hjoctyl-4H -I ,3,2-benzodioxaphosphorin 2-oxide and 2-octyl-4H -1,3,2[aryl-3Hjbenzodioxaphosphorin 2-oxide. 1. Neurochem. 64, 1680--1687.
CHAPTER
48 Cholinesterases Barry W. Wilson University of California, Davis
48.1 INTRODUCTION Cholinesterases (ChEs) are specialized carboxylic ester hydrolases that break down esters of choline. Two of special concern to the toxicology of pesticides are acetylcholinesterase (AChE, acety lcholine hydrolase, EC 3.1.1.7) and butyry lcholinesterase (BuChE, acylcholine acylhydrolase, EC 3.1.1.8). BuChE is also known as nonspecific cholinesterase or pseudocholinesterase. The preferred substrate for AChE enzymes is acetylcholine (ACh); nonspecific cholinesterase BuChE enzymes prefer butyrylcho1ine and/or propionylcholine, depending on the species (Silver, 1974). This chapter discusses these enzymes, their importance in understanding the toxicity of organophosphate esters (OPs) and carbamate pesticides (CBs), and their application to risk assessment of anticholinesterase pesticides and other agents (Taylor, 1999). ChEs are classed among the B-esterases, enzymes inhibited by OPs, and possessing a serine catalytic site (Aldridge and Reiner, 1972; Ballantyne and Marrs, 1992; Chambers and Levi, 1992; Ecobichon, 1996; Gallo and Lawryk, 1991). Other B-esterases include the broad class of carboxylesterases (CarbE, EC 3.1.1.1.), one of which is neuropathy target esterase (NTE), the enzyme associated with organophosphate-induced delayed neuropathy (OPIDN) discussed in other chapters. A different group of enzymes known as A-esterases (e.g., arylesterases, paraoxonases, and DFPases) actively hydrolyze OPs. They represent an important means of detoxification (Furlong et aI., 2000; La Du et aI., 1999; Haley et aI., 1999). There has been an immense amount of research on ChEs since 1914 when Sir Henry Dale (Dale, 1914) suggested there was an esterase capable of hydrolyzing acetylcholine in blood, and Abderhalden and Paffrath (1925) and Loewi and Navratil (1926) demonstrated tissue extracts that broke down the chemical. [There are over 14,000 research reports on ChEs listed in an ACS on-line database (Sci Finder Scholar, 1999) from 1992 through late 2000.] In the past decade, the tertiary structure and amino acid and DNA sequences of several ChEs have been elucidated. (For reviews, see Doctor et aI., 1998; Reiner et aI., 1999; Taylor, 1994, 1996.) Today, techniques such as site-directed mutagenesis and knock-out mutants enable investigators to dissect the form and function of these proteins litHandbook of Pesticide Toxicology Volume 2. Agents
erally one amino acid at a time (e.g., Faerman et aI., 1996; Gnatt et aI., 1994). Tomorrow, this knowledge will help design chemicals specifically targeted for the tertiary structure of these proteins. Specific reviews (e.g., Massoulie et aI., 1999; Taylor, 1996), conferences ( e.g., Doctor et aI., 1998; Reiner et aI., 1999), and even full-color molecular structures displayed on the Internet, help bring the reader the latest information in this rapidly moving research area. [For brevity, only selected references to a topic may be cited here. The reader is referred to these references and to the article of Gallo and Lawryk (1991), in a previous edition of this book, for citations to earlier work.] OPs with high toxicity were synthesized as chemical warfare agents in the late 1930s and early 1940s (Ecobichon, 1996; Holmstedt, 1963; Koelle, 1963). Their offspring have been adapted to agricultural use as pesticides. Synthetic CBs modeled on the natural carbamate physostigmine and specifically designed to inhibit ChEs have been in commercial use as pesticides since the 1950s (Ecobichon, 1996). Because of their potential as weapons, much research has focused on antidotes (e.g., oximes) and prophylactics to OP chemical warfare agents (e.g., National Academy of Sciences Report, 1999).
48.2 DISTRIBUTION ChEs are widely distributed across animal species (Ecobichon, 1996). Their presence in insects and other invertebrate pests have made anti-ChE agents popular and effective pesticides. Molecular forms of ChE similar to those in vertebrates have been studied in animals as varied as nematodes (e.g., Caenorhabditis; Cu1etto et aI., 1999), squid (e.g., Talesa et aI., 1999), and Amphioxus, a protochordate (Pezzementi et aI., 1998). Vertebrate-like AChE forms have been reported from Paramecium, a ciliated protozoan (Delmonte Corrado et aI., 1999). AChEs in the nervous system regulate excitation by destroying the neurotransmitter ACh. They are found at synapses, neuromuscular and myotendinous junctions, cerebrospinal fluid, central nervous system (CNS) neuron cell bodies, and axons, skeletal and smooth muscles (Silver, 1974). AChEs also are present on the surface of erythrocytes (RBCs) of mammals,
967
Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
968
CHAPTER 48
Cholinesterases
megakaryocytes, lymphocytes, and platelets (Husain, 1994; Paulus et al., 1981; Zajicek, 1957). Some research has been done on ChE in saliva (Ryhanen et aI., 1983; Yamalik et aI., 1990). Blood ChE forms are often used as surrogates for CNS enzymes in studies of toxicants. The AChE activity of human blood is restricted to its formed elements; most of the activity is vested in the RBCs (Wills, 1972). Plasma ChEs of other vertebrates (e.g., birds) often hydrolyze ACh too (Augustinsson, 1948, 1959a, b), and specific AChE enzyme activity is present in the plasma of some mammals. For example, the plasma ChE activity of rodents such as the laboratory rat is high in both AChE and BuChE (Traina and Serpietri, 1984). Neglecting plasma AChE activity may lead to misinterpretations of the extent of ChE inhibition in animals used for pesticide research and in the setting of regulations for food safety and human exposure (Wilson et aI., 1996). AChE activity also has been found in the serum of embryo mammals and birds, decreasing to adult levels after birth. AChE activity in fetal calf serum is high enough to be used as a source for purifying the enzyme (De la Hoz et aI., 1986). In contrast, adult bovine blood has relatively high RBC AChE and very low plasma ChE levels (Zajicek, 1957). Other species have a mixture establishable by studies of substrate and inhibitor specificity and, recently, by DNA analyses (Bartels et aI., 2000). Table 48.1 compares adult levels of RBC AChE for several species. The human has the highest activity. The rat has one of the lowest RBC AChE levels even though it is often used in biomedical research on anti-ChE agents. BuChEs are also found at synapses, motor endplates, and muscle fibers together with AChE (Silver, 1974). BuChE activity in blood is restricted to serum. The physiological functions of RBC and serum ChEs are unclear. The primary structure of ChEs is homologous to proteases and lipases (Taylor, 1996). One possibility is that blood ChEs evolved to protect the body from natural anti-ChE agents. A number of plant toxins have anti-ChE activity; these include the solanaceous glycoalkaloids, naturally occurring steroids in potatoes and related plants (Krasowski et aI., 1997; McGehee et aI., 2000),
Table 48.2 Plasma Hydrolysis of Choline Esters of Selected Species of Mammals
Table 48.1 Relative RBC AChE Levels of Adults of Selected Species Species Human
Sex
MIF
and the fungal territrems (Chen and Ling, 1996). The Calabar bean, Physostigma venenosum, was once used by West Africans in a "trial by ordeal" (O'Brien, 1967). Study of the action of its active anticholinergic ingredient, the CB physostigmine (eserine), helped to establish the roles of ACh and AChE in the nervous system (Engelhart and Loewi, 1930). Other examples of naturally occurring anti-ChE agents are: fasciculin from Elapsid snake venom (Marchot et aI., 1998), chaconine and solancine from tubers and nightshades (Nigg et aI., 1996), and huperzine from moss (Patocka, 1998). The association of cholinergic transmission with Alzheimer's disease is serving as a stimulus to modem studies of natural anti-ChE agents (e.g., Francis et aI., 1999; Nordberg and Svensson, 1998) as part of the search for treatments of this common disorder. Tissue ChEs may have specific but still unknown roles in addition to their regulation of neural transmission. For example, evidence has been amassing (e.g., Anderson and Key, 1999; Chiappa and Brimijoin, 1998; Layer et aI., 1998; Robitzki et aI., 1997; Sharma and Bigbee, 1998) for a developmental function for ChEs based on studies of neurite outgrowth in retinal and dorsal root ganglion cultures and embryos using immunological, sense and anti-sense oligonucleotides and inhibitors (reviewed by Layer et aI., 1998). Nevertheless, the successful development and survival of a knock-out mouse mutant lacking AChE indicates that other enzymes such as BuChE can function in its stead (Li et aI., 2000; Xie et al., 2000). There have been persistent reports concerning pesticide induced ocular damage (recognized clinically as Saku disease in Japan). These studies, reviewed by Dementi (1994), have stimulated studies by the U.S. EPA (Atkinson et aI., 1994; Boyes et aI., 1994), although without striking results. However, there are reports of visual changes and damage during growth and development (Geller et aI., 1998; Wyttenbach and Thompson, 1985). Hamm et al. (1998) found that diazinon, a widely used pesticide, damaged the development of the neural retina in Medaka, a fish.
ACh
PrCh
BuCh
AChE levels (%)
Man
135
310
360
2
100 ± 8.7
Cow
3
4
2
0
0
50
130
170
5
20 40
Guinea pig
MeCh
BzCh
Species
60
Cow
F
87.6 ± 1.9
Guinea pig
MfF
32.7 ± 3.5
Horse
130
225
365
2
28.8 ± 9.0
16
16
10
4
2
Rabbit
MIF MIF
Rabbit
21.7 ± 5.3
Rat
20
30
15
15
10
Rat
M
12.6 ± 3.0
Dog
70
115
180
6
30
MfF
30a
Cat
50
75
150
5
12
Horse
Cat
Source: Adapted from Zajicek (1957). Notes: Manometric assay at N = 3. Mean human AChE was 2180 C02/30 mintmg nitrogen. ACh substrate. aN =2.
~l
Source: Augustinsson (1959a). . Notes: Manometric assay. ~IC02/0.1 ml plasma/30 mm. N 3 or more animals. ACh: acetylcholine; PrCh: propionylchoIine; BuCh: butyrylcholine; MeCh: methylcholine; BzCh: benzoylcholine.
48.4 Multiple Molecular Forms and Life History
969
Table 48.3 Plasma Hydrolysis of Choline Esters of Selected Species of Birds, Reptiles, Amphibians, and Fish Species
ACh
PrCh
BuCh
MeCh
BzCh
Chicken
37
71
36
19
2
Duck
43
74
67
7
8
Turtle
14
103
27
7
Rana
40
90
87
2
2
5
9
12
2
Xenopus Pike
Aa
~
r
Source: Augustinsson (1959b). Notes: Manometric assay. ~IC02/0.1 m1 plasma/30 min. N 3 or more animals. ACh: acetylcholine; PrCh: propionylcho1ine; BuCh: butyrylcho1ine; MeCh: methylcholine; BzCh: benzoylcholine.
48.3 SUBSTRATE PREFERENCES AND SELECTIVE INHIBITORS AChEs prefer ACh as a substrate. Substrate preferences and activities of BuChEs vary with the species. For example, rat plasma BuChE activity has been reported to favor propionyl rather than butyryl substrates, and cows have hardly any plasma ChE activity at all (Tables 48.2 and 48.3). An important distinction between the AChEs and BuChEs is their response to substrate concentration. AChEs are inhibited by substrate in excess of a few mM; BuChEs are less sensitive (Hoffmann et aI., 1989; Wilson, 1999). In general, mammalian and avian AChEs rapidly hydrolyze ACh and its thiocholine analog acetylthiocholine (ACTh). Mammalian, but not avian, AChEs preferentially hydrolyze acetyl-,B-methylcholine. AChEs are selectively inhibited by several agents. One is the CB BW284c51 (1,5bis(4-allyldimethylammoniumphenyl)pentan-3-one dibromide) (Austin and Berry, 1953; Holmstedt, 1957; Silver, 1974); another is ARA 1327 (Augustinsson et aI., 1978). BuChEs are preferentially inhibited by iso-OMPA (tetraisopropylpyrophosphoramide; Austin and Berry, 1953), ethopropazine (Mikalsen et aI., 1986), and quinidine (Wright and Sabine, 1948). Effective concentrations for these selective inhibitors may vary by species. Useful starting points for testing are 0.1 to 0.01 mM.
48.4 MULTIPLE MOLECULAR FORMS AND LIFE HISTORY ChEs are polymorphic proteins; they occur in multiple molecular forms. AChEs consist of asymmetric and globular forms (Figs. 48.1,48.2). The asymmetric forms tend to be localized at synapses and motor endplates. They have glycosylated heads joined by sulfhydryl groups and collagen tails. The heads contain the active sites; the collagen tails attach the enzymes to cell surfaces. The globular forms are made up of the catalytic subunits (Taylor, 1996). Although these forms have similar kinetic properties, they differ in their ionic and hydrophobic interactions.
0
~L
10
GI
Jl~ 00
55
G2
G4
A4 Figure 48.1 Organization of subunits in the molecular forms of ChEs. Each circle represents a catalytic subunit. Globular forms are represented by "G." Asymmetric forms are designated by "A." Linear elements represent collagenlike tails. Disulfide bridge locations from studies of eel electroplax AChE, assumed to reflect ChEs from many sources. (After Brimijoin, 1992.)
ChE forms (Massoulie et aI., 1993, 1999) are synthesized as catalytic globular monomers (G 1) that oligomerize via disulfide bonds into multiple G2 and G4 forms (Fig. 48.1). The Gl subunits are synthesized within cells (e.g., nerve, muscle, and liver), glycosylated, and then secreted. Collagen tails are
Nucleus
% ."Y~"'"-
~~. <E)
t
70-aO%
Synthesis, Sequestering Glycosylation, Assembly of Globular Forms, Degradation
Processing, Assembly of Asymmetric Forms Transport
Vesicles Plasma Me
Secretion, Membrane Incorporation and Turnover Binding In Basal Lamina
Figure 48.2 Life history of ChEs. Enzyme is synthesized as a monomeric globular form (G 1). Up to 80% of the enzyme is degraded by intracellular proteases after ribosomal translation and before transit of the Golgi. In the Golgi, secretory forms (black triangles) are sequestered from membrane bound forms (open triangles), collagen-like tails are added to asymmetric forms, the peptide backbone is glycosytated, and the ChE becomes enzymatically active. Globular secretory forms may escape the synaptic cleft to enter extracellular fluids, blood, or external secretions. Asymmetric forms are probably bound quantitatively by adsorption or entrapment in the polyanion-rich, fibrous matrix of the extracellular synaptic basal lamina. (After Brimijoin, 1992.)
970
CHAPTER 48
Cholinesterases
Figure 48.3 Three-dimensional view of the active center of AChE. Modeled from Torpedo AChE with the addition of amino acid side chains of the mammalian enzyme. Amide backbone shown by the ribbons. The catalytic triad of the enzyme is Glu334, His447, Ser203 with hydrogen bonds indicated by dotted lines. The acyl pocket is Phe295 and Phe297; the choline subsite is Trp86, GIU202, and Tyr337; the peripheral site is Trp286, TYf72, TYf[24, and ASP74. Tyrosines 341 and 449 may help to stabilize some ligands. The catalytic triad, choline subsite, and acyl pocket are at the base of the 18-20 A gorge; the peripheral site is at its lip. (One way to see the diagram in three dimensions is to use 3-D glasses; another is to move the page toward your eyes until the images superimpose, keeping your head level. If there appear to be three images, concentrate on the middle one.) (After Taylor, 1996.)
attached to one, two, or three catalytic tetramers to yield A4, A8, and A12 asymmetric forms. Collagen-tailed forms become attached to the cell surface at specific binding sites. Globular forms are released into body fluids or bind to cell surfaces through hydrophobic amino acid sequences or glycophospholipids (Taylor, 1996). Antibodies have been prepared to several purified AChEs and BuChEs; specific protein and nucleic acid sequences have been determined and altered by site-directed mutagenesis (Doctor et aI., 1998). AChE and BuChE forms are each coded by single genes. The gene coding sequence for the collagen-tailed forms contains a C-terminal extension of 40 amino acids, the T-peptide, that interacts by a short proline rich attachment domain (PRAD) with the collagen. A single collagen gene (ColQ) is associated with both the AChE and BuChE collagen-tailed forms. The globular forms are synthesized by an alternative splice variant, the H-exon, that is expressed instead of the T-exon [see Krejci (1998) for a model and Massoulie et al. (1999) for further information]. The three-dimensional structures (Fig. 48.3) of ChEs are subjects of intense investigation (Doctor et aI., 1998; Massoulie et aI., 1999). In the past, the active site of AChE was loosely described by models based on work of Nachmansohn and Wilson (1951), in which there was a negatively charged "anionic" site and an "esteratic" site of catalytic residues. The positively charged choline moiety of ACh was hypothesized to bind to the negatively charged anionic site. A nucleophilic group, assumed to be a serine residue at the esteratic (acylation) site, was proposed to catalyze the hydrolysis [see Silman and Sussman (1998) for a more detailed historical perspective]. Recently, the crystallization of ChE proteins such as the dimeric AChE form of Torpedo californica has led to a more
detailed understanding of the form/function relationships of the enzyme (Doctor et aI., 1998). An important feature is the embedding of the active site of AChE in a "gorge" lined with 14 aromatic residues, about 20 A from the surface of the protein (Fig. 48.3). The quaternary nitrogen of choline binds through interactions with n electrons of tryptophan residues (Sussman et aI., 1991) at the "peripheral site" at the mouth of the gorge, a region conceptually corresponding to the historical concept of an "anionic site." One current view of the molecular mechanism has the ester substrate led down the gorge by molecular interactions to become hydrolyzed at the bottom by a catalytic triad of glutamate 334, histidine 447, and serine 203 residues (Doctor et aI., 1998; Taylor, 1996). Matters of current research include whether the products are ejected via a "side" door to make room for the next substrate molecule or are rapidly moved to the entrance of the gorge. The amino acid sequences around the entrance to the gorge that comprise the "peripheral site" may be important in determining the differences in substrate specificity and inhibition by excess substrate between AChE and BuChE forms (Doctor et aI., 1998; Reiner et aI., 1999). Figure 48.3 provides a three-dimensional diagram of an AChE molecule. (Other depictions may be found on the Internet by searching for "acetylcholinesterase.")
48.5 MECHANISM OF HYDROLYSIS E + AX
~+~ L]
EAX
~
EA + X
~
E+A
E: Enzyme; AX is substrate (ACh, acetylcholine) or inhibitor; EAX is reversible enzyme complex; X is Ch (choline); ks are reaction rate constants (see Rosenberry et aI., 1998; Taylor, 1996). The kinetics of ACh hydrolysis is a complicated multistep process [Fig. 48.4 from Taylor (1996) is one depiction]. It is discussed here in abbreviated form. The first step is a nucleophilic attack of the carbonyl carbon resulting in the formation of a reversible enzyme-substrate complex (EAX), acylation of the catalytic site (EA), and liberation of choline. This is followed by a rapid hydrolysis of the acylated enzyme, producing acetic acid and regenerating the enzyme (E + A). A similar reaction scheme applies to BuChEs (Taylor, 1996). The realtime kinetics of the enzyme reactions has been described by one biochemist (Quinn, personal communication) as "approaching catalytic perfection." The rate of ACh turnover (described by kcat! Km) is extremely rapid, where k cat = k2k3/ k2 + k3 (the geometric mean of the two rate constants). AChE is capable of hydrolyzing 6 x 105 molecules of ACh per molecule of enzyme per minute, a turnover time of 150 microseconds (Taylor, 1996). The upshot of this rapid rate of hydrolysis is that kl becomes the rate-limiting step for hydrolysis of ACh and its analog acetylthiocholine (AcTC). Perhaps this represents the diffusion of substrate to the active center. The deacylation step
. ~F ~, Glu334
i0A
.
-
_
~~i~4 Ser 203 ~~~~-"
Se~ 203 ~~=t~
~447
~ -,p447i~~
'&~~
-
---~l!?- ~
. -..----/~ :r -.~
~=~#/
~"¥-57
f~'
!~;:.=~
+
A
Enzyme-substrate
~rp86 . .c.~
(!/~r~~'jff;~~
complex
-~;
Glu334 ~ ~.
B
~
Glu334
Ser 203~-~">-
~l-."'S447 y~:r-
~
}Jl~ ~~~
\0
-....l ......
:.~/
Edrophonium Glu334
~Trp86
~~~Tif.~~-.
v
Ace~ enzy:e ;~~~~ ~~
->< d _~_(HiS447~~~~:;~ ~ ~;J~
4,lfi
?--o +
~
D Hydrolysis of acyl enzyme Glu 334
Ser203
Ser 203 ~::~:~:---
~-~
~rp86 Hydrolysis of dimethyl carbamoyl enzyme
.. ~'-. "~ (. Z ~,i;'?@~~
rp86 '/;'~"~~:",,, t~
.'
~20~~~
.~;, r ~~
~f&5'rt ~
f-
r::]~i~~-C] ~_
E
~p~
~
~
~J:;
.--,-.:;,:~~iJj/ -==-<--:;//
~
p-'-';-
Tetrahedral intermediate
H
~.:-;:-.-
* "' "
IDimethYI carbamoyl+enzyme
=;,~~\
Ser2o~~~~"~_
~/ 1<0519' ~~~
~ __
,.
~~,~__ ...::ifff" ~ . ';:@/
Diisopropylfluorophosphate
~e
~
rp86
J
~~'$;"n '(1!~~~\
, .'
1.1""
_""_
Diisopropyl phosphoryl enzyme
~
rp86
-
,'~~ :(/I~::::.~"
,lid·
~"._::::'::~$:"
K
.
Aged . phosphoryl enzyme
~rp86 I/n!:~lo~'-'~ ~~
L Reactivation of phosphoryl enzyme
I MOLECULE: 0 carbon • oxygen 0 nitrogen 0 hydrogen • phosphorus ® fluorine Figure 48.4 Hydrolysis of ACh by AChE and inhibition and reactivation of the enzyme. (A) Binding of ACh. (B) Attack by the serine hydroxyl, formation of a transient tetrahedral intennediate. (C) Loss of choline and formation of the acylated enzyme. (D) Deacylation of the enzyme by H20 attack. (E) Binding of the reversible inhibitor edrophonium. (F) Binding of neostigmine. (G) Fonnation of the carbamoylated enzyme. (H) Hydrolysis of the carbamolyated enzyme. (I) Binding of diisopropyl fluorophosphate. (J) Formation of the phosphoryl enzyme. (K) Formation of the aged form of the phosphoryl enzyme. (L) Attack by pralidoxime (2-PAM) to regenerate active enzyme. (After Taylor, 1996.)
972
CHAPTER 48
Cholinesterases
1,000 500 200 100
[;~ ~pH7.2
0.1
1
10
IAcetylthlochollnel (mM)
Figure 48.5 Effect of substrate concentration on activity of human blood AChE at pH 8.0 and pH 7.2. Ellman assay, pooled blood. (After Wilson et aI., 1997.)
-- 50 (!) ~
...... 20 (!)
~
0
(k3) is considered rate limiting to carbamoylating and phosphorylating agents. A major distinction between AChEs and BuChEs is the inhibition of AChE activity with increasing substrate concentration [S]. A plot of activity versus [S] for ACh, AcTC, and, in mammals, acetyl-,B-methy1choline, yields a curve with a maximum at 1-3 mM, whereas BuChE activity increases with [S] to at least 10 mM. These effects are illustrated in Fig. 48.5 from data of Wilson et al. (1997) for the human. The late Dr. A. R. Main (Hoffmann et aI., 1989) pointed out: "Because of this inhibition, methods for determining AChE activities should employ substrate concentrations at or below [S]opt." Unfortunately, his advice has not always been followed. The phenomenon of inhibition with excess substrate may be due to interactions at the peripheral site. Rosenberry et al. (1998) present evidence that excess substrate inhibition and the action of peripheral site inhibitors such as propidium are brought about by the imposition of a steric blockade in the catalytic pathway. Figuratively, a chemical cork blocks the gorge.
48.6 TOXICITIES OF ANTICHOLINESTERASES The knowledge of the three-dimensional structure of the ChEs has led to a better understanding of the mechanisms of action of drugs and chemical agents that inhibit the hydrolysis of choline esters. In general, there are three major domains for inhibitors to bind. They are the acyl and choline pockets of the active center and the peripheral anionic site. Tay lor (1996) uses edrophonium and tacrine as examples of reversible inhibitors that bind to the choline sub site near tryptophan 86 and glutamate 202; other reversible inhibitors such as fasciculin and propidium bind to the peripheral anionic site on AChE at the lip of the gorge encompassed by tryptophan 286 and tyrosines 72 and 124. CBs and OP pesticides inhibit enzyme activity by acting as alternate substrates to ACh. Carbamates give rise to a carbamoylated enzyme that is more stable than the acylated enzyme, taking minutes instead of milliseconds to rehydrolyze. Organophosphate esters are true hemisubstrates; they cova-
10
10
5
0 ....J
2
0.5 0.2
0.1
4
6 8 PI 50
10
12
Figure 48.6 Toxicity in vivo of directly acting OPs versus their inhibition of AChE in vitro. (I) Dipterex. (2) 0,0-diethyl-4-chlorophenylphosphate. (3) O,O-diethyl-bis-dimethyl pyrophosphorodiamide (sym). (4) TIPP. (5) O,O-diethylphosphostigmine. (6) Isodemeton sulfoxide. (7) Isodemeton. (8) Isodemeton sulfone. (9) DFP. (10) Diethylamidoethoxy-phosphoryl cyanide. (I\) O,O-dimethyl-O,O-diisopropyl pyrophosphate (asy). (12) Diethyl; amido-methoxy-phosphoryl cyanide. (13) Tetramethyl pyrophospnate. (14) O,O-diethyl phosphorocyanidate. (15) 0,0dimethyl-O,O-diethyl pyrophosphate (asym). (16) Soman. (17) TEPP. (18) O-isopropyl-ethylphosphone-fluoridate. (19) Tabun. (20) Amiton. (21) Diethylamido-isopropoxy-phosphoryl cyanide. (22) O,O-diethylS-(2-diethylaminoethyl)phosphorothioate. (23) Sarin. (24) O,O-diethylS-(2-triethylammoniumethyl)thiophosphate iodide. (25) Echothiophate. (26) Methylfluorophosphorylcholine iodide. (27) Methylfluorophosphorylbeta-methylcholine iodide. (28) O-ethyl-methylphosphorylthiocholine iodide. (29) Methylfluorophosphoryl-homo-choline iodide. (27), (28), and (29) with LD50 values from 0.03-0.07 mg/kg are not shown. (3\) Schradan and (32) dimefox shown on the graph were not used to calculate the regression. (After Hayes as shown in Gallo and Lawryk, 1991.)
lently bind with the serine at the active center, forming a tetrahedral configuration that resembles the transition state formed during hydrolysis of ACh. If the alkyl groups on the OP are methyl or ethyl, spontaneous regeneration may require hours, and may be even longer if tertiary alkyl groups are involved. Loss of one of the alkyl groups, a phenomenon known as aging, further stabilizes the phosphorylated enzyme, to all intents and purposes permanently inhibiting its catalytic ability. It is not appropriate to use the terms "reversible" and "irreversible"
973
48.6 Toxicities of Anticholinesterases Table 48.4
Table 48.5
Tissue IC50 Values for 4-Day and Adult Mice after Chlorpyrifos-Oxon
Necrosis of Rat Muscle after DFP and Botulinum Toxin (Btx)
Tissue Brain
Adult
Neonate
1O±0.2
9.6 ± 0.1
Liver
96±3
Plasma
18 ± 1.5
Purified
530 ± 50 330 ±28 3 nM for all tissues
Source: Mortinsen et al. (1998). Note: Values are nM, means ± SE.
to refer to the inhibitions brought about by CBs and OPs, respectively. Both classes of chemicals react covalently with the active center of the enzyme, and at some stage of the sequence of reactions, both enzyme-inhibitor complexes are rehydrolyzable. The toxicities of OPs and CBs often are correlated with the extent of their inhibitions of brain AChE. Figure 48.6 depicts the relationship between the toxicity in vivo of 30 directly acting OPs and their inhibition of AChE in vitro. However, such relationships do not necessarily signify that there are simple relationships between inhibition of AChE activity in an organ or tissue and in the test tube. For example, Mortinsen et al. (1998) measured IC50 values for chlorpyrifos-oxon with tissues from 4-day-old and adult rats (Table 48.4). The IC50s from young and adult brains were similar; the IC50s from the other tissues were not, differing by 5.5 (liver) and 20 (plasma) fold even though the IC50s of immunoprecipitated purified AChEs were the same, regardless of tissue or age. Possible interference of BuChE activities were excluded by using the specific BuChE inhibitor iso-OMPA. Factors such as A-esterase destruction, carboxyesterase binding, and sequestration of the lipophilic OP were considered possible factors to account for the differences between the in vitro and in vivo findings. One example of a carboxyesterase is serum paraoxonase (PON1), an A-esterase associated with high-density plasma lipoproteins; PONl destroys OPs such as the oxon analogs of parathion and chlorpyrifos. Direct evidence for its role in detoxifying OPs was provided by showing that mice exposed to chlorpyrifos were protected against cholinesterase inhibition and toxicity by administration of purified PONI (Li et al., 1995). Shih et al. (1998) demonstrated that knock-out PON I-deficient mice were more sensitive to chlorpyrifos and chlorpyrifos-oxon than genetically unaltered mice. Blood ChEs also have been shown to protect animals from OP toxicity. Studies on chemical warfare agents, led by the initial report of Wolfe et al. (1987), have shown that injection with purified AChE can protect mice and other animals from exposure to OPs (see Doctoretal.,1991). Many physiological actions of anti-ChEs are those expected from an excess of ACh caused by the inhibition of its catalysis. Specific symptoms depend on the chemicals and the receptors concerned (discussed elsewhere in this volume). Early signs of cholinergic poisoning likely involve stimulation of muscarinic neuroeffectors of the parasympathetic system. Symptoms in-
Treatment
EDL
SOL
0
0
85.5
28.0
Btx
0
0
DFP+Btx
0
1.59
Saline DFP
Source: Sket et al. (1991). Notes: Necroses/IOOO fibers; DFP injected 48 hours after Btx, sampled 24 hours later 1.5 mg/kg sc. EDL: extensor digitorum longus; SOL: soleus musce.
clude slowing of the heart (bradycardia), constriction of the pupil of the eye (miosis), diarrhea, urination, lacrimation, and salivation (Spencer et al., 2000; Taylor, 1996). Overstimulation at skeletal nicotinic neuromuscular junctions (motor endplates) causes muscle fasciculation (disorganized twitching) and, at higher doses, muscle paralysis. Increased ACh at cholinergic junctions of the sympathetic and parasympathetic autonomic ganglia affect the eye, bladder, heart, and salivary glands. Finally, anti-ChEs affect junctions of the central nervous system (CNS), producing hypothermia, tremors, headache, anxiety, convulsions, coma, and death. Whether or not there are consistent behavioral effects at low dose levels of OPs and CBs, such as deficits in learning and memory, is a matter of current research. In addition to affecting the nervous system, the excess ACh brought about by anticholinergic agents can cause a transient myopathy (Dettbarn, 1984). In vivo studies of Meshul (1989) and in vitro studies of the late Miriam Salpeter and colleagues using ACh receptor antagonists (Leonard and Salpeter, 1979) show this is due to an influx of Ca+ 2 and other cations into the postsynaptic cell. For example, necrosis due to diisopropyl fluorophosphate (DFP) was prevented in vivo by treatment with alpha-bungarotoxin, a snake venom agent that binds irreversibly to the nicotonic ACh receptor (Kasprzak and Salpeter, 1985). Further evidence that the necrosis is ACh mediated was provided by Sket et al. (1991). They demonstrated that botulinum toxin (a presynaptic inhibitor of ACh release) prevented muscle necrosis induced by diisopropyl fluorophosphate (DFP) in the rat (Table 48.5). ACh-induced myopathy may cause necrosis in 10 to 30 percent of the muscle fibers around the motor endplates (Dettbarn, 1984). Prolonged muscle weakness and muscle damage lasting several weeks or longer may occur. A similar transient muscle damage in humans has been termed Intermediate Syndrome (Senanayake and Karalliedde, 1987). Although it is generally accepted that most of the effects of OPs and CBs are due to inhibition of AChE, there is evidence for other modes of action of these agents. Anti-ChE pesticides have been shown to directly affect pre- and postsynaptic events (Pope, 1999). Electrophysiological studies suggest that choline itself may act as a regulator of nicotinic receptors in the CNS (Albuquerque et aI., 1998; Alkondon et al., 1997). Malathion and other OPs have been shown to affect the immune
974
CHAPTER 48
Cholinesterases
responses of mammals and fish (Beaman et aI., 1999; Rodgers and Ellefson, 1990; Rodgers and Xiong, 1997). A few OPs, including some pesticides (e.g., isofenphos, chlorpyrifos) and at least one chemical warfare agent (sarin), have been shown to cause OPIDN (organophosphate induced delayed neuropathy), a distal axonopathy. (This topic is discussed elsewhere in this volume.) Inhibition of ChEs plays a role in drug interactions. For example, cocaine is both detoxified by and is itself a reversible inhibitor of BuChEs. Studies on experimental animals indicate that depressing ChEs with anti-ChE treatments intensifies the toxic effect of cocaine (Hoffman et aI., 1992). Genetic variation between individuals can play an important role in the toxicity of anti-ChEs. One example is humans with inherited low levels of plasma BuChEs. Although usually symptomless, patients with genetically low BuChE given succinylcholine (or a similar drug) during surgery to induce relaxation of muscles are unable to speedily destroy the drug, intensifying and prolonging its activity, sometimes with serious consequences. People with such a genetic makeup can be detected by assays using dibucaine and fluoride (Silk et aI., 1979). At least two genetic polymorphisms, that of low BuChE and PONl, are considered risk factors for OP and CB pesticide exposures (Shih et aI., 1998).
48.7 ASSAY TECHNIQUES The history of ChE assays has been reviewed previously (e.g., Hoffmann et aI., 1989; Silver, 1974; Wills, 1972; Wilson, 1999; Witter, 1963). Some are "end-point" assays; others record the time course of the hydrolyses. One early assay to determine the hydrolysis of acetylcholine used a Warburg manometer to measure the C02 released from a bicarbonate containing buffer (Ammon, 1933). Although accurate, it is little used today. The particulars of this method are outlined by Wills (1972). Following World War 11, there was a widespread development of OPs for pest control. A veritable OP race ensued. By 1950, scientists were seeking rapid, accurate, and convenient clinical assays for blood ChE levels. Metcalf (1951) expressed the rationale for monitoring: "Since the cholinesterases of human blood are very sensitive to the presence of cholinesterase inhibitors, it appears that periodic estimation of blood cholinesterase levels may provide an indication of dangerous levels of overexposure to these toxicants." Several of the methods developed are still in use today. Hestrin (1949) determined the ACh remaining after incubation by reacting it with hydroxylamine under alkaline conditions to form a reddish-purple complex read at 515 nm. Metcalf (1951) adapted the method for drops of blood obtained with a spring-loaded lancet. Okabe et al. (1977) oxidized the choline released during ACh hydrolysis; the hydrogen peroxide produced was determined with an indicator reaction at 500 nm (see Abemathy et aI., 1988). Three kinds of ChE assays are commonly used today; one utilizes electrometric (pH), another uses radiometric, and a third
Table 48.6 Common ChE Assays Test
Basis
Conditions
lohnson and Russell
Radiometric
End point
(1975)
3H-ACh
Analysis Micro, fast, costly disposal
Michel (1949)
pH
Rate, ACh
Simple, slow, cheap
Ellman et al. (1961)
Colorimetric
Rate, ATCh
Micro, rapid
Adapted from Wilson and Henderson (1992).
uses colorimetric methodologies. Specific examples are listed in Table 48.6.
48.7.1 RADIOMETRIC An example of a radiometric technique is that of Johnson and Russell (1975). It is based on the differential solubility of ACh and its hydrolysis products in organic and aqueous media. Sample and tritiated ACh are reacted together; an organic solvent is added when the assay is completed. The radioactive acetate remains in the aqueous phase, quenching its scintillation; the radioactive ACh that remains in the organic phase is counted. The values are compared to a totally hydrolyzed ACh sample, usually accomplished by incubation with an excess of eel AChE. This and similar assays have high sensitivity, problems with dilution of samples encountered with "reversible" carbamate chemicals are minimized, many tubes may be measured at once, the sample size is small, and readouts may be computerized. However, radioactive assays involve a high initial investment and costly disposal of radioactive waste. Being an end-point assay, many duplicate samples are needed to run a kinetic analysis. Potter et al. (1993) applied a radioactive assay similar to that of Thomsen et al. (1988, 1989) in a field study of applicators of OPs and fumigants. In this radiometric end-point method, 14C-labeled ACh hydrolysis was stopped by ethanoVglacial acetic acid, the labeled acetic acid was evaporated, and the unhydrolyzed ACh counted.
48.7.2 pH The modified Michel method (Michel, 1949) directly determines the change in pH due to ACh hydrolysis with a pH meter or by titrating the acetic acid produced with NaOH, while keeping the pH constant (Groff et aI., 1976; Nabb and Whitfield, 1967). Potentiometric methods are reliable, they use simple reagents, and they are relatively inexpensive. However, they are limited by their relative insensitivity; they often have larger sample requirements and lower outputs than radiometric methods. Several micro pH methods have been described (see Gage, 1967); one using 10 !J.I of capillary blood was described by Mosca et al. (1995). pH assays can have relatively low variability. An early pH assay study of Rider et al. (1957) of 12 males and 12 females
48.7 Assay Techniques Table 48.7 Plasma and RBC ACh Hydrolysis Levels for 40-Year-Old Blood Donors Women
Men RBCChE
0.766 ± 0.081
0.750 ± 0.082
PlasmaChE
0.953
± 0.157
0.817+0.187
Source: Rider et al. (1957). Notes: ~pHlhour/O.02 ml red cells or plasma, 25°C, Mean ± S.D.
aged 40 years, taken from a study of 800 donors at a San Francisco blood bank, is shown in Table 48.7. Whether or not there is a difference between male and female subjects (as suggested in Table 48.7) is not clear. Many studies have noted a higher variability of plasma than RBC ChE activities. Gage (1967) points out that " ... population averages obtained by different investigators ... depend to some extent upon the method of assay used." Even so, Gage concluded on the basis of the studies available to him, that "an individual in good health with a plasma ChE 33 percent below the population average, or a red cell cholinesterase more than 20 percent below, has an abnormally low value and has probably been exposed to aChE inhibitor." Such "red-alert level" estimates have changed little in more than 30 years. 48.7.3 THIOL SUBSTRATES AND THE ELLMAN ASSAY Thiocholine substrate assays based on the work of Ellman et al. (1961) may be the most popular of ChE assays, replacing the pH methods. Many variations of the original Ellman assay have been published. The conditions of three popular commercial assays are shown in Table 48.8. Several automated versions (e.g., those of Technicon and COBIAS) are no longer marketed. The basis for the thiocholine assays is the hydrolysis of acetylthiocholine (ACTh) or related substrates by the enzyme, producing a thiol group that reacts with a sulfhydrylsensitive chromogen such as dithiobisnitrobenzoate (DTNB). In this case, peak absorption of the thionitrobenzoate produced is at 410-412 nm. The original Ellman assay used cuvettes; some
Table 48.8 Common Thiocholine-Based Assays
BIM Parameter
Ellman
Manual
Auto
Sigma
Wavelength (nm)
412
405
480
405
Substrate (mM)
ATCh
ATCh
ATCh
PrTh
0.5-1.0
5.4
5.4
5.0
pH
8.0
7.2
7.2
7.2
DTNB(mM)
0.32
0.24
0.24
0.25
Source: Wilson (1999). Notes: ATCh: acetylthiocholine; PrTh: propionylthiocholine; Boehringer-Mannheim (Roche); DTNB: dithionitrobenzoate.
BIM:
975
laboratories have modified the assay for a 96-well microplate reader in a manner similar to that described by Doctor et al. (1987). The Boehringer-Mannheim (Roche Diagnostics) kits (1981), Nos. 124117 and 450035, use ACTh as a substrate; the Sigma Diagnostic Cholinesterase (PTC) kit Procedure No. 422 uses propionylthiocholine (Sigma Diagnostics, 1989) for both AChE and BuChE. The Boehringer-Mannheim version used with a Hitachi automated spectrophotometer reads the reaction at 480 nm. This avoids possible interference of the Soret band of hemoglobin (Hb), but at the expense of reducing the sensitivity of the assay (Wilson and Henderson, 1992). The manual instrument kits recommend 405 nm. Instructions for both Boerhinger-Mannheim (Roche Diagnostics) and Sigma kits are for human blood. The reader is cautioned that the kits may not be suited to the needs of clinical veterinary laboratories or researchers without modification. In the case of the human, Wilson et al. (1997) showed that the high substrate concentration and low pH of the Boerhinger-Mannheim (Roche Diagnostics) kit introduce a difference of 40 percent in the manual assay compared to the Ellman assay run under optimum conditions (Fig. 48.5) . This is because the optimum concentration of substrate for human AChE is 1-2 mM and the optimum pH for the assay is pH 8.0, whereas the kit uses a substrate concentration of 5.4 mM and a pH of 7.2. Absorbances would have been even more reduced if the measurements for the BoerhingerMannheim (Roche Diagnostics) substrate and pH conditions were read at 480 nm (Wilson and Henderson, 1992). There have been recommendations that the Ellman assay be modified to use dithionitrobenzoic acid (DTNA), a chromogen that absorbs in the near-ultraviolet at 340 nm, to avoid the interference of Hb at 410 nm, thus divorcing the wavelength of the assay from the Soret band of Hb (Christenson et aI., 1994; Loof, 1992; Willig et aI., 1996). However, instruments reading in the near-ultraviolet are often more costly than those that register in the visual range. The stability of modem instruments should be sufficient to overcome the increased noise level of the assay at 410 nm, providing activity levels are sufficiently high. The substrate and pH of the Sigma Diagnostics Kit are optimal for neither RBC nor plasma BuChE. Augustinsson et al. (1978) proposed using propionylthiocholine, the Sigma kit substrate, as a compromise substrate for both enzymes. Under the conditions of their assay, it was no better a substrate for one than it was for the other. They recommended measuring the reaction in the ultraviolet, avoiding Hb interference that might well play a role, given the reduced sensitivity of the assay. The Sigma kit is no doubt excellent as a screen for patients with reduced BuChE activities before they undergo surgery and treatment with muscle relaxants like succinylcholine and mivacurium, but its use may be more difficult to justify, given the excellence of today's instrumentation, for the determination of RBC AChE activities to detect exposures to pesticides, since the conditions are not optimum for the enzyme. What to do? With blood enzyme assays, no one size seems to fit all. One approach is to focus on ACTh hydrolysis, and specific inhibitors such as iso-OMPA or quinidine to inhibit BuChE, establishing an estimate of BuChE
976
CHAPTER 48
Cholinesterases
Table 48.9 Relative Substrate Specificity of Plasma ChEs Species
Table 48.11 AChElBuChE Activity of Rat Plasma
Propionyl
Butyryl
Benzoyl
Dog
150
253
60
Horse
161
231
28
Cat
III
211
27
Man
155
192
36
Duck
139
153
25
Squirrel
122
144
14
Ferret
122
139
28
Hamster
153
128
24
Rat
211
119
17
Chicken
147
83
6
Mouse
139
75
11
Butyryl Favoring
Total ChE
BuChE
AChE
452 ± 17
175 ± 12
21O± 8
Source: Traina and Serpietri (1984). Notes: Mean±S.E. mU/ml; 18 rats, Ellman assay. Total with acetylthiocholine; BuChE with butyrylthiocholine; AChE with acetylthiocholine + 0.1 mM isoOMPA.
Propionyl Favoring
Source: Adapted from Hoffmann et al. (1989, Table 11.4) and Myers (1953). Note: Normalized to plasma ChE hydrolysis of ACh = 100%.
by difference, accepting, for convenience, the use of an inappropriate substrate and substrate concentration. The problems of using commercial kits with conditions designed for the human may be exacerbated when applied to other species. Many studies (some of which are summarized in Tables 48.9 and 48.10) indicate that, with other species, the relative activities of AChE and BuChEs, and even the properties of BuChEs, may differ from those of human enzymes. Lack of consideration for this may result in possible discrepancies and misinformation. For example, Harlin and Ross (1990) used adult bovine blood to establish conditions for the determination of cholinesterase activities in the only approved AOAC assay with ACTh as substrate and bovine blood. This excellent round-robin study did not discuss the report of Augustinsson (1959a, Table 51.2) that bovines had hardly any plasma cholinesterase activity. Further study is needed to establish whether the assay was determining only RBC AChE, and is useful for assaying it when serum cholinesterase is low or absent. Another problem arises when studies are undertaken with species that lack RBC AChE; the literature examined by the author suggests that only mammals have RBC AChE and that the properties of serum ChEs may vary with the species (Table 48.9). Table 48.10 Relative Acetyl Ester Hydrolysis Levels of Several Species
A further problem arises when, as is common in studies of laboratory animals that play a role in pesticide registration, the investigators assume AChE activity is restricted to the RBCs. Table 48.10 illustrates the wide differences between AChE activity in blood for humans and for two species often used in registration-geared research, the dog and the rat. Of the three, only the human has a relatively low ACTh hydrolysis rate in the plasma. Indeed, the AChE activity of rat plasma ( Table 48.11) may exceed the activity of plasma BuChE. Another source of plasma AChE is the platelets (Table 48.12). Although the AChE activity per platelet is high, its relative contribution to blood ChE is low, since the platelet content of blood is several orders of magnitude less than its content of RBCs. Studies of the particular AChE forms in the plasma of many species are lacking. An important and often unrecognized problem with thiocholine-based assays is the presence of a transient nonlinear "thiol oxidase" reaction with the color reagent DTNB in RBCs of some species. It is necessary that assays with species such as the rat that have such high "tissue blanks" be designed to either circumvent or correct for them since they can lead to indeterminate errors greater than 70 percent (Table 48.l3). One way is to include the appropriate blank in the assay. Another is to preincubate the samples for a few minutes before adding substrate until the nonlinear first 5-10 minutes ofthe assay are over (e.g., Chaney et aI., 2000). Lack of consideration of such problems in clinical assays of experimental animals submitted for regulatory purposes may have played a role in the difficulties encountered when the EPA tried to compare data from different laboratories on the rat (EPA, 1992). Errors may be compounded. For Table 48.12 Comparison of RBC and Platelet Activities of Several Species Species
Sex
Red blood cell
Platelet
Man
MIF
2180 ± 189
Cow
F
191O± 42
Guinea pig
MIF
713 ± 76
493 ± 119
Horse
MIF
627 ± 200
l500± 252
0 53 ±2l
Species
RBC
N
Plasma
N
Human
135 ± 29
60
37 ± 9.3
56
Rabbit
MIF
473 ± 116
3930 ± 208
M MlFa
274 ± 64
4240± 1070
Dog Male rat
17.9 ± 3.5
18
25.4 ± 5.5
18
Rat
9.0 ± 1.3
24
4.3 ± 1.0
45
Cat
Source: Adapted from Humiston and Wright (1967, Table 4). Notes: ACTh 0.8 mM (RBC); 7 mM (plasma) Autotechnicon analyzer, Mean± SD, N = Trials.
5450
30
Source: Adapted from Zajicek (1957). Notes: Manometric, J.4ICOz/30 minlmgN; N aN =2.
=
3.
977
48.7 Assay Techniques Table 48.14 Average Values of Cholinesterase Activity in Fasting Healthy Humans
Table 48.13 RBC DTNB Background Reaction in Various Species Species
Percent total activity
Subjects
Men
Women
Man
<10
N
40
38
Monkey
18-27
BuChE
8920 ± 2500
7490 ± 1950
Dog
30-60
AChE
121O± 200
1220 ± 123
Rabbit
50-70
AChE (Hb)
Rat
60-75
Mouse
<4
Brain (all species)
NIA
37.6 ±4.91
39.3 ± 4.49
Source: Sidell and Kamiskis (1975). Notes: BuChE in U/L, butyrylthiocholine, BoeringherlMannheim (Roche) Kit; AChE in nU/RBC and U/g Hb (Hb), acetylthiocholine according to Ellman. Mean ± standard deviation.
Source: Loof(1992). Notes: Transient activity, first 5-10 minutes, Ellman method.
example, a large and indetenninate error will occur when the Boehringer-Mannheim (Roche Diagnostics) kit assays are applied to rats without preincubating the samples. Such an error could be exacerbated by the low RBC activities and the misinterpreted plasma "BuChE" activities that are actually due to a mixture of the AChE and BuChE activities. The lesson is that, whatever the assay, it is critical that its conditions be validated for each species, tissue, and chemical under study. CBs represent a special problem because of the ease of rehydrolysis of carbamate-ChE complexes. The sensitivity of the assay requires dilution of the enzyme samples, but the act of dilution itself promotes rehydrolysis of the enzyme. Several investigators have suggested ways to minimize the problem, including Thomsen et al. (1988) for the radiometric assays and Nostrandt et al. (1993) for Ellman-type thiocholine-based assays. Regardless of the assay used, samples should be kept iced from the time of their collection to their assay. It is unfortunate that some are under the impression that OP exposures lead to "irreversible" inhibitions, and that icing a blood sample is unnecessary so long as an anticlotting agent such as EDTA or heparin has been used. To the contrary, as discussed elsewhere in this volume, ChE inhibitions by methyl-OPs such as azinphosmethyl (Guthion) have relatively rapid rates of spontaneous reactivation at room temperature; in this case, the half-life of recovery of activity for azinphos-methyl is approximately 2.5 hours (Wilson et aI., 1992b). The lack of a requirement to keep samples on ice as a part of assay protocols makes it difficult to interpret the results of otherwise excellently designed and executed studies such as that of Yeary et al. (1993). Although the instructions that accompany the commercial ChE monitoring kits discussed in this chapter recommend storing samples under refrigeration [4°C, Boerhinger-Mannheim (Roche Diagnostics), 2-6°C or -20°C (Sigma Diagnostics)], several clinical laboratories we contacted said they did not specify that samples be delivered to them on ice. 48.7.4 VARIABILITY
When it comes to studies using the Ellman assay, a paraphrase from George Orwell's classic "Animal Farm" (1945)
might be that "All Ellman thiocho1ine assays are created equal, but some are more equal than others." The variety of conditions used in field studies and laboratory experiments with thiocholine substrates, and the lack of an accepted standard assay and enzyme, make it difficult to compare the activities obtained from one experiment to another (Carakostas and Landis, 1991; Wilson et aI., 1992a). Perhaps this is what has led to the idea that thiocholine-based ChE studies are "too variable" to be relied upon for population exposure research and regulatory decisions, even though a number of carefully performed studies such as Sanz et al. (1991) indicate that ChE assays of populations can be performed with satisfactory results. For example, Sidell and Kamiskis (1975) measured RBC and plasma ChEs of a group of 22 subjects biweekly for a year, using the Technicon autoanalyzer. They found that RBC AChE levels varied less than hematocrit, Rb, or RBC counts. The annual average range of AChE values was 8% for men, 12% for women. The corresponding plasma ChE values were 25% for men, 24% for women. In general, as was true for the pR methods discussed earlier, plasma ChE values appear more variable than RBC AChE activities, whether the data are expressed on a per cell or a per Rb basis (Table 48.l4). Bellino et al. (1978) used fingersticks and saponin-hemolyzed human blood to carefully detennine the optimum conditions for the Ellman assay for human red blood cells and
Table 48.15 ChE Activities of Healthy Human Subjects Enzyme
Mean±SE
n
AChE Male
4.99+0.14
72
Female
5.18 + 0.18
71
Pooled
5.08+0.11
143
Male
2.26 +0.04
101
Female
2.46+ 0.09
71
Pooled
2.34+ 0.05
172
BuChE
Source: Modified from Bellino et al. (1978). Note: ID/ml whole blood; Ellman assay.
978
CHAPTER 48
Cholinesterases 200 ~-
6
180 160
~
'"
.. 140
4
~
:;:;
'E 120 ::> en
::I
'0
'0 100
3
'"
,g
§
z
2
.8E
80
Z
60
::>
40 20 9
11
13
15
17
19
21
23
25
27
29
31
AChE Activ itv (nmol/min/mA Hbl
Figure 48.7 Fingerstick AChE activities of VC Davis volunteers. Whole blood hemolyzed with Triton X 100 and assayed according to Ellman modified for a multiple plate reader. Activity is 14.6 ± 1.2 nmoles/minlmg hemoglobin. N = 13. (After Wilson et aI., © 1998, Plenum.)
plasma, demonstrating inhibition of AcTh hydrolysis with excess substrate, and an S-shaped substrate-concentration curve with butyrylthiocholine. They established that 0.01 mM eserine inhibited AChE and 0.3 mM eserine inhibited BuChE. Similarly, they found that 7.0 mM totally inhibited both AChE and BuChE activity, and 2.8 mM of sodium dodecylsulphate inhibited AChE but not BuChE. Their study of healthy men and women showed relatively low variability. See Table 48.15. Large sample numbers were obtained in a study of farm worker families from migrant housing centers in California by Wilson et al. (1998), in which almost 900 volunteers contributed fingersticks of blood. Ten microliters of blood were hemo1yzed at the site, transported on ice to the laboratory, stored at -70°C, analyzed under optimum assay conditions for RBC AChE using quinidine to inhibit plasma BuCh, and expressed on a Hb basis. Mean activity of the migrant housing center families (n = 894) was 14.6 ± 2.6 nmolfminlmg Hb, virtually the same as those from fingersticks and venous blood draws of approximately a dozen UC Davis volunteers. See Figs. 48.7 and 48.8. One problem to circumvent is the possible contamination of the sample with pesticide on the skin (Yuknavage et aI., 1997).
48.8 STANDARDS Several companies provide AChE standards for laboratory use using human and other species ChE preparations. Wilson et al. (2000) have been testing an AChE standard prepared by hemolyzing washed bovine RBCs. These RBC ghosts showed low variability when stored either at refrigeration (+4°C) or low-temp (-75°C) freezer temperatures for more than 250 days, suggesting that such preparations are suitable for laboratory standards. Similar results have been obtained in shorterterm studies for a lyophilized AChE from Sigma Inc.
9
11 13 15 17 19 21 23 AChE Activity (nmol /m in/mg Hb)
25
27
29
31
Figure 48.8 Fingerstick AChE activities of migrant family center residents. Blood filled capillaries kept on ice before returning to the laboratory. Whole blood assayed as in Fig. 48.7. Activity is 14.6 ± 2.6 nmoles/minlmg hemoglobin. N = 894. (After Wilson et aI., © 1998, Plenum.)
48.9 FIELD KITS Well-designed, reliable field kits for cholinesterase determinations would be valuable for monitoring the health of those who apply pesticides and those who work in agricultural workplaces subject to pesticide spraying. Enzyme-impregnated filter paper, potentiometric sensors, and colorimetric comparisons have been used in the past (e.g., Collombel and Perrot, 1970; Dahlgren, 1983; Gamson et aI., 1973; Rogers et aI., 1991). A farely new device is the EQM Test-Mate kit (Magnotti et aI., 1988; Magnotti and Eberly, 1996). It uses a solid-state device and the Ellman method to measure ChE activity in a drop of whole blood obtained by a fingerstick. Several models have been marketed. The Test-Mate has the advantages of portability, relative low cost, conveniently prepared reagents, and small sample volumes. Field and laboratory studies have been conducted with it (Keifer et aI., 1996; Prall et aI., 1998). Nevertheless, the current model (Model ChE) is not recommended by the manufacturer for field use; the instructions advise it be operated in the laboratory by a trained technician. One difficulty with using whole blood is sensitivity; blood samples are not diluted as much as they would be using a larger, more sensitive device. Under these conditions, readings may be affected by the relatively high absorption of the Soret band of Hb at 412 nm, the optimum absorbance of DTNB, the Ellman colorimetric reagent. The Test-Mate models have attempted to circumvent this by using higher wavelengths, sacrificing sensitivity of the chromogen for a lower noise level (Wilson and Henderson, 1992). A second problem with the device is that the Test-Mate does not display the raw absorbance values of the reaction. Instead, it displays values normalized to those expected at 25°C, employing a temperature sensor and a built-in algorithm that has been criticized for accuracy temperature (Amaya et aI., 1996; London et al., 1995; Wilson et aI., 1998).
48.13 Significance of Blood ehEs
48.10 REGULATORY MATTERS: ARE ChE INHIBITIONS ADVERSE EFFECTS? Shortly after the introduction of OP pesticides, the experiences of University of California scientists led to a recommendation that " ... individuals showing a 20 percent or more depletion from normal pre-exposure plasma ChE levels should discontinue participation in the work ... until cholinesterase levels have returned to normal" (Metcalf, 1951). Almost 50 years and many studies and task forces later, similar guidelines are still used. The rationale for choosing one specific decrease in ChE level over another as constituting a health hazard is not clear. One theory is that since most clinical laboratories should be able to detect a 20 percent decrease in ChE activity, and since dose/response curves for many anti-ChEs tend to be steep, this level provides a realistic statistically significant difference between test groups suitable for regulatory purposes.
48.11 BLOOD ChEs AND DETECTION OF EXPOSURE Whether or not a decrease in ChE activity in the blood constitutes an "adverse effect" raises questions concerning the short- and long-term health of individuals, populations, and their progeny that have yet to be answered to the satisfaction of most investigators. Detecting a statistically significant decrease in blood ChE levels between a putative exposed group and an unexposed group, or compared to an accepted "normal" range, is usually accepted as indicating that a potentially hazardous exposure to an anti-ChE chemical has occurred. AntiChE chemicals are not usually long-lasting within the body, and blood ChE activity can be expected to recover relatively rapidly from inhibition by an OP or CB. In the human, RBCs (and their AChE activities) are replaced at approximately 0.9 percent per day (a 120-day life span). Plasma BuChE is replaced even more rapidly (Boyer et al., 1977).
48.12 REACTIVATION OF INHIBITED AChE The discovery of Wilson and Ginsburg (1955) that oximes could displace OPs from the active site of ChEs, restoring enzyme activity, provided an important treatment for OP poisonings. It also opened the door to using reactivation of inhibited enzymes to establish that exposure had occurred when reliable unexposed control data or normal ranges of activity were lacking (Hansen and Wilson, 1999) reviewed by Wilson et al., I 992b. Useful as it may be, the application of reactivation techniques is not currently performed on a routine basis by clinical laboratories known to the author. Benschop and colleagues treated OP-inhibited serum ChE with potassium fluoride at pH 4 to chemically detect the presence of the inhibited OP-ChE. The fluoride ions reactivated the inhibited enzyme, converting the OP moiety into the corresponding phosphofluoridate and permitting its subsequent identification. Benschop's
979
group applied the technique to obtain direct evidence of exposure to sarin in several victims of the release of the nerve gas in a Tokyo subway in 1986, and from an earlier incident in Matsumoto Japan (Polhuijs et aI., 1997, 1999).
48.13 SIGNIFICANCE OF BLOOD ChEs Even if ChEs were not the targets of a multibillion-dollar, worldwide pesticide industry, they would still be of great interest to physiologists, cell biologists, and biochemists. ChEs have important roles in regulating neural transmission, and they are targets for pharmacological interventions in disorders such as glaucoma, myasthenia gravis, and Alzheimer's (Taylor, 1996). The widespread use of OP and CB pesticides and the dangers attendant upon their applications have resulted in ChE enzymes being used as biomarkers of both exposure and effect in assessing the risks to workers in agriculture and to consumers. ChEs enter into the picture in several ways. One is in the use of experimental animals, most often rats, but also dogs, rabbits, and other species, to determine no-effect levels that can be extrapolated to the human. Another is in the recommendation of safe residue levels in food. A third role is to help provide a safe agricultural workplace by monitoring those who could be exposed to dangerous levels of anti-ChE chemicals (i.e., mixer loaders, applicators). A fourth role is to decide whether poisoning episodes have involved ChE-inhibiting agents. One factor in favor of using ChEs for such clinical and regulatory ends is the ease with which they may be rapidly and inexpensively assayed compared to analyzing for the cholinergic chemicals themselves. Generations of toxicologists and public officials have worked to establish ChE assays as a simple way to help provide answers to complicated questions of health effects, exposures, and risk. Today, agencies require submission of blood and brain ChE levels from experimental animals after short-and long-term experiments as part of the registration process for pesticides. Most agree with the position that statistically significant decreases in brain ChE activities, when accompanied by knowledge of the doses involved, are useful for establishing quantitative toxicity indices. Even so, issues such as the significance of ChE levels in specific parts of the brain and the applicability of one animal model over another are unresolved. There is continuing discussion of the significance of monitoring blood ChEs of humans and other animals. One issue is the role of no observable adverse effect levels (NOAELs, the highest dose levels at which no important effect of a drug is observed) in assigning safe levels for toxic chemicals. One way to establish NOAELs is by batteries of behavioral tests under controlled laboratory conditions. Another is to measure residues on skin and clothing, urinary metabolites of agricultural workers, and fecal metabolites of laboratory and wild animals. Blood ChE levels represent standardized, relatively inexpensive measurements of a biochemical effect due to an exposure to a toxic chemical, in addition to providing evidence of the exposure itself (Nigg and Knaak, 2000). But some do not agree. For example, an industry "Acute Cholinesterase Risk Assessment
980
CHAPTER 48
Cholinesterases
Work Group" has published a review of RBC AChE, plasma ChE, and brain AChE activities focusing on their use in risk assessment (Carlock et aI., 1999). The review drew upon the literature and a compilation of toxicity data from previously unpublished industry sources in which the chemicals were listed by code and category. The work group concluded in part that: (a) plasma ChE should not be used in risk assessments since it is not an adverse effect; (b) RBC inhibition is not per se an adverse effect; (c) when available, cholinergic effects or brain AChE levels should be used ahead of RBC AChE values in setting NOELs, and that human data should take precedence over animal-derived results. In their discussion of NOELs, the group did not take into account that much of the data were derived from studies of rats and dogs and were not corrected for blood AChE levels or thiol oxidase activities of the RBCs, as discussed earlier in this chapter. A moderate position might be to use blood ChE values as biomarkers of exposure to anti-ChE inhibitors, rather than to insist they be considered quantitative indicators of physiological effects, thus supporting their use as early warning signs and as important weight-of-evidence factors.
48.14 DIRECT EFFECTS Although much of the attention of neurotoxicologists has been directed toward understanding the nature of the inhibition of ChEs and their toxicological consequences, a number of neuroscientists have been studying direct effects of anti-ChE agents on the presynaptic release of ACh (e.g., Rocha et al., 1996) and on the postsynaptic target receptors of ACh. For example, van den Beukel et al. (1998) found that micromolar levels of physostigmine, parathion, paraoxon, and phenyl saligenin cyclic phosphate blocked ACh-induced transient nicotinic inward currents in mouse NIE-115 and human neuroblastoma and locust thoracic ganglion cells. The Eldefrawis' and colleagues (Katz et aI., 1997) demonstrated that chlorpyrifos, parathion, and their oxons bind to and desensitize the nicotinic receptor (nAChR) of Torpedo, the electric ray. Narahashi's group (Nagata et aI., 1997) used rat clonal phaeochromocytoma (PCI2) cells to demonstrate that neostigmine and carbaryl blocked nAChR channels. In contrast, Albuquerque's group (Camara et aI., 1997) found that methamidophos did not affect neurotransmitter release or act directly on rat nAChR but that choline itself affected the response of the receptors (AIbuquerque et aI., 1998). The low levels at which these electrophysiological effects occur strongly suggest they should be taken into consideration when considering the effects of ChEinhibiting OPs and carbamates, such as the report of Burruel et al. (2000) of effects of methamidophos on sperm in mice.
48.15 ANTIDOTES ChE inhibitions by OPs and CBs are one of the few examples of enzyme inhibitions for which there are specific antidotes. Two
drugs in use are atropine and pralidoxime (2-PAM). Doses depend on the extent of exposure and species. Atropine binds to the muscarinic ACh receptor (mAChR), reducing the effectiveness of the excess ACh generated by the inhibition of AChE. It is given intravenously as required to relieve the symptoms of excess ACh in cases of pesticide poisoning. Oximes directly reactivate OP-inhibited AChEs (Wilson and Ginsburg, 1955; reviewed in Wilson et aI., 1992b). 2-PAM Cl (Protopam) is the oxime registered for use in the United States; its methanesulfonate salt (P2S) is used in Europe. Reactivation of the enzyme involves transfer of the substituted phosphate or phosphonate residue from the catalytic site of the enzyme to the oxime. There has been much research on treatment of exposure to ChE inhibitors focusing on chemical warfare agents (e.g., Hille et aI., 1995; Raveh et aI., 1996). One of the U.S. Department of Defense kits is "Convulsion Antidote for Nerve Agents (CANA)." It has 2 ml of the anticonvulsant diazepam, and "Nerve Agent Antidote Kit (NAAK)" containing autoinjectors with 2 mg of atropine and 600 mg of pralidoxime. Another kit, "Nerve Agent Pre-treatment Tablets (NAPS)" contains 30mg tablets of pyridostigmine bromide to be taken orally three times a day. A Swedish auto-injector contains HI-6 (500 mg), an oxime not available in the United States, and atropine (2 mg). The logic behind the use of pyridostigmine bromide is that this CB AChE inhibitor will reduce the extent to which AChE becomes irreversibly inhibited by rapidly aging chemical warfare agents such as soman, by virtue of its ability to temporarily occupy the catalytic site of AChE, interfering with its more permanent phosphorylation by soman and other OPs (Tuovinen et aI., 1999). Three 30-mg doses of pyridostigmine bromide per day were given to the Allied Forces during the Gulf War under an experimental-drug FDA license. This led to speculations that the interaction of this CB with other neuroactive chemicals present in the sector was a factor in the cluster of symptoms termed the Gulf War Syndrome (Abou-Donia et al., 1996a, b; Kurt, 1998). An alternative approach to OP poisoning is to destroy the agent within the body. Two of the most successful treatments are to inject purified ChEs to bind the agents and phosphotriesterases to destroy them (e.g., Tuovinen et aI., 1999; Wolfe et aI., 1987). Experimental evidence of nerve damage (in this case to chickens) was reported by Abou-Donia and co-workers after combined treatments of pyridostigmine bromide, DEET (an insect repellent), chlorpyrifos (Abou-Donia et aI., 1996a), or permethrin (Abou-Donia et aI., 1996b).
48.16 RISK ASSESSMENT AND ChEs ChE activity has been used for many years as a biomarker of exposure and effect for setting regulations for anti-ChE pesticide use and for safe levels of such pesticides in foods and in the environment. Almost three-quarters of a century of research has provided tools to qualitatively establish whether exposure has occurred to man and other animals in the laboratory, the clinic, and the environment. Nevertheless, problems arise when
48.17 Cholinesterases Tomorrow
data are used quantitatively, such as in setting NOAELs of a pesticide in food, or in estimating lifetime exposure levels of a chemical warfare agent. The lack of universally acceptable standards for ChE assays and the difficulty in deciding whether ChE levels are, in and of themselves, an adverse effect are two of the difficulties encountered when using ChE activities for regulatory purposes. Even so, cholinergic mechanisms became a criterion for creating a category of aggregate pesticide use in the new Food Quality and Protection Act (Mileson et aI., 1998). The U.S. Food Quality and Protection Act was passed unanimously by Congress on August 3, 1996. It created a single health-based safety standard for pesticide residues in food and removed the regulation that specified zero tolerance from pesticides that may be concentrated in processed commodities. It required a "reasonable certainty" that no harm will result from aggregate exposures, that is, exposures from chemicals with a common mode of action. The Act was health based; that is, exposures from diet, including drinking water, and nonfood exposures (e.g., residential, lawn, garden, indoor, institutional, industrial uses) were all to be considered. Children received special treatment; up to 10 additional uncertainty factors could be added for them. The Act specifies the use of sound science in making the determinations and requires a focus on health-based approaches to food safety and the promotion of safer, effective pest control methods. Applying the policies delineated by the FQPA may occupy the attention of the agricultural community, government regulators, and toxicologists specializing in agricultural chemicals for the next decade. It is safe to predict that there will continue to be sticky issues at each step of the risk assessment process, in the assessments of hazard, dose-response and exposure, and the characterization of risk. A special problem will involve determining aggregate exposures, deciding which chemicals should be included as part of a common "Risk Cup." ChE-inhibiting chemicals played a leading role in this venture when a panel of distinguished toxicologists asserted that inhibition of AChE was a "common mechanism of toxicity" for OPs. They concluded "that OP pesticides act by a common mechanism of toxicity if they inhibit phosphorylation and elicit any spectrum of cholinergic effects" (Mileson et aI., 1998). It may be too soon to say whether such a pharmacologically simplistic, but regulatorily valuable, view will win the day. Pope (1999), one of the authors of the "common mechanism" review, concluded in a separate article that "additional macromolecular targets for some OP pesticides ... may alter the cascade of events following AChE phosphorylation and thereby modify that common mechanism .... " His review analyzed the comparative toxicity of 38 OP AChE inhibitors currently in use in pesticides. Demonstrated direct effects of anticholinesterases on receptors and transmitter release discussed earlier in this chapter, and the evidence of groups like Albuquerque's that choline itself may have an effect on ACh receptors (Albuquerque et aI., 1998; Alkondon et aI., 1997), suggest that there may be holes in the "Risk Cup." Whether the leaks are large enough to lead to a modification of the "common mechanism" policy is a matter for the future.
981
48.17 CHOLINESTERASES TOMORROW The specific course of research on ChEs parallels, indeed has been at the cutting edges of some of the advances of the 20th century in biochemistry, pharmacology, physiology, cell biology, and toxicology. Many of today's disciplines had not yet been christened when ChEs were first recognized as special proteins intimately associated with the regulation of the nervous system. Although it is likely that some anticholinergic agricultural chemicals will be replaced by newer agents with different modes of action, ChEs are not likely to languish. (Registrations of several OPs already have been cancelled or restricted in use in the last ten years in the United States.) Nevertheless, use of OPs and CBs elsewhere on the planet will certainly continue for the foreseeable future, if only because some countries may not be able to afford the new generations of chemicals. The rapid development of cDNA microarrays suggests that studies of the impact of anticholinergic chemicals will soon be routinely conducted on the level of responsive genes (Gupta et aI., 1999). The advent of probabilistic methods of assessing risk of pesticides to human health and wildlife will create opportunities to apply sophisticated methods of determining risk from anticholinergic agents based on the population distributions of exposures and effects (Boyce, 1998). Research into the molecular bases of pharmaceuticals, enzyme action, development of the nervous system, and many other basic features of living systems will benefit from studies of ChEs for many years to come. And, dismayingly, the simplicity of synthesizing and deploying anti-ChEs as weapons of war and terror is likely to be as tempting to governments and terrorists of the 21st century as they have been to some in the 20th. Today's scientists leave a legacy of knowledge for those who come after, but the gift of wisdom is not theirs to bestow, much as they might wish to do so.
REFERENCES Abderhalden, E., and Paffrath, H. (1925). Beitrag der Frage der Inkret(Honnon)-Wirkung aud die motorischen Funktionen des Verdauungskanales V. Uber die synthese von cholinestem aus Cholin und Fettsauren mittels Fennenten des Dunndann. Fermentforsch 8, 299. Abemathy, M. H., Fitzgerald, H. P., and Ahem, K. M. (1988). An enzymatic method for erythrocyte acetylcholinesterase. Clinical Chemistry 34, 10551057. Abou-Donia, M. B., Wilmarth, K. R, Abdel-Rahman, A. A., Jensen, K. E, Oehme, E W., and Kurt, T. L. (l996a). Increased neurotoxicity following concurrent exposure to pyridostigmine bromide, DEET, and chlorpyrifos. Fund. Appl. Toxicology, 34,201-222. Abou-Donia, M. B., Wilmarth, K. R., Jensen, K. E, Oehme, E W, and Kurt, T. L. (I 996b ). Neurotoxicity resulting from coexposure to pyridostigmine bromide, deet, and pennethrin, implications of Gulf War chemical exposures. J. Toxicology Environ. Health, 48,35-56. Albuquerque, E. X., Pereira, E. E, Braga, M. E, and Alkondon, M. (1998). Contribution of nicotinic receptors to the function of synapses in the central nervous system, the action of choline as a selective agonist of alpha 7 receptors. J. de Physiologie 92, 309-316. Aldridge, W N., and Reiner, E. (1972) "Enzyme Inhibitors as Substrates." North-Holland!American Elsevier, Amsterdam. Alkondon, M., Pereira, E. E, Cortes, W. S., Maelicke, A., and Albuquerque, E. X. (1997). Choline is a selective agonist of alpha7 nicotinic
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Silk, E., King, J., and Whittaker, M. (1979). Assay of cholinesterase in clinical chemistry. Ann. CUn. Biochem. 16,57-75. Silman, 1., and Sussman, J. L. (1998). Structural and functional studies on acety!cholinesterase: A perspective. In "Structure and Function of Cholinesterases and Related Proteins" (B. P. Doctor et aI., eds.), pp. 2533. Plenum Press, New York. Silver, A. (1974). "The Biology of Cholinesterases." Frontiers of Biology, Vo!. 36. North-Holland, Amsterdam. Sket, D., Dettbarn, W.-D., Clinton, M. E., Misulis, K. E., Sketelj, J., Cucke, D., and Brizin, M. (1991). Prevention of diisopropylphosphorofluoridateinduced myopthy by botulinum toxin type A blockage of quantal release of acety!choline. Acta Neuropathologia 82, 134-142. Spencer, P. S., Schaumburg, H. H., and Ludolph, A C. (2000). "Experimental and Clinical Neurotoxicology," 2nd ed. Oxford University Press, New York. Sussman, J. L., Harel, M., Frolow, E, Oefner, C., Goldman, A, Toker, L., and Silman, I. (1991). Atomic structure of acetylcholinesterase from Torpedo califomica: A prototypic acetylcholine-binding protein. Science 253, 872879. Talesa, V., Grauso, M., Arpagaus, M., Giovannini, E., Romani, R., and Rosi, G. (1999). Molecular cloning and expression of a full-length cDNA encoding acetylcholinesterase in optic lobes of the squid Loligo opalescens, a new member of the cholinesterase family resistant to diisopropyl fluorophosphate. 1. Neurochem. 72(3), 1250-1258. Taylor, P. (1999). Esterases reacting with organophosphorus compounds. Chemical-Biological Interactions 119·120, 1--620. Taylor, P. (1996). Cholinesterase agents. In "Goodman and Gilman's The Pharmacological Basis of Therapeutics," 9th ed. (1. G. Hardman et al., eds.), pp. 161-176. McGraw-Hill, New York. Taylor, P. (1994). The cholinesterases: From genes to proteins. Ann. Rev. Pharmacol. Toxieol. 34, 281-320. Thomsen, T., Kewitz, H., and Pleul, O. (1989). A suitable method to monitor inhibition of cholinesterase activities in tissues as induced by reversible enzyme inhibitors. Enzyme, 42, 219-224. Thomsen, T., Kewitz, H., and Pleul, O. (1988). Estimation of cholinesterase activity (EC 3.1.17; 3.1.1.8) in undiluted plasma and erythrocytes as a tool for measuring in vivo effects of reversible inhibitors. J. CUn. Chem. Clin. Biochem. 26,469-475. Traina, M. E., and Serpietri, L. A. (1984). Changes in the levels and forms of rat plasma cholinesterases during chronic diisopropylphosphofluoridate intoxication. Biochem. Pharmacol. 33, 645-653. Tuovinen, K., Kaliste-Korhonen, E., Raushel, E M., and Hannincn, O. (1999). Success of pyridostigmine, physostigmine, eptastigmine and phosphotriesterase treatments in acute sarin intoxication. Toxicology 134, 169-178. U.S. Environmental Protection Agency (EPA) (1992). "Workshop on Cholinesterase Methodologies." Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. van den Beukel, 1., van Klcef, R. G., and Oortgissen, M. (1998). Differential effects of physostigmine and organophosphates on nicotinic receptors in neuronal cells of different species. Neurotoxieology 19,777-787. Willig, S., Hunter, D. L., Dass, P. D., and Padilla, S. (1996). Validation of the use of 6,6' -dithiodinicotinic acid as a chromogen in the Ellman method for cholinesterase determinations. Veterinary Human Toxicol. 38, 249-253. Wills, J. H. (1972). The measurement and significance of changes in the cholinesterase activities of erythrocytes and plasma in man and animals. CRC Crit. Rev. Toxicol. 1, 153-199. Wilson, B. W. (1999). Cholinesterases. In "Clinical Chemistry of Laboratory Animals," 2nd ed. (E. Quimby and W. Loeb, eds.), pp. 430-440. Taylor and Francis Inc., Philadelphia.
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Wilson, B. W., and Henderson, J. D. (1992). Blood esterase determinations as markers of exposure. Rev. Environ. Contam. Toxicol. 128,55-69. Wilson, B. w., Henderson, J. D., Bosworth. D. H., and Oliveira, G. H. (2000). Standardization of cholinesterase measurements for monitoring human exposures. Book of Abstracts, 219th ACS National Meeting, San Francisco, CA, March 26-30. Wilson, B. w., Hooper, M. J., Hansen, M. E., and Nieberg, P. S. (l992b). Reactivation of organophosphate inhibited AChE with oximes. In "Organophosphate;, Chemistry, Fate and Effects" (J. E. Chambers and P. E. Levi, eds.), pp. 107-137. Academic Press, New York. Wilson, B. w., Jaeger, B., and Baetcke, K., Eds. (I 992a). "Proc. U.S. EPA Workshop on Cholinesterase Methodologies," Arlington, VA, Dec. 4-5, 1991. Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. Wilson, B. W., McCurdy, S. A., Henderson, J. D., McCarthy, S. A., and Billitti, 1. E. (1998). Cholinesterases and agriculture, humans, laboratory animals and wildlife. In "Structure and Function of Cholinesterases and Related Proteins" (B. P. Doctor et al., eds.), pp. 539-546. Plenum Press, New York. Wilson, B. W., Padilla, S., and Henderson, J. D. (1996). Factors in standardizing automated cholinesterase assays. J. Toxieol. Environ. Health 48, 187-195. Wilson, B. W., Sanborn, J. R., O'Malley, M. A, Henderson, J. D., and Billitti, J. R. (1997). Monitoring the pesticide-exposed worker. Occupat. Med. 12, 347-363. Wilson, J. B., and Ginsburg, S. (1955). A powerful reactivator of alkyl phosphate-inhibited acetylcholinesterase. Biochim. Biophys. Acta 18, 168170. Witter, R. E (1963). Measurement of blood cholinesterase. Arch. Environ. Health 6, 537-563. Wolfe, A. D., Rush, R. S., Doctor, B. P., Koplovitz, L, and Jones, D. (1987). Acetylcholinesterase prophylaxis against organophosphate toxicity. Fund. Appl. Toxieol. 9, 266-270. Wright, C. I., and Sabine, J. C. (1948). Cholinesterases of human erythrocyte and plasma and their inhibition by antimalarial drugs. J. Pharmacol. 93, 230-239. Wyttenbach, C. R., and Thompson, S. C. (1985). The effects of the organophosphate insecticide malathion on very young chick embryos, malformations detected by histological examination. Am. J. Anatomy 174, 187-202. Xie, w., Stribley, J. A, Chatonnet, A, Wilder, P. J., Rizzino, A., McComb, R. D., Taylor, P., Hinrichs, S. H., and Lockridge, O. (2000). Postnatal developmental delay and supersensitivity to organophosphate in genetargeted mice lacking acetylcholinesterase. J. Pharmacol. Exper. Therapeut. 293(3), 896-902. Yamalik, N., Ozer, N., Caglayan, E, and Caglayan, G. (1990). Determination of pseudocholinesterase activity in the gingival crevicular fluid, saliva, and serum from patients with juvenile periodontitis and rapidly progressive periodontitis. 1. Dental Res. 69,87-89. Yeary, R. A., Eaton, J., Gilmore, E., North, B., and Singell, J. (1993). A multiyear study of blood cholinesterase activity in urban pesticide applicators. J. Toxieol. Environ. Health 39,11-25. Yuknavage, K. L., Fenske, R. A., Kalman, D. A., Keifer, M. c., and Furlong, C. E. (1997). Simulated dermal contamination with capillary samples and field cholinesterase biomonitoring. J. Toxicol. Environ. Health 51, 3555. Zajicek, 1. (1957). Studies on the histogenesis of blood platelets and megakaryocytes. Acta Physiol. Scand. 40(Suppl. 138), 1-32.
CHAPTER
49 Organophosphorus-Induced Delayed Neuropathy Marion Ehrich and Bernard S. Jortner Virginia-Maryland Regional College of Veterinary Medicine, Virginia Tech
49.1 HISTORY Neuropathy due to exposure to organophosphorus (OP) compounds was first reported in 1899, many years before this class of chemical agents was recognized for its insecticidal capabilities (Abou-Donia, 1981, 1995; Cherniack, 1986, 1988; Davis and Richardson, 1980; Ecobichon, 1994; Gallo and Lawryk, 1991; Goldstein et aI., 1988; Johnson, 1982; Lotti, 1992; Metcalf, 1984). Until the 1930s, however, organophosphorusinduced delayed neuropathy (OPIDN) appeared as isolated incidents and attracted little attention from the biomedical community. In the 1930s, between 4000 and 20,000 residents of the United States, which was under Prohibition at the time, were affected when a tricresyl phosphate-containing preparation was used as an alcohol substitute (Kidd and Langworthy, 1933). This product, called Ginger Jake, caused limb weakness and ataxia in exposed individuals from which they did not fully recover. Since the 1930s, other situations in which exposure to OP compounds has caused delayed neuropathy in humans have been identified; for relatively recent reviews, see Abou-Donia and Lapadula (1990) and Ecobichon (1994). Some of these have been isolated incidents, involving intentional or accidental exposures of individuals (Abou-Donia, 1981; Cherniack, 1988; Goldstein et aI., 1988; Lotti, 1992; Lotti et aI., 1986, 1995; Osterloh et aI., 1983). Others have resulted in the exposure of numerous people, such as those who consumed tricresyl phosphate-contaminated cooking oil in Europe, Africa, and Asia (Abou-Donia and Lapadula, 1990), those who used an adulterated abortifacient (Abou-Donia and Lapadula, 1990), and shoe-manufacturing workers in Italy exposed to an OP agent used as a plasticizer (Cavalleri and Cosi, 1980). These incidents put over 10,000 people at risk for OPIDN (Hierons and Johnson, 1978). Animals, too, can be susceptible to OPIDN following accidental exposures, and reports of OPIDN in water buffalo, cattle, horses, and sheep appear in the literature (Beck et al., 1977; EI-Sebae et aI., 1977; Perdrizet et aI., 1985; Sanders et aI., 1985). Handbook of Pesticide Toxicology Volume 2. Agents
Most of the recent victims of OPIDN have been people and animals exposed to OP agents that are used as lubricants and plasticizers, rather than insecticides. Current federal testing of OP compounds under the U.S. Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) requires that those proposed for use as insecticides be examined for their capability to cause OPIDN in relation to their capability to inhibit acetylcholinesterase, which is their insecticidal mechanism of action (EPA, 1991). Consequently, unless exposure is at concentrations that are above those necessary to cause signs of acute toxicity due to acetylcholinesterase inhibition, humans and susceptible animals are unlikely to develop OPIDN following insecticide exposure. Recent reports have suggested that OPIDN could contribute to symptoms seen in veterans of the Gulf War, based on selfreports provided by veterans (Haley et al., 1997). Soldiers on active duty were exposed to a variety of chemicals, including OP insecticides, pyrethrin insecticides, insect repellents, petroleum products, and sand dust. They were also at risk for exposure to OP compounds used as chemical warfare agents (potent acetylcholinesterase inhibitors), and, to prevent toxicity associated with such, the carbamate acetylcholinesterase inhibitor pyridostigmine was used to protect them preexposure. Soldiers were vaccinated against infectious diseases, anthrax, and botulinum toxin (NIB, 1994; President's Advisory, 1996). Investigations have determined that OP chemical warfare agents were released during this conflict; exposure levels were not determined but levels were too low to cause immediate physical symptoms. Although OPIDN appears weeks to months after exposure to some OP compounds, the OP insecticides used in the Persian Gulf and the OP chemical warfare agents are unlikely to cause this syndrome, especially in the absence of clinical signs of acute acetylcholinesterase inhibition (President's Advisory, 1996). Neurological examination of Gulf War veterans and animal studies do not appear to support the assertion that OPIDN is a clinical entity associated with soldiers that participated in the Gulf War (Abou-Doniaet aI., 1996; Haley and Kurt, 1997).
987
Copyright © 2001 by Academic Press. All rights of reproduction in any fonn reserved.
988
CHAPTER 49
Organophosphorus-Induced Delayed Neuropathy
49.2 CHEMISTRY OF
ORGANOPHOSPHORUS COMPOUNDS Not all organophosphorus (OP) compounds are capable of causing organophosphorus-induced delayed neuropathy (OPIDN). Lists of neuropathy-inducing OP compounds are contained in the 1991 edition of the Handbook of Pesticide Toxicology (Gallo and Lawryk, 1991) and other reviews of OPIDN (Abou-Donia, 1981; Abou-Donia and Lapadu1a, 1990; Cherniack, 1988; Hollingshaus, 1983; 10hnson, 1975b, 1982). Among the OP compounds responsible for reported incidents of OPIDN in man and animals or OP compounds used for laboratory studies of OPIDN are the protoxicants leptophos (O-4-bromo2,5-dichlorophenyl O-methyl phenylphosphorothioate) and triortho-cresyl phosphate (TOCP, also known as tri-ortho-tolyl phosphate, TOTP), which is metabolized to a cyclic saligenin phosphate responsible for inducing the OPIDN. Active OPIDN-inducing agents used for many experimental studies include mipafox (N, N' -diisopropy lphosphorodiamidic fluoride), diisopropyl phosphorofluoridate (DFP), and di-n-dibutyl-2,2dichlorovinyl phosphate (DBDCVP) (Fig. 49.1). None of these are currently used as insecticides. Several reports and reviews, including those listed previously and others (Abou-Donia, 1995; Davis and Richardson, 1980; Ecobichon, 1994; 10hnson et aI., 1991; Lotti, 1992; Richardson et aI., 1993; Wu and Casida, 1994; Yoshida et aI., 1994), have examined structureactivity relationships among compounds that cause OPIDN, with the following points noted:
CH 3
I
H3C-C~
N~
01-°:0-· OCH3
Leptophos
Cl
1'0
/p, /NH
F
H3 C - CH
I
CH 3
Mipafox
o
CH3CH~H~HP, 11 P-OCH=CCI2
CH3CH~H2CHP /
OFP
1. For OPIDN, sometimes referred to as Type I OPIDN (Abou-Donia and Lapadula, 1990), phosphorus must be in the pentavalent state. 2. The atom with the coordinate covalent bond attached to the phosphorus must be an oxygen; protoxicants with sulfur attached by a coordinate covalent bond can be oxidized to active neurotoxicants. 3. The neuropathy-inducing OP compounds all have at least one oxygen or amine bridge linking an R group to phosphorus. Therefore, the major subgroups producing OPIDN are either phosphates (derivatives of phosphoric acid, which has four oxygens on the phosphorus), phosphonates (derivatives ofphosphonic acid, which has three oxygens on the phosphorus), phosphoramidates (derivatives of phosphoramidic or phosphorodiamidic acids, with one or two nitrogens and two or three oxygens on the phosphorus), or phosphoroftuoridates (three oxygens and a fluoride on the phosphorus) (Fig. 49.2). 4. Alkyl substitution on an ortho site of phenyl phosphates increases the likelihood that the compound can be metabolized to a neurotoxic ant. Ortho methyl substitution rather than longer chain substitution on the phenyl ring(s) of triphenyl phosphates increases the capability to cause OPIDN. However, not all ortho-methyl-substituted phenyl phosphates induce neuropathy. Substitution at other sites on the ortho-substituted ring decreases the neurotoxicity.
Phenyl saligenin cyclic phosphate (PSP)
TOCP
OSOCVP
Figure 49.1 Chemical structures of organophosphorus compounds commonly used in laboratory studies. Included are tri-ortho-cresyl phosphate (TOCP, also known as tri-ortho-tolyl phosphate, TOTP), PSP, a cyclic saligenin phosphate metabolite congener of TOCP, leptophos, mipafox, diisopropyl phosphorofluoridate (DFP). and di-n-dibutyl-2,2-dichlorovinyl phosphate (DBDCVP).
5. Increasing the size of the alkyl substituents (up to four or five) on phosphoro- and phosphonofluoridates, phosphorodiamidofluoridates, and dichlorovinyl phosphates increases the hydrophobicity and increases the capability to cause OPIDN. 6. Chirality can contribute to neurotoxicity. Racemic mixtures tend to be less potent as inducers of OPIDN; for compounds tested to date, one enantiomer generally appears to be a more potent neuropathy-inducing agent than the other (Hollingshaus, 1983; 10hnson and Read, 1987; 10hnson et aI., 1991; Lotti et aI., 1995). Neuropathy-inducing OP compounds have a leaving group attached to a labile oxygen or nitrogen bond (Johnson, 1982). Dealkylation at this site results in a negatively charged phosphoryl group, and formation of this chemical species is needed if a phosphate or phosphonate is to induce classical delayed
49.3 Clinical Manifestations (Human)
989
49.3 CLINICAL MANIFESTATIONS (HUMAN)
Phosphates
Phosphonates
o
RHN,II
p-o-x
Phosphoramidates R-O"
o
RHN,II
,. P-o-x
Phosphorodiamidates
Descriptions of the clinical manifestations of OPIDN in humans can be found in several reviews (Abou-Donia, 1995; Cherniack, 1986; Ecobichon, 1994; Gallo and Lawryk, 1991). These descriptions follow a similar scenario. Some days (usually 6-14) after exposure, humans note tingling of the hands and feet, followed by sensory loss in the hands and feet. Electromyograms and nerve conduction studies indicate decreased firing of motor units and slowed motor conduction. It is, however, the appearance days to weeks after exposure of bilateral and symmetrical weakness progressing to flaccidity of the distal skeletal muscles of the lower and upper extremities that is characteristic of this disorder. Ataxia can be noted. Even though this may resolve with time, victims of OPIDN may still have abnormal reflexes and spasticity.
RHN
49.4 CLINICAL MANIFESTATIONS (ANIMAL) R
o -0,11 P-F
Phosphorofluoridates R -0/
Phosphinates
Figure 49.2 Basic structures of organophosphorus compounds. Phosphates, phosphonates, phosphoramidates, phosphorodiamidates, and phosphorofluoridates cause classical OPIDN (Type I OPIDN); phosphinates do not.
neuropathy [sometimes designated Type I OPIDN (AbouDonia and Lapadula, 1990)]. Deprotonation rather than dealkylation is thought to provide the necessary negative charge for neuropathy-inducing phosphoramidates (Richard son, 1995). This dealkylation occurs after the OP inhibits an esterase, specifically, neuropathy target esterase (NTE, also known as neurotoxic esterase). It has been suggested that another site on this enzyme, one other than the serine active site of NTE, is altered after this dealkylation occurs (John son, 1982). The NTE inhibition and the dealkylation reaction render the structure capable of essentially irreversibly binding to NTE, which is a necessary prerequisite to OPIDN (see Section 49.7). Another type of neuropathy induced by OP compounds, referred to as Type 11 OPIDN (Abou-Donia and Lapadula, 1990), has been reported for four trivalent phosphites [triphenyl phosphite and the tricresyl phosphites (0, rn, or p)]. The chemistry of their interaction(s) with esterases such as NTE has not been delineated in the detail provided for OP compounds that induce Type I OPIDN (Abou-Donia, 1995).
The adult hen is the recognized animal model for OPIDN (Abou-Donia, 1981; EPA, 1991). This is because clinical signs, which occur after a delay period similar to that which occurs in humans, are easy to observe as they progress, and the associated histopathologic changes are easily identified. In addition, the hen provides readily reproducible results in a relatively small, relatively accessible animal model. In the hen, a lag period of several days is needed before any clinical alterations appear. Early signs of OPIDN in this species are abnormal foot placement and leg weakness, which may affect balance. As OPIDN progresses, hens become reluctant to walk, show incoordination when they do, and may have wing droop and/or use wings for balance. More severe signs include loss of upright posture when walking, followed by loss of ability to walk. Eventually, the wings also become involved. These clinical deficits can be differentiated, using designations that range from mild effects to paralysis on a scoring system that ranges from 1 to 5 or from 1 to 8 (Sprague et aI., 1980) (Fig. 49.3). It is, however, only the adult chicken that shows these clinical signs upon exposure to neuropathy-inducing organophosphorus compounds. This manifestation of OPIDN does not seem to appear in chickens under 55-60 days of age (Funk et al., 1994a; Moretto et al., 1991). Clinical manifestations of OPIDN are seen in a variety of other species (Abou-Donia, 1981; Johnson, 1982), including other avians such as the pheasant and turkey (Johnson, 1982; Larsen et al., 1986). Humans (see the previous section) and mammals such as sheep, pigs, cattle, water buffalo, horses, cats, dogs, and ferrets are susceptible to clinical manifestations of OPIDN, as indicated by progressive ataxia (Abou-Donia, 1995; Johnson, 1982; Jortner et al., 1983; Stumpf et aI., 1989; Tanaka et aI., 1994). Clinical manifestations specifically associated with OPIDN are not, however, generally seen in laboratory
990
CHAPTER 49
Organophosphorus-Induced Delayed Neuropathy
brain and spinal cord lesions, these are not two types of OPIDN, but represent separate categories of organophosphorus-induced neurotoxicity (Lehning et aI., 1996).
4
~
49.5 NEUROPATHOLOGY
3
... 0
I/)
a;
...
~E
(3
2
~~Ii~----'-----~------------~--------~ 10 12 14 18 21 Days af1er Organophospha1e
Figure 49.3 Development of clinical signs in chickens after administration of phenyl saligenin phosphate (PSP) and tri-ortho-tolyl phosphate (TaTP). Results are presented as the mean ± SD, n = 3-9, on a scale of 1-5. PSP im 2 mg/kg (0-0), 3 mg/kg (0-0), IO mg/kg (6-6); TOTP 360 mg/kg po (e-e), 500 mg po (_ _ ). Increasing clinical scores reflect progression of deficits. Reproduced with permission from Jortner, B. S., and Ehrich, M. (1987). Neuropathological effects of phenyl saligenin phosphate in chickens. NeuroToxicology 8, 303-314. Copyright © 1987 by Intox Press, Little Rock, AR.
rodents (Ehrich et aI., 1993b, 1995; Lehning et aI., 1996; Padilla and Veronesi, 1985; Somkuti et aI., 1988; Veronesi et aI., 1991), although they have been reported to appear in rats over 6 months old or mice dosed for over 200 days (Lapadula et aI., 1985; Moretto et aI., 1992b). There have been suggestions that there are actually two types of OPIDN-OPIDN induced by pentavalent phosphorus compounds (phosphates, phosphonates, and phosphoramidates; Type I OPIDN) and OPIDN that can induced by phosphites (trivalent phosphorus compounds, with most studies done with triphenyl phosphite; Type 11 OPIDN) (Abou-Donia and Lapadula, 1990). Triphenyl phosphite has been reported to cause ataxia in species both susceptible to (hens, ferrets) and relatively resistant to (rats, Japanese quail) Type I OPIDN. The latent period is, however, shorter than that of Type I OPIDN, as seen in hen studies (Carrington et aI., 1988). In contrast to the minimal clinical neuropathy induced by pentavalent phosphorus compounds in rats, triphenyl phosphite elicited prominent clinical manifestations, such as hyperexcitability, spasticity, tail-kinking, side-to-side movements, circling, ataxia, and flaccid paralysis in these rodents (Lehning et aI., 1996; Veronesi and Dvergsten, 1987). Pathological manifestations of phosphite neurotoxicity, too, differ from those characteristic of OPIDN (see Section 49.5) (Abou-Donia and Lapadula, 1990; Lehning et aI., 1996). We have used the terms Type I and Type 11 OPIDN in this review. Some investigators feel that based on differences in species susceptibility, magnitude of NTE inhibition, onset and nature of clinical signs, and extent and nature of
Neuropathologic studies of experimental OPIDN induced by pentavalent phosphorus compounds (Type 1 OPIDN) have revealed a consistent pattern of lesions, which is felt to represent the morphologic substrate of the entity. The test compounds most often used to elicit these lesions were tri-ortho-cresyl (or tolyl) phosphate (TOCP), its neurotoxic cyclic congener phenyl saligenin phosphate (PSP), diisopropyl phosphorofluoridate (DFP), or mipafox, with the chicken and cat being the major experimental animal subjects (Bischoff, 1967, 1970; Cavanagh, 1954, 1964; Cavanagh and Patangia, 1965; Illis et aI., 1966; Itoh et aI., 1984; Jortner and Ehrich, 1987; Krinke et aI., 1979; Prineas, 1969; Tanaka and Bursian, 1989). This body of work reveals that the primary lesion is a bilateral degenerative change in distal levels ofaxons and their terminals, primarily affecting larger/longer myelinated central and peripheral nerve fibers, leading to breakdown of affected neuritic segments and secondarily of their myelin sheaths. These lesions generally begin to develop at or near the end of the postdosing symptom-free period. In experimental studies of chickens, this distal pattern of injury is manifest by clinical neuropathy associated with bilateral central nervous system long-tract involvement, such as in cervical spinal cord, medullary, and cerebellar levels of the ascending spinocerebellar tracts and fasciculus gracilis, and lumbar levels of the descending medial pontine spinal tracts (Abou-Donia and Preissig, 1976a, 1976b; Cavanagh, 1954; Classen et aI., 1996; Itoh et aI., 1984; Jortner and Ehrich, 1987; Tanaka and Bursian, 1989; Tanaka et aI., 1990a) (Fig. 49.4). The most sensitive histological indicator of OPIDN was felt to be degenerating cerebellar fibers, especially in folia IV and V (Classen et aI., 1996). Use of Fink-Heimer silver impregnation histological techniques in hens dosed with TOCP or DFP revealed more extensive distribution of central nervous system axonal and terminal degeneration, particularly extension of alterations of the lumbar medial pontine spinal tract into ventral gray matter laminae VI and VII and those of the rostral spinocerebellar system, which were seen in the deep cerebellar nuclei and mossy fiber projections to the anterior lobe of the cerebellar cortex (Tanaka and Bursian, 1989; Tanaka et aI., 1990a). Terminal and preterminal degeneration was also noted in a number of medullary structures such as the spinal lemniscus and lateral vestibular, inferior olivary, gracile, external cuneate, and lateral cervical nuclei. Degenerating presynaptic boutons and small axons (Bischoff, 1970; Dyer et aI., 1996) are the likely ultrastructural substrate of these silver-impregnated altered gray matter neurites in hens. In addition to these central nervous system lesions of OPIDN, distal regions of long peripheral nerve myelinated fibers in chickens, particularly in the legs, are similarly affected
49.5 Neuropathology
991
Figure 49.4 Cross section of a spinocerebellar tract in the cervical level of spinal cord from a hen 21 days after exposure to a neurotoxic dose of mipafox (30 mglkg ip). The section shows extensive myelinated fiber degeneration, manifest by pale-staining swollen axons (arrow) or dark-staining fibers with collapsed axons and disordered myelin sheaths (arrowhead). Toluidine blue-safranin stain; bar = 25 JllIl.
(Cavanagh, 1954; Dyeretal., 1991, 1992; EI-Fawaletal., 1988, 1990b, 1990c; Jortner, 1984; Jortner and Ehrich, 1987; Prineas, 1969). Attention has been directed to the sensitivity of the nerve supplying the biventer cervicis muscle and the large-diameter myelinated fibers of the tibial nerve branch to the lateral head of the gastrocnemius muscle in this species (EI-Fawal et aI. , 1988, 1990c; Krinke et aI., 1979). A similar neuropathologic pattern has been noted in cats exposed to neurotoxic doses of DFP or TOCP. This is manifest by distal degeneration mainly affecting large, long myelinated fibers, involving rostral (cervical, medullary, and, for some, cerebellar) levels of ascending tracts (fasciculus gracilis, spinocerebellar tracts, spino-olivary tract) and thoracolumbar spinal cord levels of descending tracts (corticospinal, reticulospinal, rubrospinal) (Abou-Donia et aI. , 1986; Cavanagh and Patangia, 1965). Degenerating nerve fiber terminals were seen in gray matter afferent nuclei (Cavanagh and Patangia, 1965; Illis et aI., 1966; Prineas, 1969). Distal levels of long peripheral nerve myelinated fibers and their terminals, such as in the hind legs and in recurrent laryngeal nerves, were similarly affected (Bouldin and Cavanagh, 1979a, 1979b; Ca-
vanagh, 1964; Drakontides et aI., 1982; Glazer et aI., 1978; Prineas, 1969). By light microscopy, the qualitative changes were best demonstrated in sections from epoxy resin--embedded preparations of spinal cord and medullary white matter, and peripheral nerve. These included axonal swelling with attenuated myelin sheaths (Figs. 49.5-49.7). The swollen axons were pale staining or debris laden and dark staining (Figs. 49.5-49.8). Another feature, common in central nervous system myelinated tracts, was the presence of contracted dark-staining axons with disordered myelin sheaths (Figs. 49.4 and 49.5). These lesions increased in affected regions as the clinical neuropathy advanced. Later lesions included fragmentation of affected fiber segments as the process of Wallerian-like degeneration ensued (Figs. 49.8 and 49.9). This was associated with formation of myelin-rich ovoids and phagocytosis of the degraded element by macrophages or Schwann cells (the latter in peripheral nerve) (Figs. 49.8 and 49.9). Ultrastructurally, changes in axonal membrane systems were early morphologic events in OPIDN. This includes proliferation of intra-axonal tubules and cisterns, resembling smooth endo-
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CHAPTER 49 Organophosphorus-Induced Delayed Neuropathy
Figure 49.5 High-power view of spinal cord long-tract degeneration in cross section from a chicken given mipafox (30 mg/kg ip) 21 days earlier shows a large central swoIIen axon with moderately dense staining axoplasm containing particulate material and a vacuole. T he myelin sheath is thin. Many of the adjacent degenerating fibers have contracted, dark-staining axons with disordered dark-staining myelin sheaths (arrowhead). Pale-staining regions of myelinated fiber loss with associated gliosis are present (asterisk). Toluidine blue-safranin stain; bar = 10 >lm.
plasmic reticulum, and vesicles in hens and cats receiving toxic doses of TOCP or DFP (Bischoff, 1967, 1970; Bouldin and Cavanagh, 1979a, 1979b; Le Vay et al., 1971; Prineas, 1969) (Fig. 49.10). Other early changes, seen in cats dosed with DFP, were peripheral myelinated fiber distal, nonterminal varicosities due to the presence of abnormal membrane-lined vacuoles in axons, inner myelin sheaths, or both, which preceded fiber degeneration (Bouldin and Cavanagh, 1979a, 1979b). These workers suggest such vacuolar alterations represent a "chemical transection" of the fiber, leading to its subsequent breakdown. Following these early ultrastructural events, a variety of subsequent degradative axonal changes are seen, progressing to fiber degeneration. One sequence involves axonal swelling with secondary attenuation of the myelin sheath. Electron microscopic study revealed that numbers of these swollen axons contained disorganized masses of normal and altered mitochondria, cytoskeletal components (neurofilaments), lysosome-like dense bodies, and membranous multilamellar bodies (Bischoff,
1967,1970; Bouldin and Cavanagh, 1979b; Jortner and Ehrich, 1987; Prineas, 1969) (Fig. 49.11). A second appearance of the swollen axons, which may be derived from the preceding, is one in which there has been granular degeneration of its contents due to lysis of the cytoskeleton and other axonal contents, leading to swollen electron-lucent axons (Jortner and Ehrich, 1987) (Fig. 49.12). This is thought to be an advanced manifestation of the neuropathy (Prineas, 1969) and may be related to increased activity of calcium-activated proteinases, associated with toxicant-induced intra-axonal elevations of calcium ions (EI-Fawal et al., 1990a). Yet a third ultrastructural appearance of degenerating fibers, particularly prominent in the central nervous system, is distal axonal collapse with illdefined electron-dense axoplasm and disordered myelin sheaths (Fig. 49.11). These axonal changes are associated with aggregation of membranous masses, altered mitochondria, and dense bodies in degenerating axon terminals (Drakontides et aI., 1982; Glazer etal., 1978; Prineas, 1969). Some workers suggest that axon terminals are the initial site of degeneration, with sub-
49.5 Neuropathology
Figure 49.6 Cross-sectioned distal (dorsal metatarsal nerve) peripheral nerve from a chicken given 2 mglkg of phenyl saligenin phosphate 14 days earlier. Fiber degeneration is manifest by swollen, pale-staining axons and thin myelin sheaths. Toluidine blue-safranin stain; bar = 100 ;4m. Reproduced with permission from Jortner, B. S., and Ehrich, M. (1987). Neuropathological effects of phenyl saligenin phosphate in chickens. NeuroToxicology 8, 303-314. Copyright © 1987 by Intox Press, Little Rock, AR.
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Figure 49.7 Longitudinal section of a dorsal metatarsal nerve from a chicken 21 days after exposure to a toxic dose (10 mglkg) of phenyl saligenin phosphate, showing dark-staining axonal debris in the paranodal region of a large myelinated fiber. Toluidine blue-safranin stain; bar = 10 ;4m. Reproduced with permission from Jortner, B. S., and Ehrich, M. (1987). Neuropathological effects of phenyl saligenin phosphate in chickens. NeuroToxicology 8, 303-314. Copyright © 1987 by Intox Press, Little Rock, AR.
Figure 49.8 This tangentially sectioned tibial nerve branch to the lateral head of the gastrocnemius muscle from a chicken 15 days after dosing with 2.5 mg/kg of phenyl saligenin phosphate shows several stages of myelinated nerve fiber degeneration (arrows). These include swollen axons with pale or moderate staining of their contents and formation of myelin-rich segments (ovoids) of Wallerian-like degeneration. Toluidine blue-safranin stain; bar = 25 !im.
sequent involvement of terminal portions of the axon, creating a true "dying-back" neuropathy (Tanaka and Bursian, 1989). This contrasts with another view, that the terminal lesions are secondary to injury in the distal, non terminal portions of the axon (Bouldin and Cavanagh, 1979a, 1979b; Prineas, 1969).
Eventually, the degrading segments of affected myelinated fibers are phagocytized, an event occurring more rapidly in peripheral than central regions of the nervous system (Fig. 49.13). In the former, most fiber debris was phagocytized and degraded by 4 weeks following a single toxic dose in hens, but these dam-
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aged neurites may persist much longer in the central nervous system (Jortner et aI., 1989). In peripheral nerve, the damaged fiber swells, fragments, and is phagocytized by macrophages or Schwann cells within the column formed by the original Schwann cell basal lamina. This resembles events in Wallerian
.. Figure 49.9 Teased peripheral nerve myelinated fiber from the tibial nerve branch to the lateral head of the gastrocnemius muscle from a chicken dosed with 10 mglkg phenyl saligenin phosphate 14 days earlier. The nerve is fragmented into myelin ovoids typical of the Wallerian-like degeneration of OPIDN. Osmium tetroxide stain; bar = 100 flm. Reproduced with permission from Jortner, B. S., and Ehrich, M. (1987). Neuropathological effects of phenyl saligenin phosphate in chickens. NeuroToxicology 8, 303-314. Copyright © 1987 by Intox Press, Little Rock, AR.
(hence the term Wallerian-like) degeneration. With advanced breakdown of the fiber in OPIDN, proliferation of Schwann cells in their basal lamina forms the band of Biingner. The latter is a site of subsequent nerve fiber regeneration, which is robust in OPIDN following a single dose of the toxic ant (Jortner et aI., 1989). The bands of Biingner (columns of proliferating Schwann cells) provide an appropriate structural and growth-enhancing environment to permit re-innervation to occur. This regeneration included replacement of degenerated peripheral axon terminals as well (Glazer et aI., 1978; Illis et aI., 1966). Consistent with other forms of nerve fiber degeneration, there is a failure of such axonal regeneration in the central nervous system in OPIDN (Jortner et aI., 1989). The prominence of spinal cord lesions relative to those in sciatic nerve of hens sacrificed 1 month or more following dosing with OPIDN-inducing insecticides was attributed to peripheral nerve regeneration (Abou-Donia and Graham, 1978; Abou-Donia and Preissig, 1976a, 1976b; Abou-Donia et aI., 1979). In the central nervous system, macrophages provide the phagocytic element acting on degraded myelinated fibers, and there is prominent
Figure 49.10 Electron micrograph showing early ultrastructural axonal changes of OPIDN (hen given PSP 2.5 mg/kg im). An increase in membrane-lined tubules and cisterns suggestive of proliferation of agranular endoplasmic reticulum is noted (straight arrow). In addition, there is a focus of granular degeneration of the cytoskeleton (arrowhead) and a small lamellar body (curved arrow). Bar = 1 flm.
49.5 Neuropathology
astrocytic proliferation in the damaged levels of spinal cord and brainstem (Jortner et al., 1989) (Figs. 49.5 and 49.13). In contrast to the lesions induced by pentavalent organophosphorus compounds (the Type I OPIDN described previously), another morphologic pattern of the neuropathy is induced by exposure to trivalent organophosphates. This has been designated Type 11 OPIDN (Abou-Donia, 1995; Abou-Donia and Lapadula, 1990) and is produced by aryl phosphites, mainly triphenyl phosphite, in monkeys, cats, rats, and chickens. The histological lesions have best been described in the latter two species. Those in the rat consist of delayed-onset, dose-related bilateral alteration of myelinated fibers and neuronal cell bodies. The fiber lesions have qualitative similarity to those of Type I OPIDN, in that intra-axonal aggregates of agranular reticulum and masses of tubulovesicular elements are seen, and the lesions progress over time to fiber degeneration (Veronesi and Dvergsten, 1987). The extent and distribution of these lesions differ from those of Type I OPIDN in rats (see the following discussion). In the brainstem, in addition to involving the spinocerebellar tracts, lesions were seen in the medial and lateral reticular formation, medial longitudinal fascicu-
995
Ius, and medial vestibular nucleus. Spinal cord white matter fiber degeneration involved the ventral and ventrolateral funiculi housing the spinocerebellar, spinothalamic, tectospinal, and reticulospinal tracts, at all levels. They were also noted in several gray matter laminae. In contrast, the prominent involvement of the rostral (cervicomedullary) levels of the fasciculus gracilis seen in Type I OPIDN was absent. Neuronal cell body lesions of chromatolysis and necrosis were a prominent feature of Type IT OPIDN, being seen in the medial vestibular and lateral reticular nuclei and, on occasion, in the lumbosacral gray matter, associated with nerve fiber degeneration (Veronesi and Dvergsten, 1987). Such neuronal lesions are not a feature of the Type I form of the neuropathy in rats. Type 11 OPIDN has also been induced in chickens by a single 1000-mglkg dose of triphenyl phosphite (Carrington et al., 1988). This avian model is characterized by neuronal cell body chromatolysis and necrosis in the anterior horns of the spinal cord gray matter and dorsal root ganglia. In addition, myelinated fiber degeneration involving the cerebellar peduncle, reticular formation, ventral and lateral spinal cord white
Figure 49.11 Electron micrograph demonstrating extensive myelinated fiber degeneration in a cross section of the medullary (distal) level of the fasciculus gracilis from a chicken administered a toxic dose (500 mg/kg po) of tri-ortho-tolyl-phosphate 21 days earlier. The morphologic presentations of fiber degeneration include swollen axons with multilarnellar membranous aggregates (arrow) and contracted electron-dense axons with disordered myelin sheaths (arrowhead). Bar = 5 ~m.
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Figure 49.12 Electron micrograph demonstrating one pattern of advanced axonal alteration. This is manifest by extensive granular degeneration of the axoplasm and axonal organelles in this cross-sectioned myelinated fiber from the dorsal metatarsal nerve of a chicken dosed with 10 mg/kg of phenyl saligenin phosphate 14 days earlier. Bar = 1 >UTI.
matter funiculi, spinal cord gray matter, and peripheral nerve is seen (Abou-Donia and Lapadula, 1990; Carrington et aI., 1988). This spectrum of lesions is somewhat different from that induced in chickens by pentavalent organophosphorus compounds, as described previously. The Fink-Heimer silver impregnation method was also employed to study the lesions of Type II (triphenyl phosphite induced) OPIDN in chickens (Tan aka et aI., 1992). This procedure revealed extensive evolving terminal and axonal degeneration, prominent in the spinal cord gray and white matter, selected medullary nuclei and fiber tracts, cerebellar granular layer in folia I-VI, and several midbrain and forebrain regions. Major systems affected included descending brainstem-spinal pathways originating in the lateral vestibular nuclei, mossy fiber afferents to the anterior lobe of the cerebellum, afferents to the lateral mesencephalic nuclei, and tracts associated with the basal ganglia and its optic tectal pathways. This represented more widespread injury than seen in Type I OPIDN, in that higher order centers responsible for processing and integrating sensorimotor, visual, and auditory information were involved (Tanaka et aI., 1992).
Neuropathological studies of triphenyl phosphite-induced delayed neuropathy in the rat, Japanese quail, and ferret have utilized the Fink-Heimer technique (Lehning et aI., 1996; Stumpf et aI., 1989; Tanaka et aI., 1990b, 1991; Varghese et aI., 1995). Although there were some species variations in affected tracts and nuclei, lesions of axonal and terminal degeneration were noted at multiple levels of the central nervous system and included thalamic and cerebral cortical (in mammals) regions in these studies. As with chickens (discussed previously), these included higher order centers.
49.6 NEUROPATHOLOGY OF MAMMALIAN ANIMAL MODELS The most reliable experimental animal model of OPIDN is obtained by single or multiple dosing in the domestic adult chicken (hen), and the spectrum of nervous system alterations in that species has been documented previously in some detail. A good deal of earlier work has employed the domestic cat, which, along with other susceptible mammalian species (sheep,
49.6 Neuropathology of Mammalian Animal Models
997
Figure 49.13 This electron micrograph shows a cross-sectioned myelinated fiber in the medullary level of the fasciculus gracilis from a chicken 21 days after a neurotoxic dose of tri-ortho-tolyl-phosphate (500 mglkg po). The degenerated axonal contents have been phagocytized by a macrophage (arrow) within the myelin tube. Processes of reactive astrocytes are seen in adjacent tissue (asterisk). Bar = 2 !-lm.
cattle, nonhuman primates, etc.), has a qualitative and distributive pattern of lesions that follows that of the hen (Abou-Donia, 1981; Cavanagh and Patangia, 1965; Jortner, 1984; Jortner et aI., 1983; Prineas, 1969). Thus, the basic nature of changes in these mammalian species has already been considered. Due to concern about regulatory reliance on an avian experimental model, there has been considerable interest in the evaluation of laboratory rodents as potential model systems for OPIDN. Most of this attention has focused on the laboratory rat (Jortner, 1988). In a series of studies in the Long-Evans strain of rat, using several dosing paradigms of TOCP, bilateral distal myelinated fiber alterations were demonstrated (Padilla and Veronesi, 1985; Veronesi, 1984). These involved distal levels of long spinal cord myelinated tracts, including ascending and descending spinal cord tracts (demonstrated in cervical levels of fasciculus gracilis and dorso- and ventrolateral white matter columns and in lumbar dorsolateral, ventrolateral, and ventral columns, respectively). This largely recapitulated the distribution of lesions seen in hens (Cavanagh, 1954; Jortner and Ehrich, 1987). The fiber lesions developed after a postdosing
latent period and were associated with a transient postdosing inhibition of whole brain neuropathy target esterase (NTE). No definitive clinical deficits were observed in affected rats. The lesions were seen in larger myelinated nerve fibers and consisted of giant axonal swellings containing accumulations of tubulovesicular profiles of smooth endoplasmic reticulum and/or vacuoles, progressing to massive accumulations of mitochondria, vesicular profiles, and amorphous bodies in granular axoplasm (Veronesi, 1984). The swellings were associated with attenuation of their myelin sheaths. Other pathological studies of rats with OPIDN, using a variety of strains (Wistar, Long-Evans, Sprague-Dawley) and TOCP or mipafox as toxicants, did not demonstrate this extensive distribution of myelinated nerve fiber lesions (Carboni et aI., 1992; Dyer et aI., 1992; Ehrich et aI., 1993a, 1993b, 1995; Inui et aI., 1993; Itoh et aI., 1985; Lehning et aI., 1996). The lesions were, in fact, primarily found in the medullary and cervical (distal) levels of the fasciculus gracilis and its afferent target nucleus, and axonal vacuolization and swelling were prominent features (Carboni et aI., 1992; Dyer et al., 1992) (Fig. 49.14). Studies by Veronesi and colleagues using mipafox
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Organophosphorus-Induced Delayed Neuropathy
Figure 49.14 Medullary (terminal) level of the fasciculus gracilis from a Long-Evans rat that had been given a toxic dose (30 mglkg ip) of mipafox 21 days earlier. There are numerous pale-staining swollen axons seen bilaterally CA). Higher power of the lesion (B) shows associated thin myelin sheaths. Toluidine blue-safranin stain.
in rats demonstrated a similar distribution of lesions (Veronesi et al., 1986). The fate of affected fibers in this rodent model is still in question, in that one study suggests they are reversible, whereas another indicates progression to degeneration and loss
of affected segments (Carboni et aI., 1992; Itoh et aI., 1984). One of the problems in evaluating neuropathic changes of this nature in the rostral fasciculus gracilis is the presence of some swollen, vacuolated axons in this region in normal rats (Car-
49.7 Pathogenesis boni et aI., 1992; Eisenbrandt et aI., 1990; Veronesi et aI., 1986). In addition, rats generally required higher dosages of toxicant to elicit pathological changes than did chickens (Ehrich et aI., 1993a, 1995). These findings, clinical insensitivity and restricted lesions, limit the utility of the laboratory rat as a model for OPIDN. Neuropathologic studies of a putative OPIDN model in laboratory mice have been done in animals dosed with 5803480 mg/kg TOCP (Veronesi et aI., 1991). Lesion distribution was not consistent with that seen in other models of the neuropathy, although myelinated fibers had qualitatively the same axonal pathology. The latter included intra-axonal vacuoles, neurofilament masses, and floccular degeneration, which were noted in cervical spinal cord white matter, medullary inferior olivary nucleus, and the fasciculus gracilis. The lesions were not associated with a specific pattern of toxicant-induced inhibition of neuropathy target esterase (NTE). The mouse, therefore, is not considered to be an acceptable model of OPIDN. Several studies of OPIDN in the ferret have brought to light another animal model of this neuropathy (Stumpf et aI., 1989; Tanaka et aI., 1991). These utilized tri-ortho-tolyl phosphate administered either orally or dermally (250, 500, or 1000 mg/kg) or DFP (2 or 4 mg/kg) given subcutaneously. The former compound was more toxic via the dermal route and, indeed, had little effect orally. Tri-ortho-tolyl phosphate-dosed animals in the 1000-mg/kg dermal group had 46% inhibition of brain neurotoxic esterase and associated clinical signs of rear leg ataxia and weakness progressing over the 11- to 49-day postdosing period. These were seen in a dose- and lesionrelated fashion. Spinal cord lesions of bilateral, degenerating axons in the cervical level of the fasciculus gracilis and lumbar regions of the dorsal spinocerebellar tract, and, on occasion, in the lumbar ventral gray matter were noted. Because these lesions were delineated using the Fink-Heimer silver impregnation method, no qualitative histological details of the lesions were detectable. The DFP-induced neuropathy in ferrets was associated with a dose-related wider spectrum of lesions, along with associated clinical deficits and inhibition of medullary/cerebellar neurotoxic esterase (Tanaka et aI., 1991). A single dose of 4 mg/kg inhibited this esterase by 86% at 4 h, with recovery of activity by 4 days. There was more prolonged acetylcholinesterase inhibition as well. These animals had rear leg ataxia, which progressed to paralysis in the 14- to 28-day postdosing period (the study was terminated at that point). Lesions detected by the Fink-Heimer method included dense axonal and terminal degeneration seen in the animals sacrificed at 21 and 28 days. In the spinal cord, these were seen in laminae VI-VII of the gray matter, in the ventral motor nucleus at the cervical level, and (as axonal degeneration) in the lateral corticospinal tract at the lumbar levels, and fasciculus gracilis and dorsal spinocerebellar tract in the cervical region. The medulla has both axonal and terminal alterations in the gracile, inferior vestibular, lateral reticular, and inferior olivary nuclei, as did the cerebellar folia of the anterior lobe. These studies indicate that the ferret has the spectrum of lesions, along with detectable clinical signs and ap-
999
propriate inhibition of brain esterases, to make it an appropriate model of OPIDN, although it has other limitations (availability, husbandry concerns) regarding its use as a standard test animal for safety assessment.
49.7 PATHOGENESIS Factors important in the development of organophosphorusinduced delayed neuropathy (OPIDN) have been discussed in a number of previous reviews (Abou-Donia, 1981, 1995; Abou-Donia and Lapadula, 1990; Carrington, 1989; Cherniack, 1986; Cranmer and Hixson, 1984; Davis and Richardson, 1980; Ecobichon, 1994; Ehrich, 1996; Gallo and Lawryk, 1991; Hollingshaus, 1983; Johnson, 1975a, 1982, 1992, 1993; Jortner, 1984,1988; Lotti, 1992, 1995, 1997; Lotti et aI., 1993; Richardson, 1995; see Chapter 50). A number of events that occur during the development of the neuropathy have been identified, yet the precise mechanism(s) remains elusive. One factor known to be important in the initiation of OPIDN is inhibition of neuropathy target esterase (NTE, also known as neurotoxic esterase). This enzyme has been suggested to be the molecular target of neuropathy-inducing OP compounds with a pentavalent phosphorus atom, yet recent evidence has suggested that NTE may be a biomarker rather than the single, specific target that initiates OPIDN (Ehrich, 1996; Johnson, 1982; Lotti, 1992; Lotti and Moretto, 1993; Lotti et aI., 1993). What is certain, however, is that NTE must be phosphorylated and extensively inhibited by the OP compound before notable OPIDN develops. In addition to inhibiting NTE, the OP compounds inducing delayed neuropathy must bind sufficiently strongly to the NTE that it is difficult to impossible to remove them (Clothier and Johnson, 1979; Johnson, 1980, 1982; Milatovic and Johnson, 1993; NostrandtandEhrich, 1993; Richardson, 1995). For this type of binding to occur, the pentavalent OP compound (phosphate, phosphonate, phosphoramidate) must have a leaving group whose removal or rearrangement results in a negatively charged moiety on the enzyme (John son, 1982, 1992, 1993). This process has been called "aging." Whether typical "aging" occurs with phosphoramidates has been debated (Johnson, 1993; Johnson and Safi, 1993; Milatovic and Johnson, 1993; Richardson, 1995), but, in any case, a strong attachment of OP compound to NTE must be part of the process between exposure and development of OPIDN. Although a 70% threshold for NTE inhibition was once thought necessary before ataxia and nervous system degeneration could occur, OPIDN induced by OP compounds with pentavalent phosphorus atoms actually follows a dose-response curve (Ehrich et aI., 1993a, 1995). The correlation between early NTE inhibition and OPIDN is sufficiently strong that registration of OP insecticides under the U.S. Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA) requires that data on NTE inhibition and on doses required for its inhibition relative to acetylcholinesterase inhibition be obtained as a biochemical determinant of potential to cause OPIDN (Ehrich, 1996; EPA, 1991).
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CHAPTER 49
Organophosphorus-Induced Delayed Neuropathy
Neuropathy-inducing phosphites also inhibit NTE (AbouDonia, 1995; Abou-Donia and Lapadula, 1990), yet they do not have a leaving group as do the OP compounds with a pentavalent phosphorus atom. This distinction, as well as differences in clinical and pathological presentation (see Section 49.5), place the phosphite-induced disorder (Type II OPIDN) in a different category than classical OPIDN (Lehning et aI., 1996; Varghese et aI., 1995; Veronesi and Dvergsten, 1987). To say that NTE is or is not the precise target for initiation of neuropathy induced by OP compounds is difficult. The physiological functions of NTE are unknown. It is a carboxy lesterase, a class of enzymes responsible for the hydrolysis of a wide variety of compounds, including OP compounds, carboxylesters, thioesters, and aromatic amides. Its molecular weight is in the range of 150,000 Da. This enzyme is found in brain, spinal cord, peripheral nerves, and in nonnervous tissue, such as the adrenal gland, lymphocytes, and platelets (Bertoncin et aI., 1985; Dudek and Richardson, 1982; Ehrich, 1996; Maroni and Bleecker, 1986; Schwab and Richardson, 1986). Suggestions have been made that activity in the peripheral nerve has a differential sensitivity than tissue from the central nervous system and that this tissue may be more predictive of potential for OPIDN (Barril et aI., 1988; Caroldi and Lotti, 1982; Carrera et aI., 1992; Lotti et aI., 1987; Moretto et aI., 1989). The nerve used for assay, however, is the sciatic nerve, which does not show exceptional pathogenic effects, except in its distal branches, following administration of neuropathy-inducing OP compounds (Dyer et aI., 1991). Furthermore, the sciatic nerve is relatively small and more difficult to use for assay than brain without providing generally recognized additional benefit (Barril et aI., 1988; Correll and Ehrich, 1991; Ehrich, 1996; Johnson, 1982). Purification of NTE has been a long and difficult process, and no studies examining the effects of a purified enzyme on cellular functions have yet been reported. NTE is a membraneassociated enzyme, and early attempts to purify the protein resulted in loss of activity when attempts were made to separate the enzyme protein from the membrane phospholipid (Clothier and Johnson, 1979; Davis and Richardson, 1987; Davis et aI., 1980; Ishakawa et aI., 1983; Pope and Padilla, 1989a, 1989b; Ruffer-Tumer et aI., 1992; Schwab et aI., 1985; Thomas et aI., 1989, 1990; Williams and Johnson, 1981). Early studies with brain tissue suggested NTE could exist in multiple forms (Chemnitius et al., 1983; Zech and Chemnitius, 1987), but later studies determined that such was not the case (Carrington and Abou-Donia, 1986; Richardson, 1992), except, perhaps, in peripheral nerve, where recent studies suggest that soluble and membrane-associated enzymes exist in that tissue (Carrera et aI., 1994; Escudero et aI., 1997; Tormo et aI., 1993; Vilanova et aI., 1990, 1993). Only recently have reports appeared that suggest an active enzyme could be purified (Glynn et aI., 1993, 1994; Johnson and Glynn, 1995; Mackay et aI., 1996; Seifert and Wilson, 1994; Thomas et aI., 1993), cloned (Glynn et aI., 1999; Lush et aI., 1998), and localized in neuronal cell bodies (Glynn et aI., 1998). Purification, characterization, and physiological effects
of OP phosphorylation and aging of NTE may be aided by use of new, very potent probes and sensitive substrates (Borhan et aI., 1995; Wu and Casida, 1994, 1995; Yoshida et aI., 1994, 1995). NTE, its biochemistry, enzymology, and isolation are discussed in more detail in Chapter 50. Although early and significant inhibition of NTE is an excellent predictor of potential for developing OPIDN, the relationship between NTE inhibition and OPIDN itself is less clear. These items have been summarized in recent reviews (Ehrich, 1996; Lotti, 1992; Lotti et aI., 1991, 1993; Richardson, 1992; see Chapter 50). For example, although NTE is inhibited shortly after OP administration, it may no longer be inhibited some days to weeks later when clinical and pathological manifestations of OPIDN are evident. If NTE were a target, its inhibition would need to initiate some series of events that no longer required its inhibition. Furthermore, NTE inhibition predictive for OPIDN is usually measured in brain and spinal cord, tissues that have a relatively low proportion of their nerve fiber population affected. The most concentrated nerve fiber lesions of OPIDN are found in distallevels of peripheral nerves, such as branches of the tibial or in the biventer cervicis in the hen, and these tissues are too small for measurement of NTE activity (El-Fawal et aI., 1988, 1990c; Jortner and Ehrich, 1987). Another consideration with the designation of NTE as the specific target for initiation of OPIDN is that this enzyme in brain, spinal cord, and/or peripheral nerve can be inhibited in species and age groups of animals that do not show notable clinical signs or pathological manifestations of OPIDN (e.g., chicks, rats, mice, and, possibly, quail [Ehrich, 1996; Funk et al., 1994a; Padilla and Veronesi, 1985, 1988; Peraica et aI., 1993; Varghese et aI., 1995; Veronesi et aI., 1991]). It appears, for instance, that chickens need to be 55-60 days of age before clinical signs indicative of OPIDN appear. Some pathological evidence of OPIDN, however, can be seen in the spinal cord of chicks that were only 2 weeks old when exposed to diisopropyl phosphorofluoridate (DFP), suggesting that age susceptibility in this animal model of OPIDN needs reevaluation (Funk et aI., 1994a). Further discordance between NTE inhibition and OPIDN can be noted in other situations. For example, clinical and pathological manifestations of OPIDN that follow administration of a single OP compound can increase in a dose-related manner when NTE is already maximally inhibited in both hens and rats (Ehrich et aI., 1993a, 1995). Although the relationship to OPIDN has not yet been proven to be direct, there are events that do occur between NTE inhibition and the development of neuropathic lesions that define the ensuing syndrome as OPIDN. These include depression of nerve muscle conduction, inhibition of retrograde transport, and perturbations of calcium-mediated intracellular biochemical processes. A number of studies, although not all (Chemnitius et aI., 1988; Shell et aI., 1988), have indicated that administration of neuropathy-inducing OP compounds altered nerve and/or muscle electrophysiological responses in humans (Roberts, 1977; Vasilescu et aI., 1984), hens (Anderson et aI., 1988; Durham
49.8 Factors Influencing the Development of Organophosphorus-Induced Delayed Neuropathy
and Ecobichon, 1984; EI-Fawal et aI., 1988, 1989, 1990b, 1990c; Lidsky et aI., 1990; Robertson et aI., 1987, 1988), dogs (Schaeppi et aI., 1984), and cats (Abou-Donia et aI., 1986; Baker et aI., 1980; Drakontides and Baker, 1983; Lapadula et aI., 1982). For example, an increase in threshold excitability was noted in peripheral nerves (sciatic, tibial), biventer cervicis) of the hen, which is the accepted animal model for OPIDN early (1-4 days) after administration ofthe OP neurotoxic ants phenyl saligenin phosphate (PSP) and di-n-butyryl dichlorovinyl phosphate (DBDCVP) (El-Fawal et aI., 1988, 1989, 1990b, 1990c; Robertson et al., 1987, 1988). Morphological damage was seen in the biventer cervicis nerve as early as 4 days after administration of PSP, even though clinical manifestations of OPIDN were not evident before day 10 (El-Fawal et aI., 1990c). Alteration of axonal transport following administration of neuropathy-inducing OP compounds has also been investigated. NTE itself was reported to be carried by anterograde axonal transport (Carrington and Abou-Donia, 1985a). Studies in the cat sciatic nerve indicated that anterograde axonal transport was accelerated 7 days after administration of diisopropyl phosphorofiuoridate (DFP), an effect the investigators suggested was secondary to the pathologic effect, as the effect was relatively small (Carrington et al., 1989). More recent studies from this laboratory, however, suggest that sciatic nerve transport of neurofilament protein in the hen is first increased (at 3 days) and then decreased (>7 days) after exposure to DFP (Gupta et al., 1997). Another laboratory examined retrograde transport in hens given di-n-dibutyl-2,2-dichlorovinyl phosphate (DBDCVP). Retrograde transport decreased in the ventral spinal cord 3 days after treatment. Maximal effects were noted 7 days after dosing, with effects both in the ventral spinal cord and in the dorsal root ganglia (Moretto et aI., 1987). Clinical deficits were not reported before day 10 after dosing. Effects of neuropathy-inducing OP compounds have also been examined on lipid constituents of axonal membranes, including cholesterol, gangliosides, and other lipids (Bush et aI., 1995; Morazina and Rosenberg, 1970; Williams et aI., 1966). Although membrane changes are evident in OPIDN, only the relatively recent work on ganglioside profiles suggests that these particular membrane lipids may be specifically involved as OPIDN develops. Much of the work on perturbations of calcium-mediated intracellular biochemical processes has been done in the laboratory of M. B. Abou-Donia, who has summarized the work in several reviews (Abou-Donia, 1993, 1995; Abou-Donia and Lapadula, 1990). The investigators in this laboratory first noted that radiolabeled neuropathy-inducing OP compounds could bind to proteins other than NTE (Carrington and Abou-Donia, 1985b). Spinal cords collected from hens exposed to 750 mglkg tri-o-cresyl phosphate (TOCP) only 1 day previously had some increase in capability to incorporate 32p, an effect that was enhanced in the presence of calcium and that increased as ataxia developed. Increased protein phosphorylation was not noted 3 weeks after treatment with a nonneuropathy-inducing OP compound (Patton et aI., 1983, 1985, 1986).
1001
The investigators proceeded to attempt to identify the protein(s) phosphorylated following exposure to a neuropathyinducing OP, TOCP. Gel electrophoresis indicated that exposure of hens to TOCP 1,6, or 21 days previously enhanced in vitro phosphorylation of tubulin, microtubule-associated protein-2 (MAP-2), and neurofilament proteins of 70, 160, and 210 kDa. These are cytoskeletal proteins important in the maintenance of axonal integrity. Because only Ca2+ -calmodulin (CaM) kinase 11 activity catalyzes the phosphorylation of these cytoskeletal proteins, effects of neuropathy-inducing OP compounds on this enzyme were examined (Lapadula et aI., 1991, 1992; Suwita et aI., 1986a, 1986b). Further experimentation indicated that there was an increase in calmodulin binding following administration of diisopropyl phosphorofiouridate (DFP), suggesting that DFP caused conformational changes that could enhance in vitro phosphorylation of cytoskeletal proteins in tissues removed from OP-treated hens. For example, phosphorylation of tubulin and all three neurofilament subunits was enhanced in vitro by treatment with CaM kinase 11 purified from hens that had OPIDN from DFP treatment 18-21 days previously (AbouDonia et aI., 1993; Gupta and Abou-Donia, 1993, 1994, 1995b; Gupta et aI., 1992). The investigators suggested that the hyperphosphorylation of cytoskeletal proteins decreases their ability to be transported down the axon, causing accumulation (AbouDonia, 1993; Gupta et al., 1997). The preceding studies measured phosphorylation of cytoskeletal proteins in vitro in tissues from animals exposed to neuropathy-inducing OP compounds. To verify if excess phosphorylation of neurofilaments actually occurred in nervous tissue, immunohistochemical techniques were used to determine the status of phosphorylated neurofilaments in affected myelinated nerve fibers of hens exposed to neuropathy-inducing OP compounds (Jensen et aI., 1992). This study demonstrated an excess accumulation of phosphorylated neurofilaments in swollen axons at 21 (distal sciatic nerve) or 7 (spinal cord dorsal columns) days after administration of 750 mglkg TOCP. This accumulation, presumably antecedent to fiber degeneration, was thought to be related to toxicant-induced conformational change of CaM kinase 11, leading to excessive phosphorylation of neurofilaments (Jensen et aI., 1992). Immunohistochemical studies in another laboratory did not demonstrate prominent excessive phosphorylated neurofilament aggregates prior to fiber degeneration in susceptible axonal populations of hens given a neuropathic dosage of PSP (Jortner et aI., 1999; Perkins et aI., 1995).
49.8 FACTORS INFLUENCING THE DEVELOPMENT OF ORGANOPHOSPHORUS-INDUCED DELAYED NEUROPATHY Although a number of events that occur between OP exposure and development of OPIDN have been identified, questions remain about susceptibilities. It is evident that certain species
1002
CHAPTER 49
Organophosphorus-Induced Delayed Neuropathy
(e.g., the accepted animal model, the hen, as well as humans, cats, ferrets, sheep, dogs, turkeys) are more susceptible than others (e.g., rats, mice) to clinical manifestations of the toxicity (John son, 1982). Certain strains of hens, the animal model of OPIDN, also appear to be differentially affected (Bursian et aI., 1989; Dunnington et aI., 1989; Ehrich et aI., 1986a), Furthermore, young animals appear to be relatively resistant to clinical manifestations of toxicity (Davis and Richardson, 1980; Funk et aI., 1994a; Johnson, 1982; Moretto et aI., 1991; Peraica et aI., 1993). NTE inhibition, however, can be significant in most susceptible and non susceptible species and age groups. In addition, although different and less extensive than that seen in the hen, neuropathological manifestations can be noted in populations once thought not susceptible to OPIDN (e.g., rats, young chicks) (Ehrich et aI., 1993a, 1993b, 1995; Funk et aI., 1994a; Padilla and Veronesi, 1985, 1988). Comparative studies on the effects of neuropathy-inducing OPs in hens and rats have been done, including studies of the progression and regression of lesions (Carboni et aI., 1992; Dyer et aI., 1991, 1992; Ehrichetal., 1993a, 1995; EI-Fawal et aI., 1990c; Jortner et aI., 1989). These studies were done with hens 18 months old and rats more than 60 days old. Indications were that lesions could repair in both species over time, with repair occurring considerably earlier in the rat than in the hen. The reason for that is that, in hens, repair is manifest only in peripheral nerve, where myelinated nerve fiber regeneration is seen over a period of weeks following fully developed OPIDN (Jortner et aI., 1989). The repair of rat lesions is thought to be related to return of swollen, vacuolated fasciculus gracilis axons to the normal state, a more rapid process (Carboni et aI., 1992). It has been suggested that such capability for repair could at least partially explain species and age differences to induction of OPIDN, including the low susceptibility of chicks and increased susceptibility of older rats (> 6 months) as measured by clinical evidence of neurological damage following treatment with neuropathy-inducing OP compounds (Funk et aI., 1994a; Lotti, 1992; Moretto et aI., 1992b; Peraica et aI., 1993). Although repair may play a role in age and species susceptibilities to clinical and pathological manifestations of OPIDN, this process may be less significant in the treatment of OPIDN. Amelioration of OPIDN has been examined in animal models of this disorder (hen, cat) using agents suggested to provide therapeutic advantage in people or experimental animals with natural or experimentally induced neurological disorders (Capildeo, 1989; Drug Facts and Comparisons, 2001; Schlimmer and Parker, 1996; USP DI, 2001). Glucocorticoids are used in the treatment of patients with acute traumatic injuries of the nervous system (Capildeo, 1989; Drug Facts and Comparisons, 2001; Schlimmer and Parker, 1996). The first studies examining the potential of glucocorticoids to ameliorate OPIDN were done in cats (Baker and Stanec, 1985; Baker et aI., 1982; Drakontides et aI., 1982). Glucocorticoids were administered shortly after diisopropyl phosphorofluoridate (DFP) and treatments continued over the next 19-20 days. The depression of repetitive neural discharges and muscle contractile response usually seen in cats with OPIDN
4.0 TOTP and 300 ppm eOrtlcosterone TOTP IInd 200 ppm corticosterone
3.0
TOTP alone
TOTP and SO ppm corticosterone
TOTP and 30 ppm corticosterone
1.0
" Days After TOTP
Figure 49.15 Effect of 30-300 ppm corticosterone on clinical signs of delayed neuropathy induced by po administration ofTaTP (360 mg/kg) to chickens. Results are presented as mean±SD, n = 8-12. A score of 0 = no clinical signs, 1 = mild ataxia, 2 = moderate ataxia, 3 = severe ataxia, 4 = paralysis. Chickens given corn oil or corticosterone without TOTP had scores of O. Reproduced with permission from Ehrich, M., Jortner, B. S., and Gross, W. B. (1986). Dose-related beneficial and adverse effects of dietary corticosterone on organophosphorus-induced delayed neuropathy in chickens. Taxicol. Appl. Pharmacol. 83,250-260. Copyright © 1986 by Academic Press, San Diego.
did not appear and morphological damage to motor nerve terminals was much attenuated. Glucocorticoids could also ameliorate OPIDN in hens, an effect that was dependent on dose of both corticoid and OP compound. Doses of OP compound that were not overwhelming but still sufficient to induce neuropathy and relatively low concentrations of glucocorticoids were protective, as indicated by clinical, electrophysiological, and morphological endpoints (Ehrich and Gross, 1982; Ehrich et aI., 1986b, 1988; Lidsky et aI., 1990). When doses were higher, glucocorticoids (and extreme stress) could exacerbate OPIDN (Ehrich and Gross, 1983, 1986; Ehrich et aI., 1985, 1986a, 1986b, 1988) (Fig. 49.15). The myopathy caused by large doses of glucocorticoids could have been exaggerating the neuropathic effects of the OP compounds in this situation (Schlimmer and Parker, 1996). The mechanism for glucocorticoid-induced alteration of OPIDN was not related to its effect on esterase activities, as NTE was equivalently inhibited whether or not the hens received corticoids. Measurements of esterase and microsomal enzyme activities also suggested that the glucocorticoids did not affect the metabolism of the neuropathy-inducing OP compound. However, even though the mechanism(s) for the interaction of neuropathy-inducing OP compounds and glucocorticoids remains undefined, the possibility remains that stress and/or corticoids could be a contributing factor in OPIDN. Calcium channel blockers are also used to treat neurological disorders, especially those related to ischemia (Brailowsky, 1988; Drug Facts and Comparisons, 2001; USP DI, 2001). The rationale for studies on amelioration of OPIDN is based on the general role of calcium in neuronal degradation (Sch1aepfer, 1971,1987; Schlaepfer and Hasler, 1979; Schlaepfer and Zim-
49.8 Factors Influencing the Development of Organophosphorus-Induced Delayed Neuropathy merman, 1984). Axonal degeneration, which is a feature of OPIDN (see Section 49.5), has been suggested to result from an increase in calcium-dependent proteinase (CANP or calpain) activity (Sch1aepfer, 1987). Ca1pain activity increased in brain, sciatic nerve, and muscle of hens treated with TOTP or PSP, with activity significantly increased in sciatic nerve as early as 2 days after treatment with PSP (El-Fawa1 et aI., 1990a). Total nerve calcium was also increased, with this effect noted 4 days after PSP treatment. Increases in calpain activity were blocked by administration of 4 daily doses of the calcium channel blocker nifedipine, when initiated 1 day before PSP treatment. Calcium channel blockers were demonstrated to ameliorate PSP- and TOTP-induced OPIDN, as indicated by clinical, electrophysiological, and morphological endpoints (El-Fawal and Ehrich, 1993; El-Fawal et aI., 1989, 1990a, 1990b). Clinical signs developed later in hens treated with the calcium channel blockers verapamil and nifedipine. In addition, excitability thresholds of nerve-muscle preparations from hens given PSP and calcium channel blockers approached levels in preparations from control animals, and the pathological effects of PSP on myelinated fibers of the biventer cervicis nerve were markedly attenuated (Figs. 52.16 and 49.17). Calcium channel blockers have also been demonstrated to decrease lethal effects of DFP, another neuropathy-inducing OP compound (Dretchen et aI., 1986). In another laboratory, however, using a different approach, a slight decrease in calpain activity was reported along with a decrease in certain cytoskeletal proteins (e.g., a neurofilarnent subunit, NF-H, vimentin, GFAP), with the suggestion made that the proteinase was being degraded, and that axonal degeneration was related to decreased cytoskeletal proteins (Gupta and Abou-Donia, 1995a). The amelioration of OPIDN by calcium channel blockers may, be related to their effects on calpain (EI-Fawal and Ehrich, 1993; EI-Fawal et aI., 1990a) or to their action against differential vascular effects induced by neuropathic and nonneuropathic OP compounds (McCain et al., 1995, 1996). Calcium channel blockers did not affect NTE. Neuropathy-inducing PSP increased peripheral vascular resistance, response to vasoactive agents, and circulating levels of norepinephrine and epinephrine. The calcium channel blocker verapamil attenuated all of these responses. The effects of PSP on the cardiovascular system did not occur in hens exposed to paraoxon, an OP compound that does not cause OPIDN, suggesting that OP effects on the cardiovascular system may contribute to development of OPIDN. As noted previously both inhibition and "aging" of NTE are expected before OPIDN will occur. OP compounds that do not age do not cause OPIDN (Johnson, 1982). In fact, nonaging inhibitors of NTE, given prior to neuropathy-inducing OPs, will prevent OPIDN. These NTE inhibitors include carbamates, phosphinates, and sulfonyl fluorides. These compounds appear to protect the NTE from the OP compound; that these compounds protect this enzyme has been a primary reason for designating NTE as the primary target for initiation of OPIDN (Johnson, 1980, 1982, 1993; Johnson and Read, 1993; Richardson, 1992).
1003
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DAYS AFTER ADMINISTRATION
Figure 49.16A Development of clinical deficits and partial recovery after administration of phenyl saligenin phosphate (PSP = e-e), nifedipine plus PSP (NP = 0-0), or verapamil plus PSP (VP = o-{]). Verapamil, 7 mg/kg/day im, was given for 4 days beginning one day before PSP, 2.5 mg/kg im. Nefedipine, 10 mg/kg/day, was given for 5 days beginning one day before PSP. Results are presented as mean±SE, n = 5-10. Differences between the group of hens given only PSP and hens given nifedipine or verapamil plus PSP are denoted by asterisks (ANOVA with Newman-Keuls test for mUltiple comparisons, p < 0.05). I = altered gait, 2 = difficulty in walking and standing; 3 = severe ataxia, 4 = leg paralysis, 5 = paralysis with both leg and wing involvement. Hens not given PSP had scores of O. Reproduced with permission from EI-Fawal, H. A., Jortner, B. S., and Ehrich, M. (1990). Modification of phenyl saligenin phosphate-induced delayed effects by calcium channel blockers: in vivo and in vitro electrophysiological assessment. NeuroToxicology 11,573-592. Copyright © 1990 by Intox Press, Little Rock, AR.
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Figure 49.16B Log-log plot for inflection region (40-500 ).is) of strengthduration curves from biventer cervices nerve muscle preparation days 15-16 after treatment of hens with PSP (2.5 mg/kg im). The dosing regimen is given in the legend to Fig. 49.16A. Control (C) e_, PSP (P) = A-A, nifedipine plus PSP (NP) = v-v, verapamil plus PSP (VP) = 0-0. Reproduced with permission from El-Fawal, H. A., Jortner, B. S., and Ehrich, M. (1990). Modification of phenyl saligenin phosphate-induced delayed effects by calcium channel blockers: in vivo and in vitro electrophysiological assessment. NeuroToxicology 11, 573-592. Copyright © 1990 by Intox Press, Little Rock, AR.
Recently, it was discovered that administration of certain NTE inhibitors after neuropathy-inducing OP compounds will initiate or exacerbate clinical manifestations of OPIDN (Lotti et aI., 1991; Pope and Padilla, 1990). This phenomenon, called
1004
CHAPTER 49
Organophosphorus-Induced Delayed Neuropathy
Figure 49.17 Cross sections of distal levels of the biventer cervicis nerve from chickens dosed with 2.5 mg/kg im of phenyl saligenin phosphate 15 days earlier. The nerve in A is from a hen that only received the toxicant and shows extensive loss of myelinated fibers. The nerve in B was from a hen that had received the toxicant, plus the calcium channel blocker verapamil at 7 mg/kg/day for 4 days, beginning one day prior to the phenyl saligenin phosphate administration. Examination of this nerve (B) shows that the verapamil dosing was protective to the myelinated nerve fibers, many of which have a normal morphological appearance (arrow). Toluidine bluesafranin stain; bar = 100 }.tm.
promotion or potentiation, has subsequently been reproduced in other laboratories, with clinical manifestations and nervous system lesions included among the endpoints (Johnson and Read, 1993; Massicotte et aI., 1999; Pope et aI., 1992; Randall et aI., 1997; Richardson, 1995). With administration of NTE inhibitors after dosing with a neuropathy-inducing OP compound, OPIDN in hens has been exaggerated beyond what would be expected if the neuropathy-inducing OP compound were given alone. The exacerbation of OPIDN appears to be due to a quantitative rather than qualitative difference, as observed in hens given several different OP compounds (e.g., DFP, DBDCVP, PSP) and several different NTE inhibitors, with phenyl methanesulfonyl fluoride (PMSF) being used most often (Massicotte et aI., 1999; Moretto et aI., 1992a; Osman et aI., 1996; Peraica et aI., 1995; Randall et al., 1997). To date, all promotors of OPIDN are NTE inhibitors, yet most are those that do not lose a side group after attachment to the enzyme (in other words, the promotor-enzyme complex does not have to "age"). There have been reports recently of promotion occurring with a dose of an NTE inhibitor below that necessary for inhibition of the enzyme (Moretto et al., 1994; Osman et aI., 1996). During promotion, OPIDN can appear at subclinical doses of a neuropathy-inducing OP compound or in test subjects normally not susceptible to this condition (e.g., chicks, rats) (Harp et aI., 1997; Lotti et aI., 1993, 1995; Moretto et aI., 1992a,
1992b; Pope et aI., 1992, 1993, 1995). A recent report indicated that the only enzyme consistently inhibited by promoters was NTE, suggesting that some fraction or isoform of NTE may be the molecular target for promotion of OPIDN (Milatovic et aI., 1997). However, because NTE can be maximally inhibited without subsequent OPIDN by the OP compound first given, with administration of a second NTE inhibitor being followed by OPIDN, others have suggested that NTE is unlikely to be the target of OPIDN for promotion (Gardiman et aI. , 1999; Johnson, 1993; Lotti, 1992, 1995, 1997; Lotti and Moretto, 1999; Lotti et aI., 1993; Moretto et aI., 1994; Osman et al., 1996; Pope et aI., 1993). Factors other than NTE inhibition may be involved in OPIDN and promotion of OPIDN, because a recent study indicated that a soluble factor released in the spinal cord after exposure to a neuropathic OP compound had dramatic effects on cell growth (Pope et aI., 1995).
49.9 TESTING FOR ORGANOPHOSPHORUS-INDUCED DELAYED NEUROPATHY Registration of OP compounds for pesticide use under FIFRA recommends that they be tested in hens 8-14 months old without designation of breed or strain. Since 1991, this testing has
49.10 Summary
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(B) Figure 49.18 Concentration response curves for inhibition of acetylcholinesterase (AChE) and neuropathy target esterase (NTE) in neuroblastoma cells of human and murine origin by organophosphorus compounds. Cells were incubated with OP compounds for 1 h before assay. Paraoxon causes acute cholinergic crisis (AChE inhibition) rather than organophosphorus-induced delayed neuropathy (OPIDN); PSP causes OPIDN. Point-to-point composite curves are provided to aid visualization (Prism; GraphPad, San Diego). Each curve represents at least three different assays that included at least three concentrations of OP compounds that provided values between 10 and 90% of values in vehicle-treated cells. Reproduced with permission from Ehrich, M., Correll, L., and Veronesi, B. (1997). Acetylcholinesterase and neuropathy target esterase inhibitions in neuroblastoma cells to distinguish organophosphorus compounds causing acute and delayed neurotoxicity. Fundam. App!. Taxiea!. 38,55-63. Copyright © 1997 by Academic Press, San Diego.
included NTE and acetylcholinesterase determinations, clinical observations, and neuropathology following single- and multiple-dosing procedures (EPA, 1991). In the initial testing procedure, brain and spinal cord samples are collected from a subset of the dosed hens within 48 h of administration of a single dose of the test OP insecticide and NTE and acetylcholinesterase activities determined. The remaining hens are observed over the next 3 weeks, with in situ perfusionfixation and removal of brain, spinal cord, and peripheral nerves for histopathological examination at that time. Multipledose testing (28 days) may also be necessary. With these tests, the relative sensitivity of NTE to inhibition compared to acetylcholinesterase inhibition identifies those OP compounds
1005
capable of causing OPIDN even before clinical signs and morphological changes appear. Suggestions have been made that NTE measurements in cultured cells could be used to predict potential for OPIDN without the need to run this test in animals (Barber et aI., 1999a, 1999b; Ehrich and Veronesi, 1995, 1999; Ehrich et aI., 1994, 1997; Funk et aI., 1994b, 1994c; Knoth-Anderson and Abou-Donia, 1993; Knoth-Anderson et aI., 1992; Nostrandt and Enrich, 1992, 1993; Sogorb et aI., 1996, 1997; Veronesi, 1992; Veronesi and Ehrich, 1993; Veronesi et aI., 1997). Investigations indicated that NTE activity could be found in both primary cultures of avian and bovine origin and continuous cell lines of human and rodent origin (e.g., SH-SY5Y, PC-12, NB41A3). A recent, thorough concentration-response study with 11 active esterase inhibitors, including 7 that cause OPIDN and 4 that do not, indicated that either a human cell line or a murine cell line was capable of identifying the neuropathy-inducing OP compounds based on the relative sensitivity of NTE to inhibition compared to acetylcholinesterase (Ehrich et aI., 1997) (Fig. 49.18). Concentrations of OP compounds needed to inhibit NTE and acetylcholinesterase were far below those cytotoxic to the cultures. A similar result was noted in another study in which new, very sensitive NTE inhibitors were examined in cell lines of rodent origin (Li and Casida, 1997). Although it appeared that cell cultures did not have sufficient oxidative capability to convert protoxicant phosphorothioates to active enzyme inhibitors (Ehrich, 1995; Ehrich and Veronesi, 1995; Ehrich et aI., 1994, 1997), recent studies have indicated that this can be overcome by preincubation of OP protoxicants with a bromine solution or a microsomal preparation (Barber et aI., 1999a, 1999b). The results of recent studies enhance the possibility that OP compounds may one day be screened for potential to induce OPIDN by using an in vitro system.
49.10 SUMMARY Organophosphorus-induced delayed neuropathy (OPIDN) is a generally progressive, irreversible disorder that causes clinical manifestations appearing days to weeks after humans and certain species of animals are exposed to OP compounds that can essentially irreversibly inhibit most of the available neuropathy target esterase (NTE, neurotoxic esterase). The severity of OPIDN, as indicated by clinical and anatomical manifestations, depends on species and age of test animals and extent of NTE inhibition. Chickens have proven the most sensitive test species. OPIDN is manifest clinically by ataxia and weakness progressing to paralysis, associated with bilateral degeneration of distal and terminal regions of long myelinated nerve fibers. The neuropathy can be prevented by pretreatment with NTE inhibitors; yet these same compounds promote OPIDN when given after a neuropathy-inducing OP compound. Although the precise mechanism of OPIDN has not been determined, there appears to be a role for calcium, as calcium blockers ameliorated the effect, and changes on CaM kinase II activity and cytoskeletal protein phosphorylation appear after
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administration of neuropathy-inducing OP compounds. Recent studies indicate that NTE purification is imminent, and that neuropathy-inducing OP compounds and their effects on NTE can be studied in cultured cells.
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Tanaka, D. J., and Bursian, S. J. (1989). Degeneration patterns in the chicken central nervous system induced by ingestion of the organophosphorus delayed neurotoxin tri-ortho-tolyl phosphate: A silver impregnation study. Brain Res. 484, 240-256. Tanaka, D. J., Bursian, S. J., Lehning, E. J., and Auterich, R. J. (1991). Delayed neurotoxic effects of bis(1-methylethyl) phosphorofluoridate (DFP) in the European ferret: A possible mammalian model for organophosphorusinduced delayed neurotoxicity. NeuroToxicology 12, 209-224. Thomas, T. C., Ishakawa, Y., McNamee, M. G., and Wilson, B. W. (1989). Correlation of neuropathy target esterase activity with specific tritiated di-isopropyl phosphorofiuoridate-labelled proteins. Biochem. 1. 257, 109116. Thomas, T. c., Szekacs, A., Hammock, B. D., Wilson, B. W., and McNamee, M. G. (1993). Affinity chromatography of neuropathy target esterase. Chem.-Biol. Interact. 87,347-360. Thomas, T. c., Szekacs, A., Rojas, S., Hammock, B. D., Wilson, B. W., and McNamee, M. G. (1990). Characterization of neuropathy target esterase using trifluoromethyl ketones. Biochem. Pharmacol. 40, 2587-2596. Tormo, N., Gimeno, 1. R., Sogorb, M. A., Diaz-Alejo, N., and Vilanova, E. (1993). Soluble and particulate organophosphorus neuropathy target esterase in brain and sciatic nerve of the hen, cat, rat, and chick. 1. Neurochem.
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47,543-548. Veronesi, B. (1984). Effect of metabolic inhibition with piperonyl butoxide on rodent sensitivity to tri-ortho-cresyl phosphate. Exp. Neurol. 85, 65 1-660. Veronesi, B. (1992). The use of cell culture for evaluating neurotoxicity. In "Neurotoxicology" (H. A Tilson and C. L. Mitchell, eds.), pp. 21-49. Raven Press, New York. Veronesi, B., and Dvergsten, C. (1987). Triphenyl phosphite neuropathy differs from organophosphorus-induced delayed neuropathy in rats. Neuropathol. Appl. Neurobiol. 13, 193-208. Veronesi, B., and Ehrich, M. (1993). Using neuroblastoma cell lines to examine organophosphate neurotoxicity. In Vitro Toxicol., 6, 57-65. Veronesi, B., Enrich, M., Blusztajn, J. K., Oortgiesen, M., and Durham, H. (1997). Cell culture models of interspecies selectivity to organophosphorous insecticides. NeuroToxicology 18,283-297. Veronesi, B., PadiIIa, S., Blackmon, K., and Pope, C. (1991). Murine susceptibility to organophosphorus-induced delayed neuropathy (OPIDN). Toxicol. Appl. Pharmacol. 107,311-324. Veronesi, B., Padilla, S., and Lyerly, D. (1986). The correlation between neurotoxic esterase inhibition and mipafox-induced damage in rats. NeuroToxicology 7, 207-216. Vilanova, E., Barril, J., and Carrera, V. (1993). Biochemical properties and possible toxicological significance of various forms of NTE. Chem.-Biol. Interact. 87, 369-381. Vilanova, E., Barril, J., Carrera, v., and PeIIin, M. C. (1990). Soluble and particulate forms of the organophosphorus neuropathy target esterase in hen sciatic nerve. 1. Neurochem. 55, 1258-1265. WiIliams, C. H., Johnson, H. J., and CasterIine, J. L. (1966). Cholesterol content of spinal cord and sciatic nerve of hens after organophosphate and carbamate administration. 1. Neurochem. 13,471-474. WiIliams, D. G., and Johnson, M. K. (1981). Gel-electrophoretic identification of hen brain neurotoxic esterase, labelled with tritiated di-isopropyl phosphorofluoridate. Biochem. 1. 199, 323-333. Wu, S. Y., and Casida, J. E. (1994). Neuropathy target esterase inhibitors: Enantiomeric separation and stereospecificity of 2-substituted-4H-l,3,2benzodioxaphosphorin 2 oxides. Chem. Res. Toxicol. 7, 77-8\.
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CHAPTER 49
Organophosphorus-Induced Delayed Neuropathy
Wu, S. Y., and Casida, J. E. (1995). Ethyl octylphosphonofluoridate and analogs: Optimized inhibitors of neuropathy target esterase. Chem. Res. Toxicol. 8, 1070-1075. Yoshida, M., Tomizawa, M., Wu, S. Y., Quinstad, G. B., and Casida, J. E. (1995). Neuropathy target esterase of hen brain: Active site reactions with 2-[octyl-3Hjoctyl-4H-,3,2-benzodioxaphosphorin 2 oxide and 2-octyl-4H-, 3,2-[aryl-3Hjbenzodioxaphosphorin 2 oxide. J. Neurochem. 64, 16801687.
Yoshida, M., Wu, S. Y., and Casida, J. E. (1994). Reactivity and stereospecificity of neuropathy target esterase and a-chymotrypsin with 2-substituted4H-l,3,2-benzodioxaphosphorin 2 oxides. ToxieoZ. Lett. 74, 164-176. Zech, R., and Chemnitius, J. M. (1987). Neurotoxicant sensitive esterase: Enzymology and pathophysiology of organophosphorus ester-induced delayed neuropathy. Prog. Neurobiol. 29, 191-218.
CHAPTER
50 Understanding the Toxic Actions of Organophosphates Kai Savolainen Finnish Institute of Occupational Health
50.1 INTRODUCTION Organophosphates (OP) are widely used as insecticides and thus exposure to these compounds still represents a genuine health risk. The overall mechanisms of action of organophosphate-induced neurotoxic effects are well known, but the underlying molecular mechanisms of toxic actions are surprisingly poorly known. However, the introduction of a number of OP pesticides and highly toxic OP nerve agents has emphasized the importance of understanding the mechanisms of toxicity of these OP compounds in detail. It is fundamental to appreciate that their toxicity stems largely from excess acetylcholine (ACh) due to the inhibition of acetylcholinesterase (AChE) and subsequent accumulation of ACh in the target tissues, especially those cells in the vicinity of cholinergic receptors that are responsible for mediating the effects of ACh. The dramatic effects seen in OP intoxication include brain activation, epileptiformic convulsions, muscular tremors, which lead ultimately to flaccid paralysis, increased sweating and salivation, profound bronchial secretion, bronchoconstriction, increased activity of the intestine and diarrhea, miosis, hypertension, lowered body temperature, and hyperglycemia. When the effects of OP compounds are compared, marked differences are evident between them. This is most likely due to the marked differences in their ability to bind with their prime target, AChE, and the differences in the rapidity of ACh accumulation in and close to the targets of ACh. The consequences of excess ACh are primarily mediated via cholinergic muscarinic and nicotinic receptor activation. Muscarinic receptors are found in the central nervous system (CNS), blood vessel walls, and endocrine and exocrine glands. Nicotinic receptors are located in autonomic nervous ganglia, in the CNS, in the adrenals, and in the neuromuscular junction, the area specialized for transmission of neuronal impulses to striated muscles. Muscarinic receptors are G-protein-coupled, slow reacting transmembrane proteins. After activation, their effects are mediated into the cells via formation of calcium-mobilizing phosphoinositide-derived second Handbook of Pesticide Toxicology Volume 2. Agents
messengers or inhibition of adenylate cyclase, leading to increased formation of cyclic adenosine monophosphate (cAMP). Nicotinic receptors are ion channels. Their activation leads to increased influx of sodium into the cell. There are several subtypes of both receptors, and the mode of action of different muscarinic receptors and different nicotinic receptors may markedly differ from each other. In the CNS, there are more muscarinic than nicotine receptors, and muscarinic receptor activation in the brain, as in peripheral tissue, has profound effects on neuronal signalling, and can alter the numbers of many different receptors, as well as modify gene expression and the expression of proteins encoded by these genes. Cholinergic muscarinic activation also dramatically facilitates brain metabolism and induces major electrophysiological effects, often associated with overt convulsions. The CNS effects of OPs can be modified by drugs, typically cholinergic antagonists such as atropine, but also with y-aminobutyric acidergic (GABAergic) agonists such as benzodiazepines and antagonists of glutamatergic receptors. In fact, anticholinergics, like atropine, and diazepam, which belong to the benzodiazepine group of drugs, are the most effective antidotes against OP poisoning. In addition, interaction of cholinergic stimulation with lithium markedly amplifies cholinergic-induced neuronal signalling and convulsions, most likely due to an interaction at the G-protein level. Nicotinic receptors have their most dramatic effects in autonomic ganglia and the neuromuscular endplates. These alterations are characterized by a decrease in membrane potential, membrane resistance, and a decrease in afterhyperpolarization. Many of these effects can, surprisingly, be inhibited by atropine, most likely due to an interaction of muscarinic and nicotinic receptors in autonomic ganglia. The most effective blockers of nicotinic receptors are d-tubocurarine and its more modem analogs. At the neuromuscular endplate, OPs induce (1) repetitive activity in response to single nerve stimulus and (2) decremental responses to repetitive nerve stimulation. OPs typically also induce accelerated spontaneous release of ACh, leading to increases in the miniature endplate potentials
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CHAPTER 50 Toxic Actions of Organophosphates
(MEPP) frequency, but at high OP concentrations, the end result is depolarization of the neuromuscular endplate and endplate regions. The effects of OPs on neuromuscular endplates can be prevented with AChE reactivators such as pyradine-2-aldoxime methiodide (2-PAM) that can restore, in part, neuromuscular transmission. The cardiovascular and respiratory systems are particularly sensitive to the effects of OPs because both are under strict cholinergic control. A more detailed understanding of the effects of these toxic agents seems to be warranted because of their dramatic effects on the CNS. In particular, muscarinic receptor-mediated effects have been overlooked in the past. Recent observations also suggest that cholinergic stimulation of cholinergic muscarinic receptors might, in fact, be a trigger that activates many other neurotransmitter systems. For example, it is now becoming clear that cholinergic brain stimulation is the trigger that sets off the propagation of convulsive waves.
Table 50.1 Mllior Actions of Organophosphate AntichoIinesterases at Various Sites in the Body Receptor
Target organ
Symptoms and signs
Central
Central nervous system
Giddiness, anxiety, restlessness, headache, tremor, confusion, failure to concentrate, convulsions, respiratory depression
Muscarinic
Glands Nasal mucosa
Bronchorrhea
Sweat
Sweating
Lachrymal
Lachrymation
Salivary
Salivation
Smooth muscle Iris
50.2 HISTORY AND BACKGROUND Organophosphorous compounds were first synthetized in 1854, but their remarkable toxicity was not recognized until the 1930s (see Minton and Murray, 1988). The first synthesized OP pesticide, tetraethyl pyrophosphate (TEPP), was developed in Germany before World War 11 to replace the highly toxic and lipid-soluble botanical insecticide, nicotine. At the same time, the highly toxic nerve gas agents, tabun and sarin, also were developed, but they were not used during the war (Minton and Murray,1988). Anticholinesterase OP compounds have been widely used as insecticides because it was quickly appreciated that insects are highly susceptible to these agents, even though their toxicity to mammals also proved to be high (Minton and Murray, 1988). OPs represent a large group, many members of which cause toxicity by inhibiting the key enzyme, acetylcholinesterase. This enzyme promotes the hydrolysis of acetylcholine, the principal physiological cholinergic agonist of nicotinic and muscarinic receptors in the body. Thus, the toxicity of OP compounds in many respects can be considered to be ACh toxicity. Due to the ubiquitous distribution of both nicotinic and muscarinic cholinergic receptors throughout the body, exposure to OP compounds has widespread toxic consequences in several target organs. The actions of OPs include neuronal excitation in the brain and subsequent epileptic convulsions, muscular tremors, increased sweating and salivation, increased bronchial secretion, bronchoconstriction, increased activity of the intestine and diarrhea, miosis of the eyes, blurred vision, tachycardia and hypertension, hyperglycemia, and lowered body temperature. For example, when adult baboons were exposed to soman, a highly toxic anticholinesterase OP nerve agent (for soman, see Churchill et aI., 1985), via an intravenous infusion, the animals had a very rapid onset of OP-type cholinergic signs of intoxication, including overt muscular fasciculations resembling epileptic convulsions, stridorous breathing, copious secretions, and atrioventricular arrhythmias (Anzueto et al., 1986).
Rhinorrhea
Bronchial mucosa
Nicotinic
Miosis
Ciliary muscle
Failure of accommodation
Gut
Abdominal cramps, diarrhea
Bladder
Frequency, involuntary micturition
Heart
Bradycardia
Autonomic ganglia
Sympathetic effects: pallor,
Skeletal muscle
Weakness, fasciculation
tachycardia, hypertension
Source: Modified from Fuortes et al. (1993), and Marrs (1996).
OP-induced effects are mediated through the activation of nicotinic or muscarinic receptors. The distribution of these receptors varies in different parts of the body. Table 50.1 summarizes the most important toxic actions of OP compounds and classifies them according to the receptor type that is behind each toxic action. A number of recent reviews have summarized the key toxic actions of different OP compounds (Choi et al., 1995; De Bleecker et al., 1992; Ecobichon, 1996; Fuortes etal., 1993; Gunderson et al., 1992; Gutmann and Besser, 1990; Marrs, 1996; McDonough and Shih, 1997; Minton and Murray, 1988). Intoxication with OPs can take place a number of ways. Previously, when less was known about their toxicity, careless use of OPs often lead to accidental acute poisonings, and even some deaths of workers using these agents. It was estimated in 1990 that the number of annual pesticide poisonings was about 3 million cases in the world and that the incidence was thus about 57 cases per potentially exposed 100,000 individuals. There were about 20,000 cases of pesticide-induced fatalities in the world, most due to excessive exposure to OP compounds. In industrialized countries, occupational exposure to OP compounds is, for the most part, well controlled and the number of poisonings is relatively small. However, in developing countries, in which the use of OP compounds is particularly widespread because of the hot climatic conditions, the number of deaths may be high. For example, pesticide poisonings are relatively common in countries such as Sri Lanka, Venezuela, Indonesia, South Africa, and Brazil (see Choi et aI., 1995).
50.2 History and Background
Whereas OP compounds are usually highly lipid-soluble, they readily penetrate the skin, and exposure in the occupational environment takes place mainly through the skin. This is illustrated in Fig. 50.1 by the excellent correlation between alterations of the levels of mevinphos, a highly toxic OP insecticide, on the foliage in greenhouses, and the dermal contamination rate, as well as the decline in the acetylcholinesterase activity in red blood cells in the exposed workers as a function of time (Kangas et al., 1993). In some cases, exposure via the lungs also may play a role, but it is usually of minor importance (Savolainen and Kangas, 1995). In some cases, inhalation can be relatively important as the route of the compound into the body. These situations are, however, unlikely to be associated with an exposure to OP that would have remarkable health consequences. A rare example could be the use of OP nerve agents when OP concentrations in the air are likely to be high. However, in occupational environments and in situations in which OP-induced effects are likely to take place, dermal exposure predominates (Kangas et al., 1993; Savolainen and Kangas, 1995; Storm et al., 2000). Exposure to OPs via food, in household use, and thorough oral routes is nonsignificant when suicidal cases are excluded. Lethalities associated with OPs in the past contained a large number of suicides; for example, in the 1950s, parathioncontaining formulations became popular for this purpose in several countries (Choi et al., 1995; Hayes et aI., 1978; 10vanovic et al., 1990), including Finland, until their availability was strictly restricted (AI ha, 1967; Minton and Murray, 1988; Marrs, 1996). However, the most toxic OP compounds are the anticholinesterase OP nerve agents, which were created by defence industries of many countries. These include sarin (isopropyl methylphosphonoftuoridate), soman (pinacolyl methylphosphonofluoridate), tabun (ethyl N -dimethylphosphoramidocyanidate), and YX (O-ethyl-S-[2(diisopropylamine)-ethyl)methyl-phosphonothionate) (see Abdallah et al., 1992). Even though these compounds have been used rarely (e.g., the Iraqis against the Iranians and the Kurds, as well as in the well known attack by a terrorist group in a Tokyo underground station), their very existence carries a potential continuous threat (Gunderson et aI., ] 992; Ecobichon, 1996). The extreme toxicity of many OP compounds highlights the need for a more complete understanding of their mechanisms of toxic actions. Even though the overall toxicity and the general mechanisms of toxic actions of OP compounds have been rather well clarified over the years and seem to be quite similar within the group, a more thorough understanding of the cascades of cellular and subcellular toxic events of these compounds in the nervous systems is needed for effective prevention and treatment of OP-induced poisonings. For this purpose, studies on mechanisms of OP nerve agents are important, because state-of-the-art studies on mechanisms of occupationally used OPs are, for the most part, lacking.
1015
Mevlnphos (og/cm1) 80r---------------------------~
20
oL---------------------------~ 10 20 30 40
o
Time (br) aIler appllcatlon
(a) Dermal exposure rate ijlg/hr) 20 ,------nr-------------------~
16
12
8
•
ID
20
.to
30
40
50
Time (br) after appUcatlon
(b) Mevlnpbos (ngfanl) 1~,-------------------------~~ 120
90
8
12
16
20
Dermal exPOSRre rate ()lIIbr)
(c)
Figure 50.1 Dennal absorption is important in the absorption of organophosphates. Absorption of mevinphos into the body after exposure in greenhouses is used as an example in an occupational setting. (a) The decrease in the amount of mevinphos in foliar samples from flowers grown in greenhouses as a function of time. The equation of the curve is y = -0.026 + 1.789; r = 0.96. (b) The dennal exposure rate of workers exposed to mevinphos after the application of the compound to fl owers grown in greenhouses. The equation of the curve is y = 3352.7- 2.4; r = -0.67. (c) The correlation (y = 7.2 + 3.5; r 0.97) between the decrease of the amount of mevinphos on the foliage and the dennal exposure rate to mevinphos via the hands. These data provide evidence that in this situation the skin is the most important exposure route to mevinphos. Reprinted with pennission from J. Kangas et al., Am. J. lnd. Hyg. 54(4), 150--157 (1993).
=
1016
CHAPTER 50
Toxic Actions of Organophosphates
50.3 CHEMISTRY OF
ORGANOPHOSPHORUS COMPOUNDS Anticholinesterase OPs are derivatives of phosphonic or phosphoric acid or their sulfur-containing analogs, notably phosphorothioic, phosphonothionic, phosphorodithioic, or phosphonodithioic acids (see Fig. 50.2). Phosphonic acid or its derivatives does not generally inhibit AChE activity. OPs with AChE activity usually have two alkyl groups and a third group, the leaving group, that is often an aryl group or a heterocyclic group. However, in most of the OP warfare nerve agents, the leaving group contains fluorine (see Fig. 50.3; Holmstedt, 1963; Marrs, 1996; Minton and Murray, 1988; World Health Organization, 1986). The leaving group is more susceptible to hydrolysis than the alkyl groups. Typically, OPs with the P=S
Type ofOP
configuration have little or no inhibitory action on AChE unless they have been activated through enzymatic or nonenzymatic oxidative desulfuration to the corresponding oxon that contains the P=O configuration. Such compounds are termed indirect inhibitors of AChE, and include many important insecticides such as malathion and parathion (Aldridge, 1996; Ecobichon, 1996; Hirvonen et al., 1993; Savolainen et al., 1991). Holmstedt (1963) classified the OPs into four categories based on the structure of the leaving group. In category I, the leaving group contains a quaternary nitrogen. An example of thc compounds in this group is the drug ecothiopate. Category 11 includes the warfare nerve agents soman and sarin, as well as diisopropyl phosphofluoridate (DFP) (see Churchill et aI., 1987; Savolainen et al., 1988a, b), where the leaving group is fluorine. In category Ill, the leaving group is cyanide, cyanate, thio-
Example
Structure
Phosphates
Monocrotophos Chlorfenvinphos Chlorpyrifos-methyl Dichlorvos Tri-o-cresyl phosphate
Phosphonates
Trichlorfon
Phosphinates
Glufosinate"
Phosphorothioates
Pirimiphos-methyl Bromophos Diazinon Triazophos
(8=)
EPN Leptophos
Phosphonothioates (8=1
Demeton-S-methyl Ecothiopate
Phosphorothioates (S·substituted)
Phosphonothioates (S-substituted)
R
vx
(continuedl Figure 50.2 Organic derivatives of phosphoric acid.
50.3 Chemistry of Organophosphorus Compounds
cyanate, or a halogen other than fluorine. For example, the nerve agent tabun belongs to category III (see Holmstedt, 1963). The fourth group, category IV, is the most heterogeneous in terms of the structure of the leaving group, and contains a large number of pesticides. Derivatives of pyrophosphoric acid include compounds such as TEPP, sulfotep, monothiotep, schradan, and tetraisopropyl pyrophosphoramide (iso-OMPA) that is extensively used in laboratories as a specific inhibitor of butyrylcholinesterase (see Koelle, 1963; Savolainen et al., 1984). These compounds do not, in fact, conveniently fit into the classification created by Holmstedt (1963). What they have in common is an ability to inhibit AChE activity (Marrs, 1996). Some of the other OPs that express some AChE activity do not have a clearly defined leaving group. For example, S,S,S-tributyl phosphorotrithioite (DEF) and phosphorotrithioite (merphos), where all three substituents are S-butyl moieties, and perhaps ethephon, an accelerator of fruit ripening and a mono(chloroalkyl)
derivative of phosphonic acid, all belong to this poorly defined group of OPs (Marrs, 1996). In terms of OP chemistry and nonenzymatic degradation, it is important to keep in mind that most OPs are poorly watersoluble, have a high oil-water partition coefficient, and a low vapor pressure. Most of the OPs, with the exception of dichlorvos, are not particularly volatile, and all are degraded by hydrolysis, especially in alkalic conditions, yielding water-soluble products that are generally considered to be nontoxic. Knowledge of these general chemical properties of OPs has practical implications in decontamination of skin exposed to OP compounds because scrubbing the skin with (an alkali) soap causes rapid hydrolysis of the compound. Extensive reviews on the treatment and management of acute OP compound poisoning are available (see De Bleecker et al., 1992; Ecobichon, 1996; Marrs, 1996; McDonough and Shih, 1997; Minton and Murray, 1988; World Health Organization, 1986).
o
Rs-~-SR
Phosphorodithioates
Malathion Azinphos-ethyl and methyl Dimethoate Disulfoton Phosmet Phosalone
dR
o
Phosphorotrithioates
RS-~-SR s~ o
Phosphorarnidates
Phosphoramidothioates
RO-~-~ I
DEF menos
R Fenamiphos
,
OR
R
S
R
dR
or ' R
RO-~-~
0
R
dR
Methamidophos R Propetamphos
Rs-~-~
Phosphorofiuoridates
DFP
Phosphonofiuoridates
Soman Sarin
GF Pyrophosphates
(Continued).
0
"
I'
RO- P -O-P-O-R
"Not an anticholinesterase. Figure 50.2
o
6R
1017
6R
TEPP Sulfotep Schradan (OMPAl
1018
CHAPTER 50 Toxic Actions of Organophosphates
Examples of the Four Main Categories of Organic Phosphorus Compounds Group
X constituents
Example
substituted quaternary nitrogen
<;H 5-O, /-0
_
p.~
/'\..
C2Hs -O
I
S-CH2-CH2-W-(CH~
Ecothiopate isodide
11
F
CH 3 ,
/CH-O, qO CH3 ... CH) F
p,
Sarin (iC3H70h P(O)-F DFP III
CN, OCN, SCN, or halogen other than F
CH 3 ,
N ,to
CH/ 'p.~ 3
/'
C2Hs- 0
_
C=N
Tabun
Parathion
~ ° Q~"O -
CH
3
IV
alkyl; alkoxy or alkylthio; aryl or heterocyclic; aryloxy, arylthio, or one of their heterocyclic analogs; nitrogen; or disubstituted phosphoryl groups
Q-~ -
/
°
'\..0
n
CH)
U
CH) TriorthocresyJ phosphate
Figure 50.3 Examples of the four main categories of organic phosphorous compounds according to Holmstedt (1963). Reprinted with permission from Handbook of Pesticide Toxicology, Vo!. 2, p. 918, Academic Press, San Diego (1991).
50.4 ACETYLCHOLINE AND ACETYLCHOLINESTERASE, THE TARGET OF ANTICHOLINESTERASE ORGANOPHOSPHOROUSCOMPOUNDS ACh is one of the most important neurotransmitters in the human body. It can be considered an excitotoxic transmitter because when it is present in excess, it readily causes toxicity (Jope et aI., 1989; Lallement et al., 1994a; McDonough and Shih, 1997; Meldrum and Garthwaite, 1990; Naarala et al., 1997; Olney et al., 1986; Savolainen et al., 1991, 1998; Solberg and Belkin, 1997). The term excitotoxicity implies that a receptor agonist, usually a physiological neurotransmitter or
its analog, causes overt excitation of neuronal cells through receptor activation. This overt neuronal excitation then usually leads to elevated levels of intracellular calcium, activation of proteases and endonucleases, and ultimately cell death through apoptosis (programmed cell data) or necrosis (Meldrum and Garthwaite, 1990; Olney et aI., 1986; Savolainen et aI., 1998; Solberg and Belkin, 1997). The toxicity induced by OP compounds is, in fact, the toxicity of excess free ACh in the target tissues because hydrolysis by ACh is by far the most important route for inactivation ACh accumulates (Ecobichon, 1996; McDonough and Shih, 1997; Savolainen et aI., 1995). In man, the neurotransmitter ACh is present in cholinergic nerves in the CNS, at the terminal nerve endings on all postganglionic parasympathetic nerves (where it activates muscarinic receptors), innervating salivary, lacrimal, and sweat glands, at neuromuscular junctions (activating nicotinic receptors), and in the autonomic nervous system, that is, sympathetic and parasympathetic ganglia (activating nicotinic receptors). The loss of activity of AChE due to the accumulation of ACh results in excessive nervous system stimulation that may rapidly lead to respiratory failure and death (Ecobichon, 1996; Marrs, 1996; Minton and Murray, 1988). The signs of OP-induced toxicity include those that result from stimulation of muscarinic receptors in the parasympathetic nervous system (increased secretions, bronchoconstriction, miosis, gastrointestinal cramps, diarrhea, urination, bradycardia) and those that result from the stimulation and subsequent blockade of nicotinic receptors. Nicotinic receptor stimulation excites receptors in the ganglia of the sympathetic and parasympathetic divisions of the autonomic nervous system, as well as the junctions between nerves and muscles (tachycardia, hypertension, muscle fasciculations, tremors, weakness, and/or flaccid paralysis). Furthermore, stimulation of cholinergic receptors in the CNS results in restlessness, emotionallability, ataxia, lethargy, mental confusion, loss of memory, generalized weakness, convulsions, cyanosis, and coma (see also Ecobichon, 1996). 50.4.1 ACETYLCHOLINE SYNTHESIS IS
CATALYZED BY CHOLINE ACETYLTRANSFERASE ACh synthesis is catalyzed by the enzyme, choline acetyltransferase (CAT). Synthesis is regulated by the availability of choline and acetyl coenzyme A (CoA) (Jope, 1979). Choline for ACh synthesis is supplied by active choline uptake into the cell, membrane phospholipid breakdown, and blood-derived lysophosphatidylcholine (see Pepeu, 1983). Choline is taken up by cholinergic neurons from the extracellular space by active high and low affinity uptake systems (Antonelli et al., 1981). The supply of choline, which is directly proportional to the effectiveness of choline uptake into the cell, is rate limiting for de novo ACh synthesis. There is also evidence that some of the choline for acetylcholine synthesis can come from removal at the head group of the membrane phospholipid, phosphalidylcholine (Caulfield, 1993; Gutmann and Besser, 1990).
50.4 Acetylcholine and Acetylcholinesterase, the Target of Anticholinesterase Organophosphorous Compounds
Normally, the concentration of choline in the plasma is about 10 !-lmolJl. Choline does not cross-diffuse through the blood-brain barrier and, therefore, it requires an active transport system to help it cross biological membranes. Cholinergic neurons have a membrane-bound, energy- and Na+ -dependent, active choline transport system. The Km of the choline pump is 1-5 !-lmol/l (i.e., less than the concentration of choline in the extracellular space), and thus the availability of choline is not likely to be a limiting factor in the synthesis of ACh, even though it is possible that the rate of ACh, synthesis in the neurons is regulated by altering the transport capacity (Vrnax ) of the choline pump. 50.4.2 CHOLINE UPTAKE IS A TARGET FOR THE EFFECTS OF ORGANOPHOSPHOROUS COMPOUNDS
Choline uptake by neuronal cells may also be a target for the effects of OP compounds. In synaptosomes obtained from rats injected with 120 !-lg/kg of sarin or soman subcutaneously (s.c.) and studied ex vivo, sodium-dependent, high-affinity choline uptake was transiently but markedly decreased in the cortex and the hippocampus, and increased in the striatum. Similar effects on high-affinity choline uptake by soman or sarin could not be demonstrated in rat synaptosomes in vitro. The authors concluded that the OP effect was not due to a direct action of these compounds on the uptake process nor did it depend on AChE inhibition (Whalley and Shih, 1989). Earlier, Harris et al. (1982) showed that inhibition of ACh synthesis with hemicholinium, a competitive inhibitor of carrier-mediated uptake of choline across the nerve endings, effectively inhibited soman-induced elevations in cerebral ACh levels and also protected against soman toxicity. These authors concluded that excess ACh is the primary cause of OP intoxication and that the CNS is very sensitive to excess ACh. The possibility that these effects are compensatory and are aimed at protecting the organism against excessive accumulation of ACh in target tissues cannot be excluded. Liu and Pope (1998) exposed adult male rats to high doses of OP anticholinesterase compounds, chlorpyrifos (280 mg/kg) or parathion (6.6 mg/kg). They found that when AChE was maximally (82-96%) inhibited, both compounds markedly inhibited high-affinity choline uptake in striatal synaptosomes ex vivo, although chlorpyrifos was less effective than parathion. Both OPs prevented ACh release in the presence of physostigmine and the muscarinic agonist atropine; thus both chlorpyrifos and parathion most likely affected muscarinic presynaptic autoreceptor activity. ACh released from postsynaptic nerve endings binds to postsynaptic muscarinic and nicotinic receptors. Receptor activation is followed by inhibition of the hydrolysis of ACh by AChE due to ACh accumulation (Massoulie and Bon, 1982). With a continuous accumulation of free ACh at the nerve ending of all cholinergic nerves, there is a continuous activation of the postsynaptic neurons or activation of other types of postsynaptic cells, notably ganglion cells, muscle cells, or endocrine or
1019
exocrine gland cells (Ecobichon, 1996). There are several excellent reviews on the role and characteristics of ACh and AChE (Ecobichon, 1996; Erulkar, 1989) and reference will be made to these sources.
50.4.3 REACTION OF ANTICHOLINESTERASE ORGANOPHOSPHOROUS COMPOUNDS WITH ACETYLCHOLINESTERASE
The reaction between an OP compound and the active site in the AChE protein, a serine hydroxyl group, leads to the formation of a transient intermediate complex that partially hydrolyzes the leaving group. The resulting stable, phosphorylated, and largely unreactive inhibited enzyme can be reactivated, under normal circumstances, only at a very slow rate. In many cases, reaction with a number of anticholinesterase OP compounds will lead to irreversible inhibition of the enzyme (Fig. 50.4), and the signs and symptoms will be prolonged and persistent, requiring vigorous medical attention and active treatment with specific antidotes of cholinergic receptor stimulation and enzyme reactivation (Choi et aI., 1995; Gutmann and Besser, 1990). Also, once an irreversibly inhibited OP compoundenzyme complex is formed, synthesis of new AChE enzyme molecules is required to restore the normal rapid hydrolysis of ACh. Before that, several events, including cholinergic receptor desensitization (see Gutmann and Besser, 1990) and downregulation (Churchill et aI., 1984a), will take place (see later) to attenuate the effects of ACh-induced excitation of postsynaptic cholinergic receptors and subsequent excitotoxicity (Savolainen et aI., 1995), and other toxic events (McDonough and Shih, 1997). Even though much is known about the effects of OPs on AChE in animals and in in vitro systems, there is very limited information available on the effects of OP compounds on AChE in the human CNS. Postmortem examination of the distribution of AChE inhibition in the brains of two victims of parathion intoxication and of two control brains, matched for age and sex, indicated that paraoxon-induced AChE inhibition was regionally selective. The largest decrease (6085%) occurred in the cerebellum, in some thalamic nuclei, and the cortex. Only a moderate decrease (10-30%) was detected in the substantia nigra and basal ganglia, and no effects were seen in the white matter. The authors (Finkelstein et aI., 1988) concluded that a detailed knowledge of the brain regions affected by OP poisoning may explain some of the clinical manifestations of poisonings by these compounds. These results were in partial agreement with findings obtained in experimental animal studies (Churchill et aI., 1985, 1987).
1020
CHAPTER 50
Toxic Actions of Organophosphates
1988a, b). This is because the rate of AChE inhibition and subsequent ACh accumulation may be critical for these effects (Olney et aI., 1986). In a number of studies, DFP failed to produce convulsions, whereas both paraoxon and soman readily induced tonic-clonic convulsions in a subpopulation of rats; these differences may be explained by the rate of AChE inhibition by various OPs (see Savolainen et al., 1991 ; 1988a, b). The rate of AChE inhibition with paraoxon is at least 45 times greater than which can be achieved with DFP over a wide range (10- 6-10- 10 M) of paraoxon concentrations in vitro (Chemnitius et al. , 1983; Liu and Tsou, 1986). Soman inhibits AChE even more effectively than paraoxon (Aldridge and Reiner, 1972; Hoskins et al. , 1986; Koelle, 1963; Sterri et al., 1985). In addition, soman increases brain ACh concentrations much more rapidly than DFP (Fonnum and Guttormsen, 1969; Shih, 1982). Hence, the OP-induced inhibition rate for AChE may be important for OP-induced toxic effects such as convulsions and may explain some of the differences in the seizurogenic potentials of DFP, paraoxon, soman, and other OP compounds. It is also important to keep in mind that several OP-induced effects may, in fact, have an association, or even a causal relationship, with OP-induced alterations in brain metabolic events and neuronal signalling, in addition to convulsions and other behavioral effects.
A
B
c NOz
D
CzH,-o,l /p
CzH s-
°
11
° Figure 50.4 Reaction of acetylcholine (A), ecthiopate (B), and paraoxon (C) with acetylcholinesterase and positioning of 2-PAM (see Section 50.8.2) for reactivation of the enzyme inhibited by diethoxyphosphate (D) derived from either of the two inhibitors. Reprinted with permission from Handbook of Pesticide Toxicology, Vo!. 2, p. 922, Academic Press, San Diego (1991 ).
50.4.4 ORGANOPHOSPHOROUS COMPOUND
DIFFERENTIALLY INHIBIT ACETYLCHOLINESTERASE There are marked differences between different OP compounds in their ability to inhibit AChE. In fact, some of the dramatic differences between OP compounds that cause some of their serious toxic effects such as convulsions are related to their affinities for AChE (see Churchill et al., 1987; Savolainen et aI.,
50.5 DISTRIBUTION OF CHOLINERGIC SYSTEMS IN THE CENTRAL NERVOUS SYSTEM Because the emphasis of this review will be on the mechanisms of toxic actions of OP compounds in the CNS, central cholinergic pathways will be discussed in more detail. Central cholinergic pathways were initially identified by CAT immunochemistry in cat (Kimura et al., 1981), guinea pig, and rat (Kimura et al., 1980) brains. CAT reactive cells are found throughout the medial septal nucleus , diagonal band area, neostriatum (caudoputamen), nucleus accubens, olfactory tubercle region, and fields of the medial fore brain bundle. CAT reactive neurons are not found in other areas, such as neo- and piriform cortex (Kimura et al., 1980). These immunohistochemical techniques have identified several important cholinergic pathways in the brain, such as the septohippocampal pathway, and intrinsic cholinergic neurons in the caudate-putamen complex (Pepeu, 1983). In addition to immunohistochemical identification of CAT reactive neurons, information on the distribution of cholinergic receptors in the central nervous also provides insight into the cholinergic connections in the CNS and helps provide an understanding of the unique effects of OP- and other cholinergicmediated toxic actions in the CNS. It is noteworthy that the density of muscarinic receptors in the CNS as determined by eH]ACh binding is about 100 times greater than that of the nicotinic receptors (Schwarz and Kellar, 1982). Autoradiographic techniques show the highest densities of brain muscarinic receptors in the corpus striatum (nucleus
50.6 Receptors as Targets of Acetylcholine-Mediated Organophosphorus Compound Effects
caudatus-putamen), the cerebral cortex, and the hippocampal formation (Churchill et aI., 1984a, b; Kuhar and Yamamura, 1975). More specifically, stratum oriens, radiatum, and molecular layer in the hippocampus, and cingulate cortex and piriform cortex in the cerebral cortex contain particularly high densities of muscarinic receptors (Kuhar and Yamamura, 197 S). The density of muscarinic receptors is low in the thalamus, and very low or nonexistent in the midbrain and the cerebellum (Churchill et aI., 1984a, b; Kuhar and Yamamura, 1975). Thus, the distribution of muscarinic receptors resembles the distribution of the histochemical and immunohistochemical markers of cholinergic neurons (Jacobowitz and Palkovits, 1974; Kimura et al., 1980; Palkovits and lacobowitz, 1974; Pepeu, 1983). Particularly in the hippocampus, the muscarinic receptor distribution corresponds to that of cholinergic nerve terminals (Kuhar and Yamamura, 1976). Ml (specific pharmacological agonist pirenzepine; Caulfield, 1993) and M2 (specific pharmacological agonists are experimental drugs AF-DX 116 and methoctramine) receptors are distributed heterogeneously (for characteristics of the various muscarinic receptors, see the next chapter) (Mash and Potter, 1986). High densities of Ml muscarinic receptors exist in olfactory tubercle, caudate putamen, nucleus accumbens, hippocampus, amygdala, and cerebral cortex. M2 receptors are distributed throughout the brain with high densities in regions that contain large numbers of cholinergic nerve bodies (Hammer et aI., 1980; Mash and Potter, 1986). The density of receptors with a stereospecific binding site for nicotine (Romano and Goldstein, 1980) in the mammalian brain is only 1% of that of muscarinic receptors (Pepeu, 1983). In the brain, the highest concentrations of nicotinic receptors are found in the thalamus, cortex, superior colliculus, and striatum, whereas the lowest concentrations occur in the piriform cortex and hippocampus (Schwarz and Kellar, 1982). Thus, the distribution of nicotinic receptors in the CNS clearly differs from that of muscarinic receptors. It is also quite evident that most of the cholinergic effects of OPs in the CNS are mediated via muscarinic rather than nicotinic receptors. This is important because the most dramatic toxic actions of OPs are mediated via their effects on cholinergic receptors in the CNS and subsequent stimulation of other neurotransmitter systems in the brain, as well as via cholinergic receptor stimulation in other target organs, subsequent to the initial effects of OPs on AChE and the cholinergic systems (see Savolainen et al., 1995).
50.6 RECEPTORS AS TARGETS OF ACETYLCHOLINE-MEDIATED
ORGANOPHOSPHORUS COMPOUND EFFECTS Based on the great diversity of pharmacological and toxicological actions of ACh, it was not difficult to conclude that several types of receptors that mediate the effects of ACh must exist in different organs of the body. Muscarinic receptors predominate in the target tissues innervated by the parasympathetic nervous
1021
system, endocrine and exocrine glands, and vessel walls. Furthermore, muscarinic receptors also predominate in the CNS. The effects of ACh on muscarinic receptors can in most cases be blocked by atropine. Nicotinic receptors predominate in autonomic nervous ganglia, in the adrenals, and within the neuromuscular junction, in the muscular endplate. The effects of ACh on nicotinic receptors cannot be blocked by atropine (Caulfield, 1993). 50.6.1 MUSCARINIC RECEPTORS Muscarinic receptors (mAChRs) are the main cholinergic receptors in the CNS and are expressed at high concentrations (Fisher et aI., 1983). The nicotinic receptor differs and is much like the other ligand-gated ion channels, such as the GABA acid receptor, causing rapid signal transduction in response to ligand binding. The muscarinic receptor is a member of the family of (7- transmembrane (7- TM» surface proteins (Perelta et aI., 1988). Muscarinic receptors transduce their signals across the membrane by interacting with guanosine S'-triphosphate (GTP) binding proteins (G proteins) (see Gilman, 1987). There is a cascade that involves a number of macromolecular interactions in muscarinic receptor activation, which means that the responses of muscarinic receptors are slow compared with those mediated through the nicotinic receptors. The mAChRs are also termed calcium-mobilizing receptors, because they are usually coupled with phosphoinositide (PI) signalling and may either excite or inhibit neurons (Putney, 1987). At somatic neuromuscular junctions, mAChRs also mediate a diverse range of physiological actions, including the regulation of cardiac and smooth muscle contraction, and exocrine gland secretion (Leiber et aI., 1990). The stimulation of mAChR leads to activation of membrane integral PI-specific phospholipase C (PLC), which in turn promotes facilitated bifurcating PI metabolism and increased formation of two second messengers, inositol-1,4,S-trisphosphate (InsP3), and diacylglycerol (DAG). This latter compound activates a key cellular enzyme, protein kinase C (PKC), which is involved in cell activation, regulation of differentiation, and even apoptosis (Dypbukt et aI., 1994; Orrenius, 1997). Furthermore, the activation of mAChRs leads to the inhibition of guanylate cyclase (Felder et aI., 1989; Gilman, 1987; Hanley and Iversen, 1978), the activation of cAMP-specific phosphodiesterase (Meeker and Harden, 1982), and the synthesis of prostanoids (Busija et aI., 1988). The relationships between these responses are not well understood and, in fact, multiple mAChRs may be responsible. Also, the mAChRs seem to be involved in the regulation of Ca2+ -dependent K+ channels (Baraban et aI., 1985) and voltage-dependent Ca2+ channels (Hesheler et al., 1987). Due to their structural characteristics, muscarinic receptors are also termed seven transmembrane-domain receptors, indicating that the receptor has seven hydrophobic transmembrane segments that penetrate the cell membrane (Fig. SO.S; Caulfield, 1993). Guanine nucleotides inhibit muscarinic agonists from
1022
CHAPTER 50 Toxic Actions of Organophosphates Table 50.2 Muscarinic Acetylcholine Receptor Sublypes, Localization, and Effects on Second Messenger Systems Receptor sUbtype
M]
M2
M3
Gene
ml
m2
m3
m4
+
+
mSb
Gene localizationc Brain
+ +
Heart Sal ivary glands Figure 50.5 G-protein-coupled seven transmembrane domain receptors are composed of a single subunit with seven presumptive transmembrane domains. Muscarinic receptors belong to this category of receptors that are coupled through G proteins to an effector enzyme, phospholipase C. Its activation leads to increased formation of calcium· mobilizing second messengers, hence also the name calcium-mobilizing receptor (for functional details, see Fig. 50.7).
Intestines
+
Trachea
+
Urinary bladder
+ + + +
+ +
Cell line (NG108-1 5) Effects on second messengersd
PI cAMP
binding to cell membrane receptors, but they do not prevent the binding of muscarinic antagonists to these receptors (Hesheler et al., 1987). Muscarinic receptor signalling is discussed in more detail in Section 50.7.1, which deals mainly with PI metabolism and events associated with PI signalling. This is because PI signalling-related events are more relevant to OPinduced neuronal excitation than inhibition of adenylate cyclase (Cockcroft, 1986; Mei et al., 1989). Initially, mAChRs were classified into two groups according to their pharmacological properties: one in the brain with a high affinity for pirenzepine (Md and the other in the peripheral organs with a low affinity for pirenzepine (M2; Cross et aI., 1984). Activation of M) receptors most likely stimulates PI metabolism and formation of the two second messengers, InsP3 and DAG (Berridge, 1989), whereas M2 receptors inhibit cAMP formation. However, today three types of mAChRs can be defined based on their affinities toward selective agonists. In addition, five genes that code for the muscarinic receptors have been identified (Liu et al., 1986; Mei et al., 1989). The different types of mAChRs are coupled with different and specific second messenger systems (see Table 50.2). It is evident that the Ml and M2 receptors are encoded by m) and m2 genes, respectively (Perelta et al., 1987). Later, the protein encoded by the m3 gene was identified as the M3 mAChR (Buckley et aI., 1989; PinkasKramarski et aI., 1988). The specific pharmacological antagonist of the M3 receptor is hexahydrociladephinole (Caulfield, 1993). The pharmacology of the mAChRs encoded by the Ill4 and ms genes is not clear. When expressed in cells, the Ill4 geneencoded receptor shows a relatively high affinity for pirenzepine and thus resembles the M) receptor, which is coupled with PI hydrolysis and formation of InsP3 and DAG. These two receptors are, however, two distinct receptor subtypes because the Ill4 encoded receptor seems to couple to adenylate cyclase instead of PLC (Ashkenazi et al., 1989). The possibility that different muscarinic receptor subtypes differ in their ability to couple with different G proteins cannot be exclude. The mAChR sUbtypes also exhibit a distinct regional distribution in the brain. The main consideration with ACh-induced stimulation of mAChRs is the subsequent activation of the target cells, which may lead to overt excitation and toxicity, especially in the CNS.
+ +-
* *
* *
aFor references, see the text. bThe gene ms has been cloned from a rat brain cDNA library (Liao et aI., 1989). c+, detectable; +-, barely detectable; -. not detectable. d*, stimulation.
Direct effects of OP compounds on muscarinic receptors were studied by using rat brain membranes or cultures of human neuroblastoma N lE-lIS cells (Bakry et al., 1988). Sarin, soman, or tabun had no effect on the receptors, but VX and echothiopate inhibited, in a competitive manner, the binding of l-quinuclidinyl(phenyl-4eH])-benzilate ([3H]QNB) and of [3H]pirenzepine ([3H]PZ) to muscarinic receptors, with VX being the most potent in this respect. The authors (Bakry et aI., 1988) suggested that OP compounds may directly act on muscarinic receptors if their concentration in the circulation is at or above micromolar levels. However, the mechanism of this receptor-OP compound interaction remains to be elucidated. Ward et at. (1993) explored the interaction of eight OP compounds with muscarinic receptors with regard to their ability to inhibit AChE activity in vitro in tissue homogenates from rat hippocampus and frontal cortex. Of the compounds tested, only ecothiopate competed for eH]QNB binding and only at concentrations exceeding 100 f.lM. The OP anticholinesterases did compete, however, with a muscarinic receptor agonist, eH]CD (eH]cis-methyldioxolane) that binds with a high affinity to 10 and 3% of muscarinic receptors in the frontal cortex and hippocampus, respectively. Ecothiopate and DFP were potent inhibitors of eH]CD binding as were the active oxon forms of parathion, malathion, and disulfoton. A similar pattern of potency was observed for the inhibition of brain AChE activity, indicating that there was a strong correlation between the abilities of OP compounds to inhibit eH]CD binding and to inhibit AChE activity.
50.6 Receptors as Targets of Acetylcholine-Mediated Organophosphorus Compound Effects
1023
50.6.2 NICOTINIC RECEPTORS
The nicotinic ACh receptor is the best characterized neurotransmitter receptor, and understanding its function is essential for understanding the mechanisms of action of OP compounds. Electric organs of the Torpedo species have served as a rich source of nicotinic receptors. The electrical discharge of Torpedo depends solely on a postsynaptic excitatory potential that results from depolarization of the postsynaptic membrane due to an interaction between the receptor and a nicotinic receptor agonist. Depolarization arises directly from the opening of receptor channels. In skeletal muscle and in the fresh water electric eel, Electrophorus electricus, depolarization at the endplate activates a voltage-sensitive Na+ channel that causes the depolarization to spread across the surface of the muscle or electric organ. The density of receptors in Torpedo electrical organs is about 100 pmoVmg protein, as compared with 0.1 pmoVmg protein in skeletal muscle (see Taylor and Brown, 1989). Several snake a toxins, including a-bungarotoxin, irreversibly inactivate receptor function in intact skeletal muscle, and this property of the toxin has been utilized to identify and isolate the nicotinic ACh receptor from Torpedo (Changeux et aI., 1984). Labeled ex toxins have been utilized as markers of the nicotinic receptor during its solubilization and purification. The nicotinic ACh receptor belongs to the group of ligandgated ion channels and consists of five subunits arranged around a pseudoaxis of symmetry. In the nicotinic receptors of skeletal muscles and the electric organ of Torpedo, one of the subunits, designated a, is expressed in two copies; the other three subunits, f3 , elY, and 8, are present as single copies (see Fig. 50.6; Taylor and Brown, 1989). The receptor is thus a pentamer of molecular mass near 280 kDa. However, the neuronal nicotinic receptors contain only ex and f3 subunits in different combinations with two ex subunits and three f3 subunits. Structural studies have shown that the subunits are arranged around a central cavity, with the largest portion of the protein exposed toward the extracellular surface. The central cavity is most likely the ion channel that, in its resting state, is impermeable to ions. However, once activated, it opens to form a 6.5 A diameter pore. The open channel is selective for cations, and permeation of the channel by a particular cation seems to be controlled by the diameter of the open channel. Both of the ex subunits have a site for binding ACh (plus other nicotinic agonists). ACh must occupy both sites to permit receptor activation and subsequent channel opening (Maelicke, 1986). This then leads to a brief surge of Na+ ions into the cell. The influx of Na+ ions causes a change in membrane potential, and this induces a localized depolarization of one part of the cell membrane. This depolarization of the neuromuscular endplate is termed the endplate potential (epp). If this local depolarization is sufficiently large, it can trigger an action potential that spreads throughout the cell. In a muscle cell, the generation of an action potential leads to muscular contraction. Compounds that are competitive antagonists of ACh, such as d-tubocurarine, compete for the ACh binding site in the ex subunit of the nicotinic receptor. They are
Figure 50.6 Molecular structure of the nicotinic cholinergic receptor. The structure of the receptor is described in the text. The figure shows a longitudinal view with the y subunit removed. The remaining subunits, two copies of a, one of {3, and one of y, surround an internal channel with an outer vestibule and a narrowing located deep in the membrane bilayer region. Acetylcholine binding sites, indicated by arrows, are found at the ay and ao (not visible) interfaces. Reprinted with permission from N. Unwin, J. Mol. Bioi. 229, 11011124 (1993).
termed competitive inhibitors because they can be displaced from the binding site by increasing the concentration of ACh in the vicinity of the receptor (see Taylor and Brown, 1989). It is noteworthy that a and f3 subunits of neuronal nicotinic receptors differ from those in skeletal muscle nicotinic receptors, but are structurally and chemically identical in their amino acid content, and they are evolutionary homologs (Unwin et aI., 1988). Neuronal nicotinic receptors differ in structure and pharmacological characteristics from skeletal muscle nicotinic receptors. Ca2+ ions permeate more easily through neuronal nicotinic receptors than they pass through skeletal muscle nicotinic receptors; changes in the levels of free intracellular Ca2 + concentrations may markedly modify those responses mediated through neuronal nicotinic receptors. The problem with nicotinic receptors is, in fact, more complex than described above. To date, genes that encode 16 different subunits of vertebrate nicotinic receptors have been cloned. These subunits are identified as a 1-ex9, f31-f34, y, 8, and e. However, because neither the subunit stochiometry nor the arrangement of most of the receptors is known with certainty, this issue will not be tackled in more depth in this context (Lukas et al., 1999; McGehee and Role, 1995; Sargent, 1993). Instead, a simplified approach will be utilized. OP anticholinesterases may have direct actions on nicotinic receptors. There are data to suggest that OP anticholinesterases bind to allosteric sites of the cholinergic nicotinic receptors as identified by inhibition of eH]phencyclidine binding, but some can also bind to the receptor's recognition site because they inhibit [125 I]a-bungarotoxin binding (Bakry et al., 1988). Soman and ecothiopate at micromolar concentrations acted like
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CHAPTER 50 Toxic Actions of Organophosphates
partial agonists of the nicotinic receptors and induced receptor desensitization. On the other hand, VX acted like an open channel blocker of the activated receptor (i.e., a compound that can only gain access to the ion channel when it is in the open configuration) and also enhanced receptor desensitization. The authors (Bakry et aI., 1988) suggested that the toxicity of OP compounds may include some direct actions on the nicotinic receptor if their concentration in the circulation exceeds the micromolar level. The mechanism of this nicotinic receptor-OP compound interaction remains to be elucidated. Chi and Sun (1995) found that soman, sarin, tabun, and phencyclidine did not modify the binding of eSI]a-cobratoxin to the nicotinic receptor. Katz et al. (1997) also reported that incubation of Torpedo membrane-bound nicotinic receptors with the muscarinic agonist carbacholine stimulated the binding of [3H]thienylcyclohexylpiperidine ([3H]TCP), which binds to the receptor's noncompetitive agonist binding site in its ionic channel. This agonist stimulated binding of eH]TCP was inhibited in a dosedependent manner by OPs such as chlorpyrifos oxon, chlorpyrifos, parathion, and paraoxon. The OPs did not have any effect on equilibrium binding of [a 125 I]bungarotoxin to the receptor's ACh binding site, but preincubation of the membranes with OPs increased the site's affinity for carbachol. In the absence of an agonist, the OPs increased the binding of [3H]TCP markedly. These data suggest that, in addition to AChE inhibition, OPs bind directly to a site on the nicotinic receptor that is distinct from the ACh or TCP binding sites and that this binding induces nicotinic receptor desensitization. In summary, these results indicate that OP compounds may exert direct actions on nicotinic receptors.
50.7 CENTRAL NERVOUS SYSTEM CHOLINERGIC EFFECTS OFORGANOPHOSHOROUS COMPOUNDS AND OTHER CHOLINERGIC AGONISTS Stimulation of neuronal cholinergic, predominantly muscarinic receptors (Pepeu, 1983) in the CNS with high doses of direct or indirect cholinergic agonists such as ACh, OP compounds, pilocarpine, carbachol, and oxotremorine has profound effects on neuronal aerobic glucose, phospholipid, sphingolipid, and RNA metabolism, neuronal signalling, and osmoregulation (Churchill et al., 1984a, b; Hoskins et aI., 1986; Jope and Morrisett, 1986; Jope et aI., 1986; Pazdernik et aI., 1985; Savolainen et aI., 1988a, b; Wade et aI., 1987), as well as a having a profound impact on neuronal electrophysiological effects (see, e.g., Gutmann and Besser, 1990). These events are associated with dramatic behavioral effects such as generation of epileptic foci, and clonic and tonic-clonic convulsions. If these effects are not interrupted, they inevitably lead to the demise of experimental animals and humans (Churchill et al., 1984a, b; McDonough and Shih, 1997; Pazdernik et al., 1985, 1986;
Savolainen et aI., 1988a, b; Wade et aI., 1987). Long-lasting convulsions are often associated with serious cell losses and brain damage, in addition to dramatic downregulation of cholinergic muscarinic receptors (Churchill et al., 1985; Pazdernik et al., 1985). In the following section, some of the most important effects of cholinergic brain stimulation will be described, with a special emphasis on OP compounds. 50.7.1 EFFECTS ON CEREBRAL CHOLINERGIC SIGNALLING
It can be said that OP-induced brain stimulation is predominantly a consequence of the activation of cerebral mAChRs and, hence, activation of receptor-coupled G protein (Fain et al., 1988; Gilman, 1987) and subsequent activation of PLCmediated PI signalling (Berridge and Irvine, 1989). Thus, events associated with facilitated PI hydrolysis play an crucial role in OP-induced brain effects. The following paragraphs give a brief description of the signalling cascade. PIs play a key role in muscarinic cell signalling as precursors of second messengers that are responsible for transducing the signal from the cell surface muscarinic receptors into the cell (Berridge, 1989). PIs are different from all other membrane phospholipids because kinases are able to further phosphory late their inositol head groups. Although PIs account for about 10% of the total phospholipid composition of the cell membrane in most cells, phosphatidylinositol-4,5-bisphosphate (PIP2) is a minor membrane component that makes up between 1 and 10% of the total PI pool. Its concentration is higher in the brain than in any other tissue, which suggests that it plays an important role in the specialized functions of the nervous system. Stimulation of calcium-mobilizing receptors, to which muscarinic receptor subtypes belong, initiates a bifurcating hydrolysis pathway of PIP2, an acidic membrane-bound phospholipid. Hydrolysis of this membrane phospholipid results in the formation of two second messengers, InsP3 and DAG (Downes and Michell, 1981). DAG stimulates PKC, an enzyme vital for several important cellular functions, including receptor-mediated activation (Nishizuka, 1988), whereas InsP3 diffuses into the cytosol to release calcium from nonmitochondrial internal stores and, perhaps indirectly, to stimulate the entry of extracellular calcium into the cell (Berridge, 1987). Ultimately, the PI pathway leads to the reformation of PIP2, and the cycle is again primed. Lithium inhibits the metabolism of inositol phosphates in the final dephosphorylation step (see Fig. 50.7). Thus, lithium is likely to reduce the supply of free inositol required to maintain the formation of lipid precursors used for cell signalling. These pathways regulate several cellular processes, including metabolism, contraction, neural activity, and cell proliferation (Berridge, 1989; Berridge and Irvine, 1989). OP compounds induce their effects by inhibiting AChE, leading to accumulation of ACh and excessive activation of mAChRs. Thus, OP-induced brain effects and neurotoxicity are, in fact, the toxicity of excess ACh (Savolainen et aI., 1998). Katz and Marquis (1992) exposed human SK-N-SH
50.7 Central Nervous System Cholinergic Effects of Organophoshorous Compounds and Other Cholinergic Agonists
Calcium stores Figure 50.7 Receptor-mediated phosphoinositide turnover: An agonist (A) such as ACh binds to receptor (R), causing the activation of a G protein (Gp) that, in turn, stimulates PLC. The PLC hydrolyses PIP2, generating InsP3 and DAG. InsP3 binds to a specific receptor on the membrane of a nonmitochondial cell organelle that contains Ca2+, causing the release of calcium from the intracellular stores. The binding of InsP3 to its receptors is inhibited by heparin and increased intracellular H+ and Ca2+. InsP3 can further be phosphorylated to InsP4, which may facilitate the entry of calcium into the cell through the plasma membrane or may trigger the movement of calcium within the cell. DAG activates PKC, which phosphorylates a large number of substrates. Activation of cAMP-dependent protein kinase A (PKA) leads to phosphorylation of InsP3 receptor protein.
neuroblastoma cells to low concentrations of paraoxon or carbachol, a direct muscarinic agonist. Paraoxon inhibited the N-[3H]methylscopolamine (eHlNMS) muscarinic receptor binding. Paraoxon, also at low concentrations (0.1 nM), caused a time-dependent increase in the PI turnover, whereas high concentrations of carbachol were required for the same effect. Both pertussis toxin, a G-protein inhibitor, and neomycin, a PLC inhibitor, inhibited cholinergic-induced facilitation of PI hydrolysis. It seems that paraoxon may modulate signal transduction in neuronal cells by indirect activation of muscarinic receptors, that is, by elevating levels of ACh, as well as by acting at a site distal to the receptor (Katz and Marquis, 1992). Bodjarian et al. (1992) demonstrated that soman also facilitates PI hydrolysis in hippocampal slices from rats. The effect was mediated through muscarinic receptor subtypes MI and M3 subsequent to AChE inhibition and ACh accumulation. Even though the M2 muscarinic receptor subtype is preferentially coupled with inhibition of adenylate cyclase, leading to reduction of levels of cAMP, it was also shown to be associated with PLC-mediated hydrolysis of PIs (Mei et al., 1989). Thus, findings from in vitro studies are consistent with the assumption that OP compounds affect neuronal PI signalling and that this is mediated via cholinergic muscarinic receptor activation. Savolainen and co-workers have shown in a series of in vivo studies that exposure of experimental animals to several OP
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compounds, such as DFP, malaoxon, paraoxon, and soman, dramatically increases PI signalling, as measured by the accumulation of inositol-I-monophosphate, that is closely associated with cholinergic-induced clonic or tonic-clonic convulsions (Hirvonen et al., 1989, 1990, 1993; Savolainen et al., 1988a, b, 1991; Savolainen and Hirvonen, 1992). In these studies, rats were usually pretreated with lithium (3-5 mequiv/kg), because lithium amplifies the cholinergic-induced PI turnover and its associated convulsions (Hirvonen et al., 1990; Honchar et al., 1983; Savolainen et al., 1988a, b). When rats were given soman with or without lithium pretreatment, tonic-clonic convulsions and increased PI turnover could be prevented by pretreating the animals with atropine or benzodiazepine, indicating that the events are associated with stimulation of mAChR and are modifiable by increasing the GABAergic tone in the CNS (Savolainen et al., 1988a, b). In another study (Hirvonen et al., 1990), malaoxon-induced convulsions were associated with marked increases in PI signalling and early neuronal injury, especially in the hippocampus (Hirvonen et al., 1990). During an investigation of the differences between male and female rats, Savolainen and Hirvonen (1992) observed that malaoxon stimulates PI signalling in the brain of the rats' offspring and that females seem to be more sensitive than males toward the OP-induced cholinergic brain stimulation and alterations in PI signalling. Lithium seemed to increase the sensitivity of experimental animals toward OP-induced convulsions. Furthermore, increased PI turnover in the hippocampus may indicate a lithium-induced lowering of the seizure threshold for OP in limbic regions (Savolainen et al., 1991). Hirvonen et al. (1993) gave a single convulsive dose of pilocarpine and then overt convulsions were abruptly terminated with diazepam after 15 min. This did not prevent the occurrence of serious brain damage and facilitated PI signalling was still seen 5 days after the cessation of the convulsions and exposure to pilocarpine. At this stage, the entire dose of pilocarpine had been excreted via urine and all convulsions had been over for 5 days. The authors concluded that pilocarpine served as a trigger for seizures and convulsions, and, subsequently, other neurotransmitter systems, most likely systems coupled to PI turnover, became activated and were responsible for the residual stimulation of PI turnover as well as the associated brain damage. The candidate receptors most likely to become activated after initial cholinergic neuronal stimulation are the glutamatergic receptors (see Savolainen et al., 1998). The hypothesized pathway is as follows: AChE inhibition by OP compounds leads to ACh accumulation and subsequent excessive cholinergic muscarinic stimulation. There are convulsions that lead to release of glutamate, activation of glutamate receptors, influx of calcium into the neurons, and neuronal apoptosis or necrosis. Similar effects can also be induced by stimulating neuronal cells with direct cholinergic muscarinic agonists such as pilocarpine or carbachol (Felipo et al., 1998; Hirvonen et al., 1993; Savolainen et al., 1995, 1998; Solberg and Belkin, 1997). Bodjarian et al. (1995) provided evidence that, in addition to cholinergic receptors, histamine HI subtypes and glutamate metabotropic receptors also are involved in the facilitated PI
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CHAPTER 50 Toxic Actions of Organophosphates
signalling that is associated with soman-induced convulsions. In contrast, 5-HT2 or ar adrenoreceptors-receptors also coupled to PI signalling-were not associated with these somaninduced signalling events. Cholinergic brain stimulation by OP compounds also affects the levels of cyclic nucleotides-AMP and cyclic GMP-in different brain regions due to inhibition of the cyclase enzymes (Liu et al., 1986). In another study, Liu et al. (1988) observed that decreases in the levels of AMP or cyclic GMP in the striatum, cerebellum, and spinal cord were not related to soman-induced convulsions in rats. This is consistent with the assumption that the facilitation of PI signalling, rather than the activity of adenylate cyclase, is essential for cholinergicinduced convulsions, associated second messenger formation, and neuropathology (Hirvonen et aI., 1990, 1993). 50.7.2 EFFECTS ON RECEPTORS The density of various receptors and their subtypes in different brain regions markedly affects brain cholinergic signalling and subsequent metabolic events, and the pathological consequences. Yagle and Costa (1996) exposed Sprague Dawley (SD) rats to doses of 2 mglkg per day of disulfoton (S-(2(ethylthio)ethyl)phosphorothionate) for 14 consecutive days, and measured messenger ribonucleic acid (mRNA) levels of muscarinic receptor mr, m2, and m3 subtypes, immediately after the cessation of the exposure, as well as after a 28-day recovery period. There was a marked reduction in the levels of muscarinic receptor subtypes in several brain regions immediately after the exposure, but after the recovery period, only the m2 subtype mRNA levels remained decreased, indicating that this receptor subtype may be more sensitive than the others toward OP-induced alterations. Also, marked reductions in eH]QNB binding were seen immediately after the cessation of the exposure, indicating a marked reduction in muscarinic receptor numbers. The findings of Yagle and Costa (1996) are consistent with earlier observations by Doebler et al. (1983a), who showed that repetitive s.c. injections of soman at a 0.5 LDso dose level caused a marked and progressive RNA depletion in caudate and cortex. Somaninduced reductions of overall brain RNA levels were mediated via muscarinic receptor stimulation because they could be completely blocked by pretreatment with atropine when given together with pralidoxime (Doebler et aI., 1983b). Feeding mice with parathion (0.4-500 mglkg/day) in their diet for 14 days inhibited mouse brain AChE activity and transiently reduced the maximal binding of eH]QNB, [3H]NMS, and eH]4DAMP ([3H]-4-diphenylacetoxy- N -methylpiperidine methiodide) binding without affecting receptor affinities for these ligands (Jett et al., 1993). Inhibition of whole brain AChE varied between 10 and 80% in a dose-dependent fashion. These results suggest that dietary doses of parathion induced a transient downregulation of different muscarinic receptor sUbtypes in the mouse brain. Churchill et al. (1984a, b) found that [3H]QNB binding to muscarinic receptors in the rat fore brain decreased after con-
vulsions induced by a single dose of soman, as well as after a single dose of conic acid, an analog of an excitotoxic neurotransmitter, glutamate (Churchill et al., 1990). The reason for the decreased binding of [3H]QNB was a decrease in the number of muscarinic receptors rather than the affinity for the ligand. Blockade of the convulsions by diazepam also inhibited alterations in the eH]QNB binding. Consistent with the findings of Churchill et al. (1990), Chaudhuri et al. (1993) found that high doses of parathion or chlorpyriphos, both anticholinesterase OP compounds, markedly inhibited [3H]QNB binding in cortex and striatum. Abdallah et al. (1992) found that 100 fl-M paraoxon markedly reduced the Bmax of eH]4-DAMP binding, an M3 muscarinic receptor subtype-specific antagonist, without any significant alterations in its affinity toward the receptor in rat submaxillary gland (SMG) cells, indicative of a noncompetitive inhibition of the binding by paraoxon in these cells. The authors suggested that paraoxon may bind to two different sites in these SMG cells. One might be an allosteric site on the M3 muscarinic receptor that modulates receptor function. The other site could be in the G; protein-adenylate cyclase complex (Gilman, 1987). These findings complicate our understanding of OP toxicity, because they also mean that OPs can have direct effects on muscarinic receptors. Viana et al. (1988) demonstrated that when PC12 pheochromocytoma cells were treated with nerve growth factor (NGF), the number of [3H]NMS binding sites increased twofold, but NGF did not change the Kd for this ligand. Exposure of these cells to 50 fl-M soman decreased the number of binding sites in both cells with and without NGF treatment. Other OP compounds, including sarin, tabun, and VX, also reduced eH]NMS binding. These reductions in muscarinic binding were not reversed by atropine. Interestingly, similar changes induced by carbacholine were reversed by atropine. It thus seems that the decreases in muscarinic receptor binding in PC12 cells induced by OP compounds occur via a mechanism that does not necessarily involve agonist-induced receptor desensitization. Blanchet et al. (1986) previously reported results that concurred with those of Viana et al. (1988) that exposure of mouse NS-20 and N1E-115 neuroblastoma cells to soman markedly reduced the number of eH]NMS binding sites and inhibited carbacholine-induced cyclic GMP formation. Thus, exposure of neuronal cells to an OP was able to induce muscarinic receptor downregulation and subsequent desensitization of muscarinic receptor-mediated responses. Low levels of paraoxon also have been found to block M2 and M3 muscarinic receptors in homogenates of calf caudate nuclei as indicated by modulation of paraoxon-induced inhibition of [3H]QNB binding by specific antagonists of M2 (AF-DXl16) and M3 (4-DAMP) receptors (Katz and Marquis, 1989). To explore whether a critical period exists for OP-induced alterations in the density of mAChRs and spontaneous behavior, neonatal mice were exposed to a single dose of 1.5 mglkg ofDFP on neonatal days 3,10, or 19, and followed for changes in receptor density and behavioral alterations at the age of 4 months (Ahlbom et al., 1995). Mice exposed on days 3 or 10 showed a decrease in mAChR density and spontaneous mo-
50.7 Central Nervous System Cholinergic Effects of Organophoshorous Compounds and Other Cholinergic Agonists tor behavior, but those exposed on day 19 no longer showed any effects. The lack of effects in mice exposed on neonatal day 19 was not due to differences in AChE activity; thus, we cannot exclude the possibility that a critical period exists for OP-induced cholinergic effects during neonatal life. Jett et al. (1994) found that high doses of parathion, resulting in 84-90% inhibition of AChE, to adult male rats for 21 days did not affect the affinities of different mAChR subtypes toward ligands. However, there was a significant downregulation of the Ill4 receptor subtype gene product, and mt mRNA and m3 mRNA in the frontal cortex as well as the Ill4 mRNA in the striatum. However, no changes in mAChR subtype gene products or mRNAs were found in the hippocampus. These findings indicate that paraoxon-induced ACh accumulation causes a marked depletion in the numbers of receptors of several muscarinic receptor subtypes. Whereas the degree of AChE inhibition in all brain areas was of the same magnitude, differences in parathion concentrations do not explain the lack of effect in the hippocampus, a key brain region in the regulation of seizure and convulsive activity. One possibility, though, is that in the hippocampus, OP-induced effects were less severe because of quicker renewal of the muscarinic receptors in this brain region. Churchill et al. (1985) used muscarinic receptor autoradiography after soman administration sufficient to cause a pronounced weight loss in a subgroup of rats to reveal a consistent pattern of cell loss with extensive neuronal necrosis. Receptor autoradiography indicated that these changes were associated with a dramatic decrease in the numbers of muscarinic receptors in the piriform cortex and the thalamus. These authors concluded that quantitative receptor autoradiography provides, in addition to kinetic information and topographical distribution, radiohistochemical evidence of neuronal damage that most likely took place subsequent to early effects of soman, such as alterations in total or muscarinic receptor-specific mRNA levels. Aas et al. (1987) demonstrated, however, that exposure to a low concentration of soman in the inhalaled air did not induce any marked reduction in muscarinic receptor binding characteristics in the hippocampus and the neostriatum of rats. This finding emphasized a fundamental aspect of toxicology-the significance of dose-even when we are dealing with highly toxic agents such as soman. Rocha et al. (1996) investigated the presynaptic effects of paraoxon in rat cultured hippocampal neurons, and found that paraoxon (30 !-lmol-1 mM) blocked the ion channels of glycine, y-aminobutyric acidA (GABAA), N -methyl-D-aspartic acid (NMDA; glutamate receptor subtype), and nicotinic ACh receptors, but not the ion channels of kainate- and a-amino3-hydroxy-5-methyl-4-isoxazolepropionic acid -(AMPA) -like glutamate receptors. The authors suggested that the combined effects of paraoxon on the functions of severalligand-gated receptors may constitute actions relevant to the neurotoxicity of paraoxon. Even if these findings that reflect the direct effects of paraoxon on a number of ligand-gated ion channels are difficult to interpret, they are consistent with a number of recent observations that implicate the involvement of glutamate receptors in the effects of cholinergic neuronal stimulation (Jope et aI.,
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1986; Loikkanen et al., 1998; Naarala et aI., 1997; Savolainen et al., 1998; Solberg and Belkin, 1997). It is also of interest that the nerve agent, VX, but not soman or sarin, markedly inhibited [3H]NMS binding to muscarinic receptors and also reduced CH]s-piperone binding to dopamine D2 receptors in the rat striatum (Naseem, 1990). This OP seemed to differ somewhat from other OPs because they did not mirror its effects. It seems that OP compounds have both direct and indirect effects, mediated mainly via ACh accumulation subsequent to AChE inhibition. Even if OPs have some direct effects on a wide variety of receptors and their subtypes, it seems that they only slightly modify the overall effects of these agents. The vast majority of their actions are attributable to ACh accumulation on cholinergic receptors and subsequent glutamatergic activation. 50.7.3 EFFECTS OF BRAIN METABOLISM Chemical-induced seizures, and pure cholinergic agonist-induced brain activation, also in the absence of seizures, increase functional brain activity when assessed by local cerebral glucose utilization (LCGU; Churchill et aI., 1984a, b). Churchill et al. (1987) gave s.c. injections of soman or DFP to SD rats. Soman rapidly induced tonic-clonic convulsions, whereas DFP only occasionally induced transient seizurelike activity. Soman induced LCGU in most of the cortex, striato-pallidonigral pathway, limbic system and thalamic nuclei, whereas DFP increased LCGU in a very limited fashion, primarily in the dorsal striato-pallido-nigral pathway. Both soman- and DFPinduced facilitation of LCGU could be blocked or markedly inhibited by a mixture of muscarinic agonists, trimedoxime, atropine, and benactyzine, indicating that cholinergic stimulation was responsible for the activation of the striato-pallidonigral pathway, which is known to be important in muscarinic receptor-mediated convulsions. These authors suggested that even though soman and DFP activate this cholinergic pathway, only soman causes the spread of activity throughout the pathway, leading to overt motor convulsions. Possible explanations for this difference in response to these OPs are differential responses in cholinergic actions within specific brain regions or some noncholinergic action of soman. Also, delayed effects of soman on LCGU were studied by giving rats [t4C]2_ deoxyglucose (2-DG) prior to a single s.c. dose of 120 !-lg/kg (0.9 x LDso) sarin and by following the rats for up to 72 h. At later time points, there was a marked reduction in the LGGU, and this was associated with neuropathology in many brain regions, with the most marked damage occurring in the piriforrn cortex and the amygdala (Pazdernik et al., 1985; Samson et aI., 1985). These data suggest (Samson et al., 1985) that shortly after OP administration, there is an initial and marked increase of LCGU that reflects cerebral activation. Later there is a secondary reduction in cerebral LCGU. This is most likely due to convulsion-associated neuropathology and subsequent cell loss in the most severely affected brain regions. Khan and Hasan (1988) investigated the effects of methyl parathion at doses of 1.0, 1.5, or 2.0 mg/kg on the levels of gan-
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CHAPTER 50 Toxic Actions of Organophosphates
gliosides and glycogen in the cerebral hemisphere, cerebellum, brain stem, and spinal cord following seven daily i.p. injections. The authors observed a marked and dose-dependent decrease in the levels of gangliosides in all brain regions. An identical decrement in the levels of glycogen was found in different brain regions. The authors concluded that the OP-induced, dose-dependent, cholinergic brain stimulation was responsible for the changes in the levels of gangliosides and glycogen in different brain regions. The applied dose induced overt toxicity in the rats, including hyperexcitability, muscular fasciculations, tremors, and convulsions. After 6-7 days, all experimental rats became lethargic. Control rats exhibited no abnormality.
50.7.4 EFFECTS ON GENE EXPRESSION Convulsions may be associated with rapid and major increases in gene expression. This may be a direct consequence of convulsions, or at least be causally associated with them (Ceccatelli et al., 1989; Zimmer et al., 1997a, b). A high dose of soman [77.7 J.Lg/kg body weight (bw)] caused tonic-clonic convulsions in the exposed rats and induced a robust progressive expression of an immediate early gene, c-fos, a reliable indicator of neuronal activation (Greenberg et aI., 1986; Murphy et al., 1994), in the piriform cortex and the noradrenergic locus coereleus. Later, c-fos expression also occurred in the entorhinal cortex, the endopiriform nucleus, the olfactory tubercle, the anterior olfactory nucleus, and the main olfactory bulb. At 2 h the c-fos expression achieved its maximum and was then present also in the cerebral cortex, thalamus, caudate-putamen, and the hippocampus, brain regions, typically metabolically activated subsequent to soman exposure (Pazdernik et aI., 1985). At 8 hand beyond, c-fos expression returned to the control level (Zimmer et aI., 1997a, b). In general, c-fos promotes the transcription of additional genes, including those that encode proteins that are required for metabolic and physiologic activities of the cell. Thus, c-fos expression indicates that the cell is adapting to external stimuli by producing the proteins necessary for continued cellular function. Under extreme stress, c-fos also promotes the transcription of genes that encode proteins that are critical to cell survival (Sheng and Greenberg, 1990). Several other investigators (Arenander et aI., 1989; Greenberg et al., 1986; Seuwen et aI., 1990) also have shown that muscarinic receptor-mediated activation of PKC induces the immediate early genes c-fos and c-jun. These are genes that encode nuclear proteins (Maki et aI., 1987) and act in tandem as a dimeric complex that binds to a specific DNA consensus sequence in target genes to stimulate their transcription. Muscarinic receptor activation has induced c-fos expression in PC 12 pheochromocytoma cells (Arenander et al., 1989; Greenberg et aI., 1986), and both c-fos and c-jun expression in fibroblasts that express MJ muscarinic receptors (Seuwen et aI., 1990) and glial cell lines (Ashkenazi et aI., 1989). The changes induced by OPs may produce permanent changes in the gene levels in these cells. It is clear that cholinergic-induced convulsions are associated with increased expression of immediate
early genes. The exact role of these genes, whether they are consequences of neuronal excitation or causally linked with it, remains to be elucidated.
50.7.5 EFFECTS ON NEUROTRANSMITTER LEVELS IN THE BRAIN Exposure to OP compounds also has been shown to markedly affect the release and metabolism of a number cerebral neurotransmitters. Convulsive doses of soman (31.2 J.Lg/kg) inhibited guinea pig brain AChE by 90%, and elevated ACh levels in most brain areas, with levels of ACh remaining high for long periods of time. In all brain regions, soman reduced noradrenaline (NA) levels and the levels of dopamine (DA) were unchanged, but the levels of dopamine metabolites increased. The levels of 5-HT were unchanged, but those of its metabolites showed a modest increase. Changes in the levels of amino acid neurotransmitters correlate well with alterations in ACh levels: aspartate levels fell whereas those of GABA rose. It was concluded that these changes in the levels of several cerebral neurotransmitters are secondary to the initial increase in ACh content, and include an increased DA and 5-HT turnover, and release of NA and excitatory and inhibitory amino acid neurotransmitters (Fosbraey et aI., 1990). In contrast, a reversible AChE inhibitor, physostigmine, decreased levels of GABA in the hypothalamus, striatum, cerebellum, and the rest of the brain, whereas OP compounds paraoxon and soman had no effect on brain GABA levels. The lack of any effect of OPs on brain GABA levels may have been due to short-term exposure to paraoxon and soman (Coudray-Lucas et aI., 1984). Fosbraey et al. (1990) and EI-Etri et al. (1992) both reported that a dose of soman that induced tonic-clonic convulsions in most of the exposed animals also decreased cerebral noradrenaline levels in several brain regions in convulsing rats, but no change in NA was seen in nonconvulsing rats. Also consistent with the observations of Fosbraey et al. (1990), levels of DA and 5-HT were unaltered, but the levels of corresponding metabolites were elevated (EI-Etri et aI., 1992). Thus, it is possible that rapid and sustained NA release plays a role in the induction and/or maintenance of OP-induced convulsions, whereas the changes in the levels of the other neurotransmitters, notably 5-HT and DA are likely to be secondary to the convulsions. Consistent with the observations of EI-Etri et al. (1992), Shih and McDonough (1997) found that ACh levels increased rapidly after soman administration and that NA levels already started to decline 5 min after seizure onset and this process continued. However, levels of DA and its metabolites 3,4-dihydroxyphenylacetic acid and homovanillic acid also were elevated 5 min after seizure onset and thereafter. The brain aspartate levels were decreased within 20 min after the onset of seizures, and those of glutamate were decreased within 80 min after seizure initiation. Levels of GABA were markedly increased in the cortex, but concentrations of glutamine, glycine, and taurine were unchanged. The results are consistent with the notion that inhibition of AChE and elevation of ACh initiate the
50.7 Central Nervous System Cholinergic Effects of Organophoshorous Compounds and Other Cholinergic Agonists
seizure process, resulting in secondary changes in DA turnover and release of NE, and later changes in excitatory (aspartate, glutamate) and inhibitory (GABA) amino acid neurotransmitters. Thomsen and Wilson (1986) found earlier that repeated sublethal doses (300 Jl.g/kg bw) of paraoxon decreased transmitter release as measured via MEPPs. The authors suggested that this effect was attributable to a decrease in the transmitter store and mobilization ability that could account for the behavioral tolerance observed during long-term OP intoxication. Liu et al. (1994) studied catecholamine secretion and calcium influx in bovine adrenal chromaffin cells after their exposure to OP compounds, and found that catecholamine secretion and 4SCa2+ uptake evoked by a nicotinic receptor agonist DMPP (1,1dimethyl-4-phenylpiperazinium) were inhibited by both methyl parathion and malathion. The authors suggested that in addition to AChE, voltage-gated Ca2+ channels and nicotinic receptors also may be sites of OPs action in the mammalian nervous system.
50.7.6 EFFECTS ON ELECTROPHYSIOLOGY AND ASSOCIATED EVENTS A supralethal dose of soman (180 Jl.g/kg bw), 1 min prior to a 2 mg/kg dose of atropine and 30 min after a dose of the oxime HI-6 (125 mg/kg) caused epileptiform tonic-clonic seizures in rats. The convulsions were associated with extensive neuronal damage 1, 3, 10, or 30 days later. The severity of neuronal damage was associated with significantly, but transiently, increased /) (0-3.5 Hz) frequency in the electroencephalographs (EEGs) recorded 24 h after the exposure. Particularly sustained damage occurred in cortical areas, with piriform and perirhinal cortices exhibiting the most serious morphological alterations (McDonough et al., 1998). Even though there was a clear correlation between the occurrence of /) frequency and the severity of damage, EEG alterations could not be used to predict the morphological alterations. The EEG normalized within 10 days. When two doses of chlorphenvinfos, an OP insecticide [(2chloro-l-(dichlorophenyl)vinyl ethyl phosphate) (CVP)], were given at a 3 month intervals, EEG and behavioral alterations were induced in rabbits. The effects were less pronounced after the second dose. In rats, a symptomatic dose of 3 mg/kg of CVP given s.c. on 10 consecutive days induced subtle changes in complex behavior-neophobia in the open field test-and increased the EEG arousal response to an external painful stimulus (Gralewicz and Socko, 1997). Generalized convulsive status epilepticus is the most common and potentially most damaging form of status epilepticus. It exhibits a typical electrophysiological pattern: phases 1-5 are seen in the EEGs. Koplovitz and Skvorak (1998) found that soman induced all stages in 12 out of 15 rats, but phases 2-5 occurred in all rats. The findings suggest that the sequence of EEG changes is independent of the initiating cause, reflect a common electrical response to generalized convulsive status epilepticus and point to a common underlying neurochemical
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mechanism. These conclusions do not conflict with the hypotheses that acetylcholine and the convulsive excitatory amino acids may be the common final pathway in generalized convulsions even if the agent that initiates the cascade of events differs (Savolainen et aI., 1998; Solberg and Belkin, 1997). OP compounds also induce more subtle electrophysiological changes, which can be reflected in potential problems with learning and memory. Subchronic ex vivo and in vitro exposure to parathion resulted in an increase of the field excitatory postsynaptic potentiation (EPSP) after a subthreshold tetanic stimulation. Furthermore, the occurrence of a late effect reduced long-term potentiation (LTP; Schmuck et aI., 1998). LTP is a long-lasting increase in the efficacy of synaptic transmission that is considered to underlie the plastic changes associated with learning and memory (Doyere and Laroche, 1992). In the CAl region of the hippocampus, tetanus-induced LTP is dependent on activation of the NMDA receptors and on the rise in intracellular calcium concentration (Collingridge et aI., 1992; Lynch et aI., 1983). Schmuck et al. (1998), suggested that the effects of parathion on EPSP and LTP may be due to elevated synaptic ACh and stimulation of cholinergic muscarinic receptors, even if no persistent neurobehavioral changes that correlate with these electrophysiological effects have been demonstrated (lvens et aI., 1998).
50.7.7 SEIZUROGENIC AND BEHAVIORAL EFFECTS Dissociation between the motor and electrical aspects of convulsive status epilepticus induced by OP compounds is problematic in many cases. Sparenborg et al. (1993) used the nerve agent soman (200 Jl.g/kg bw) to induce cholinergic convulsions and electrical brain discharges in guinea pigs, and administered cholinergic or GABAergic antagonists to investigate the association between behavior and electrical brain events. All animals that received soman without other treatments developed severe status epilepticus associated with continuous electrographic seizure activity. Despite the presence of continual motor convulsions in all animals challenged with soman and pretreated with diazepam, increased electrographic seizure activity did not take place in most of these animals. Likewise, scopolamine also inhibited soman-induced increased electrographic activity, even though it did not terminate the motor convulsions. Neuronal necrosis was found in the hippocampus, thalamus, amygdala, and cerebral and piriform cortices in animals that exhibited both increased electrographic activity and motor convulsions, but not in those that exhibited motor convulsions alone. Thus, it is important to note that overt convulsions may take place, at least in experimental animals, with no electrographic alterations present in the brain that lead to actual necrosis. Soman given as an infusion at a dose of 13.1 Jl.g/kg bw to adult baboons induced the onset of intoxication within 2-3 min, manifested as hyperactivity and severe grand mal convulsions (Anzueto et aI., 1986). Stamper et al. (1988) exposed preweanling rat pups to daily doses of 1.3 or 1.0 mg/kg of parathion during postnatal days
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CHAPTER 50 Toxic Actions of Organophosphates
5-20, a time period critical for development of behavior and biochemical maturation of the cholinergic systems in the brain. Even though the doses of parathion were quite high, they were not lethal to these experimental animals. Had the exposure been to paraoxon, the situation would have been different. The exposure resulted in a dose-dependent reduction in AChE activity and muscarinic receptor binding in the brain, and was later associated with multiple deficits in spatial memory. Thus, impairment of memory functions due to alterations in brain cholinergic systems were seen even though the doses of OP did not provoke severe signs of toxicity. Repeated sublethal doses of DFP or soman were given to rats for 4 weeks to explore behavioral tolerance toward anticholinesterase OP compounds (Van Dongen and Wolthuis, 1989). Even though there was progressive AChE inhibition, the rats' behavior, as measured with shuttle-box performance, remained virtually normal 24 h after the DFP injections throughout the study. However, behavioral tolerance did not develop toward soman. The results indicate that the differences between DFP- and soman-induced behavioral tolerance could not be explained by (1) inhibition of AChE de novo synthesis, (2) muscarinic receptor downregulation, (3) differences in the number of nicotinic receptors, or (4) the activity of phosphorylphosphatase (DFPase or somanase) in different organs. The authors concluded that the behavioral tolerance induced by DFP is probably due to presynaptic alterations or increased synthesis of de novo synthesis of OP-binding proteins (see also Meldrum and Garthwaite, 1990).
50.8 MODULATION OF EFFECTS OF ORGANOPHOSPHOROUSCOMPOUNDS Whereas OP anticholinesterase compounds are indirect cholinergic agonists that induce their effects through stimulation of cholinergic muscarinic and nicotinic receptors, it is conceivable that modulation of the responses of these receptors could be used to modulate the effects of these agents. The cholinergic muscarinic receptors also communicate with other receptor systems, especially glutamatergic (Felipo et aI., 1998; Savolainen et aI., 1998; Solberg and Belkin, 1997) and GABAergic receptors (Hirvonen et aI., 1993; McDonough and Shih, 1997; Savolainen et aI., 1988a, b), and thus their responses can be modulated not only by antagonists of muscarinic receptors, but also with agonists or antagonists of other receptors. Furthermore, because muscarinic receptors are G-protein-coupled (Gilman, 1987) Ca2+ mobilizing receptors (Berridge, 1989), they permit modulatory interactions with compounds that affect this protein (Jope et al., 1989). It is not likely that nicotinic receptors can be modulated other than via their antagonists such as d-tubocurarine and its analogs (Taylor and Brown, 1989). However, nicotinic receptor functions may be modulated directly by OP compounds and indirectly by cholinergic muscarinic agonists (see Katz et al., 1997).
50.8.1 ATTENUATION OF CHOLINERGIC EFFECTS BY MUSCARINIC RECEPTOR INHIBITION Several studies indicate that atropine, the nonsubtype-selective antagonist of all muscarinic receptors, is effective in blocking muscarinic receptor-mediated effects of OP compounds (John son and Lowndes, 1974; Lundy et aI., 1978; Savolainen et aI., 1988b; Shih, 1990, 1991; Shih et aI., 1991a, b). High doses of atropine (32 mg/kg) given prior to soman (100 mg/kg) effectively blocked soman-induced convulsions in rats and attenuated the increased cerebral PI hydrolysis (Savolainen et aI., 1988a, b). In another rat study (Shih, 1990), atropine (12 mg/kg) given prior to soman (100 I-lg/kg) did not block soman-induced convulsions and did not affect AChE activity or degree of OP-induced inhibition. The difference in the effectiveness of atropine to prevent soman-induced convulsions in these two studies (Savolainen et al., 1988a, b; Shih, 1990) may have been due to the different doses of atropine used. Longterm exposure to sublethal doses of soman results in behavioral supersensitivity to atropine, possibly due to downregulation in the number of muscarinic receptors in the brain (Modrow and McDonough, 1986). Pazdernik et al. (1983) reported that administration of large doses of soman (2 x LDso) to rats protected with TAB [a mixture of trimedoxime (TMB-4), atropine, and benactyzine] resulted in approximately twofold reductions of LCGU in most brain regions. This was in contrast to the marked increase in LCGU that occurred in conjunction with the convulsions after a LDso dose of soman. The results indicated that TAB is effective in protecting against soman-induced convulsions, but only at the expense of a severe decrease in LCGU after soman exposure. These findings are consistent with those of Shih et al. (1991a, b) (see subsequent text), who demonstrated a marked protection against soman-induced toxicity by atropine or benactyzine. When rats were given a convulsive dose of soman subsequent to pretreatment by diazepam, atropine, or benactyzine, both diazepam and benactyzine prevented convulsive activity, but atropine had no effect. All pretreatments attenuated LCGU with a unique pattern. In the pathology phase, 72 h postsoman, the marked reduction in LCGU and the conspicious brain damage associated with soman-induced convulsions was minimized with all three pretreatments (Pazdernik et aI., 1986). Shih et al. (1991 a, b) studied the efficacy of several drugs, especially antagonists of cholinergic muscarinic receptors, to prevent soman-induced toxicity, OP-induced convulsions, and death. In the absence of atropine sulfate, only tertiary anticholinergic drugs (scopolamine, trihexyphenidyl, biperidene, benactyzine, bentzatropine, azaprophen and aprophen), ceramiphen, carbetapentane, and the NMDA-receptor antagonist MK-80l were effective anticonvulsants. In the presence of atropine sulfate, the benzodiazepines (diazepam, midazolam, clonazepam, loprazolam, and alprazolam), mecamylamine, flunazirine, phenytoin, clonidine, CGS 19755 and Organon 6370 protected against soman-induced convulsions. The authors concluded that central muscarinic cholinergic mechanisms are pri-
50.8 Modulation of Effects of Organophosphorous Compounds marily involved in eliciting the convulsions following exposure to soman and that subsequent recruitment of other excitatory neurotransmitter systems and loss of inhibitory control may be responsible for sustaining the convulsions and for producing the subsequent brain damage. These overall conclusions are consistent with a number of research articles (see Hirvonen et al., 1993; Olney et aI., 1986; Savolainen et al., 1988a, b, 1995, 1998; Solberg and Belkin, 1997). 50.8.2 ATTENUATION OF CHOLINERGIC EFFECTS BY ACETYLCHOLINESTERASE REACTIVATION Because OP-induced cholinergic intoxication is due to inhibition of the enzyme that hydrolyzes ACh (i.e., AChE), it is natural that attempts have been made to develop drugs that alleviate OP-induced AChE inhibition (Shih et aI., 1991a, b). Emergency treatment is always necessary when dealing with OP poisoning. Even if cholinergic muscarinic agonists such as atropine and scopolamine are the first line of treatment in these cases, a supplementary treatment goal is AChE reactivation, which is necessary in many cases (see Shih et al., 1991a, b; Ecobichon, 1996). Oximes (pralidoxime chloride or 2-PAM, pralidoxime methanesulfonate or P2S) are give intravenously to reactivate the nervous tissue AChE, and to alleviate nicotinic and muscarinic symptoms. The use of oximes is not necessary in mild intoxications, but may greatly amplify the effect of direct cholinergic antagonists to attenuate the cholinergic symptoms. The therapeutic action of oximes is due to its ability to reactivate AChE without having marked toxic actions of their own. The basic requirements for a reactivating molecule of AChE consist of a rigid structure with a quaternary ammonium group and an acidic nucleophile. This nucleophile must be complementary to the phosphorylated (i.e., deactivated) enzyme in such a way that the nucleophilic oxygen is positioned close to the electrophilic phosphorus atom. These requirements lead to the development of 2-PAM. The reactivation is an equilibrium reaction, where the oxime reacts with the phosphorylated enzyme or with the free unbound OP ester. Development of new oximes has produced P2S, obidoxime [bis( 4-formy 1- N -methy1pyridinium oxime)ether dichloride], TMB-4 [N,N-trimethylene bis(pyridine-4-aldoxime)bromide], and, more recently, the H-series compounds (Ecobichon, 1996; Shih et al., 1991a, b; see subsequent text). Ligtenstein and Moes (1991) utilized two OPs as AChE inhibitors. One is S-diethylaminoethyl-O-cyclohexyl-methylphosphonothionate, a tertiary amine that readily penetrates the CNS and, therefore, exhibits both central and peripheral AChE inhibiting properties. The other compound is the methiodide derivative of this agent, which has a strong and pHindependent ionic character that does not allow it to penetrate though the blood-brain barrier (BBB); thus, it does not gain access to the brain. Atropine sulfate inhibited convulsions and
1031
toxicity induced by the former compound, but not lethality induced by the latter. The oxime (HI-6) used as the reactivator of AChE was more effective than atropine in both cases. Atropine and HI -6 together had a synergistic effect in the case of brain-penetrating compounds, but not in the case of an ionized OP agent. The authors concluded that a combination of atropine and an oxime is an effective combination when an OP has both central and peripheral actions. Several other investigators have confirmed the synergistic protective effect of the combination of atropine and various oximes against OPinduced poisoning in both humans and experimental animals (De Neef and Porsius, 1982; Endres et al., 1989; Lallement et aI., 1997; Shih et al., 1991a, b; Singh et al., 1998). When diazepam was combined with atropine and an oxime, it further amplified the protection offered by the atropine-oxime combination (Lallement et aI., 1994a, b, 1997). In humans, administration of the oxime pralidoxime produced neurophysiological amelioriation in 11 out of 15 cases. The authors (Singh et aI., 1998) also emphasized that although the administration of three compounds-pralidoxime, magnesium sulfate, and pancuronium-resulted in the reversal of the neuroelectrophysiological defects, only pralidoxime was of true therapeutic value, but because of its short duration of action, frequent administration is required.
50.8.3 ATTENUATION OF CHOLINERGIC EFFECTS BY INCREASING GABAERGIC TONE IN THE CNS Diazepam (5 mglkg) effectively prevented soman-induced (100 J.lglkg) overt convulsions, consistent with earlier observations (Lundy et aI., 1978), and markedly attenuated the somaninduced facilitation of PI signalling in rat brains (Savolainen et aI., 1988a, b). Doebler et al. (1985), in turn, observed that diazepam almost completely prevented soman-induced depletion of RNA in two brain regions, but did not affect the mean time of soman-induced death or 24-h survival. These data suggest that excessive neuronal activity per se may underlie the genesis of soman-induced central metabolic impairments. We could also conclude from these results that it is possible to dissociate epi1eptiform activity from the lethal action of soman. These data also provide evidence that cerebral GABAergic tone is involved in controlling brain metabolic events and the initiation of convulsions. These conclusions are in agreement with the findings of Gant et al. (1987), who suggested that there may be an interaction between GABA receptors and OP compounds in a manner that modifies OP toxicity. 50.8.4 THE ROLE OF EXCITATORY AMINO ACIDS Excitatory amino acid (EAA) neurotransmitters have been implicated in the initiation and propagation of seizures, and the involvement of the NMDA class of glutamatergic EAA receptors has been demonstrated in several models of seizures (Din-
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CHAPTER 50 Toxic Actions of Organophosphates
gledine et aI., 1990). Glutamate, acting through a number of EAA receptor subtypes including NMDA receptors, is known to evoke neuronal cell death (Meldrum and Garthwaite, 1990). Glutamate also has been implicated in cholinergic agonistinduced seizures (Hirvonen et al., 1990, 1993; McDonough and Shih, 1997; Savolainen et aI., 1988a, b, 1998; Solberg and Belkin, 1997). Dizocilpine (MK-801), a noncompetitive inhibitor of NMDA receptors, was given to guinea pigs either before or after a convulsive dose of soman. Pretreatment of the animals with MK-801 did not prevent or delay the onset of electrical seizure activity, but did diminish its intensity and led to its rapid termination. A large dose of MK-801 (5 mglkg) was even able to prevent the appearance of seizures. Posttreatment of the animals with MK-801 prevented, arrested, or reduced seizure activity, convulsions, and neuronal death in a dose-dependent manner, pointing to a possible role for the NMDA receptor in the spread and maintenance of cholinergically induced seizures, although NMDA receptors themselves are not involved in seizure initiation. Shih (1990) made similar observations on the effects of MK-801 on soman-induced convulsions in a rat model. He found that MK-801 was an especially effective antidote against soman-induced convulsions when administered in conjunction with diazepam. Deshpande et al. (1995) observed that soman did not induce cytotoxicity in cultured hippocampal neurons, whereas exposure of these cells to glutamate did not induce death in 80% of the exposed cells. Memantine, a glutamate receptor antagonist, significantly protected the neurons against glutamate toxicity. When rats were pretreated with memantine 1 h prior to soman (0.9 x LD50), the severity of convulsions as well as brain damage were significantly reduced, indicative of a role for glutamatergic receptors in soman-induced neurotoxicity. The role of glutamatergic receptors in soman-induced chlolinergic neurotoxicity was emphasized by the finding that a noncompetitive antagonist of NMDA receptors (thieny1cylohexylpiperidine) offered useful protection against somaninduced (62.5 J.lglkg) seizures and lethality in guinea pigs (Carpentier et al., 1994; Lallement et aI., 1994a, b). Lallement et al. (1994a, b) found that administration of NBQX (2,3-dihydroxy-6-nitro-7 -sulfamoylbenzoquinoxaline), a selective inhibitor of AMPA glutamatergic receptors, prevented the onset of soman-induced convulsions in rats. When NBQX was given after the soman injection, it also reduced the intensity of convulsions. These results clearly indicate that glutamatergic receptors are involved in OP-induced convulsions. This conclusion can be extended from NMDA glutamatergic receptors (see Carpentier et aI., 1994; Loikkanen et al., 1998; Naarala et al., 1997) to include non-NMDA glutamatergic receptors (Lallement et al., 1994a, b) also. These studies (Carpentier et al., 1994; Lallement et al., 1994a, b) also indicate that atropine amplifies the attenuation of soman-induced convulsions obtained with both glutamatergic NMDA (TCP) and non-NMDA-receptors (NBQX) antagonists, highlighting the interaction between g1utamatergic and cholinergic receptors in the initiation and propagation of convulsions (see Felipo et aI., 1998; Fig. 50.8).
Soman
Brain
Excessive stimulation of muscarinic receptors
Neuronal hyperexcitatlon
Convulsions
Excess of glutamate
NMOA-receptor activation
Figure 50.8 Pictorial representation of the role of glutamate in neuronal damage caused by intoxication with an organophosphate nerve agent, soman, subsequent inhibition of acetylcholinesterase, and acetylcholine accumulation in the central nervous system. Typically, excessive activation of muscarinic receptors leads to neuronal stimulation, subsequent convulsions, release of excessive amounts of glutamate, and NMDA receptor activation, which is followed by an influx of calcium to the neurons and subsequent programmed (apoptotic) or necrotic neuronal death. These terminal events are preceded by increased production of cellular messengers, such as nitric oxide and phosphoinositides, accumulation of free calcium in the cell, oxidative stress, alterations in gene expression, and serious dysfunction in the maintenance of cellular homeostasis.
50.8.5 DOWNREGULATION OF MUSCARINIC RECEPTORS
Lim et al. (1991) observed that simultaneous and continuous administration of physostigmine and trihexyphenidyl had protective effects against soman-induced toxicity in guinea pigs. Combination of these two compounds provided greater protection against soman-induced toxicity than either of the compounds alone. The antimuscarinic properties of trihexyphenidyl and protection of AChE activity by physostigmine against irreversible enzyme inhibition by soman may be responsible for the protective effect of this drug combination. When physostigmine was given alone, it also provided protection against so-
50.9 Effects on Autonomic Ganglia man toxicity; this protection was amplified when scopolamine was given simultaneously with physostigmine. Physostigmineinduced protection was due to the well-known tolerance toward its AChE-inhibiting properties, whereas scopolamineinduced protection was due to its antimuscarinic effects. Jointly these protective effects seem to provide marked protection against soman-induced convulsions and lethality (Philippens et al., 1998). Shih et al. (1993) reported that the antiparkinsonian drugs biperidene and trihexyphenidyl provide protection against soman-induced toxicity; this is likely due to their antimuscarinic effects rather than their secondary effects of cerebral dopamine turnover. Aronstam et al. (1987) even found that clonidine, a centrally acting a-2 adrenergic agonist (Buccafusco and Aronstam, 1986), given prior to soman administration prevented the decrease in receptor number and decreased the extent of AChE inhibition caused by soman. The authors concluded that clonidine may protect AChE from irreversible inhibition by soman, thereby decreasing the extent of cholinergic overstimulation with its attendant downregulation of muscarinic receptors. Buccafusco and Aronstam (1986) found that clonidine provided protection against soman-induced toxicity including tremors, convulsions, and Straub tail, as well as excessive salivation that results from activation of peripheral muscarinic receptors. Atropine further amplified the clonidineinduced protection against soman intoxication. Clonidine noncompetitively inhibited AChE in vitro and in vivo, and also inhibited ligand binding to cortical muscarinic receptors in vitro. The authors concluded that the protective effects of clonidine are likely to involve multiple effects, including blockade of acetylcholine release and postsynaptic muscarinic receptors, and transient inhibition of AChE.
50.8.6 AMPLIFICATION OF CHOLINERGIC-INDUCED CONVULSIONS Inhibition or attenuation of receptor-mediated responses is usually easier than amplification. In the case of cholinergic receptors, a number of antagonists exist for both cholinergic muscarinic and nicotinic receptors (McDonough and Shih, 1997; Minton and Murray, 1988). Even if amplification of cholinergic receptor-mediated responses does not have any therapeutic or practical value, understanding the mechanisms involved in amplification of cholinergic-induced responses may provide valuable insight into the receptor mechanisms. Amplification of muscarinic receptor-mediated cellular and molecular events, and subsequent behavioral changes such as convulsions have been intensively investigated. Honchar et al. (1983) initially showed that pretreatment of rats with lithium chloride greatly amplified cholinergic-induced convulsions and further facilitated the associated cerebral PI signalling. Savolainen et al. (1988a, b, 1991) and Hirvonenetal. (1989,1990,1993) showed in a number of studies that lithium pretreatment greatly amplified convulsions induced in rats by OPs such as soman, DFP, paraoxon, and malaoxon, as well as convulsions produced by
1033
a direct cholinergic agonist, pilocarpine. In the case of several OPs, pretreatment of rats with lithium increased their sensitivity to cholinergic-induced convulsions by about two- to threefold (Savolainen et aI., 1991). Several investigators (Honchar et al., 1983; Berridge, 1989) hypothesized that lithium's ability to decrease the threshold for cholinergic-induced convulsions or amplify these convulsions is due to the noncompetitive inhibition of PI metabolism. This blockade of myo-inositol-1-phosphatase leads to a dramatic accumulation of myo-inositol-1-phosphate in the brain. Later studies provided evidence that most likely this is not the case. Instead, the interaction between lithium and cholinergic stimulation seems to take place at the G-protein level, that is, lithium may affect the coupling mechanisms of cholinergic muscarinic receptors (Jope, 1988; Jope et aI., 1986; Savolainen et aI., 1995).
50.9 EFFECTS ON AUTONOMIC GANGLIA Heppner and Fiekers (1992) explored the prevention and the reversal of the effects of soman on the electrical properties of sympathetic ganglion neurons in vitro from the adult bullfrog Rana catesbeiana. Atropine pretreatment (10 !J.M) blocked the soman-induced decrease in the membrane potential, membrane resistance, and duration of the afterhyperpolarization. Atropine posttreatment restored the soman-induced decrease in the membrane potential, but was ineffective in reversing either the change in membrane resistance or the duration of the afterhyperpolarization. These authors concluded that the effects of soman on the electrical properties of these neurons are mediated by the activation of muscarinic receptors, and that following receptor activation, different cellular mechanisms may be involved in the regulation of the electrical properties of the neuron. The results also indicated that pre-OP rather than postOP treatment with atropine was more effective in blocking these direct effects of soman.
50.10 EFFECTS ON THE NEUROMUSCULAR ENDPLATE AND MUSCLE CELLS Acetylcholinesterase inhibition can lead to severe neuromuscular dysfunction. For example, the OP anticholinesterase, paraoxon, produces a dose-dependent necrosis in the rat skeletal muscle fibers even after a single administration. The pathology, initially concentrated around the neuromuscular endplate region, already is evident 30 min after paraoxon administration. By 24 h, a generalized breakdown of muscle fiber architecture is apparent with an accompanying infiltration of phagocytes. Electrophysiological studies indicate that paraoxon increases neurotransmitter release and causes spontaneous and impulserelated antidromic nerve activity, both of which can be reduced markedly by reactivation of inhibited AChE with the oxime 2-PAM. Furthermore, the severity of the pathology is positively correlated to the degree and duration of AChE inhibition: the
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CHAPTER 50 Toxic Actions of Organophosphates
critical loss of AChE activity is 85% provided the inhibition lasts at least for 2 h (see Wecker et aI., 1978). 50.10.1 ELECTROPHYSIOLOGICAL EFFECTS
E1ectromyographic (EMG) features of OP poisoning include (1) repetitive activity (RA) in response to single nerve stimuli and (2) two types of decremental responses to repetitive nerve stimulation (RNS). The smallest amplitude can either occur in the second response, with a subsequent gradual recovery (decrement-increment phenomenon), or the amplitude can progressively decline toward further responses (decrement phenomenon; Besser et aI., 1989). Paraoxon and fenthion both induced RA and decrements in RNS in rats as the main EMG findings in both types of OP intoxications. Various types of impairment of neuromuscular transmission seem to coexist, most likely due to variations present at distinct stages of anticholinesterase poisoning (De Bleecker et al., 1994). Paraoxon (10- 9 -10- 3 M) produced repetitive nerve terminal activity as well as accelerated MEPP frequency at the neuromuscular junction in an isolated muscle. At high concentrations, paraoxon also depolarized the muscle membrane, especially in the endplate regions. The depolarizing effects of paraoxon, due to ACh accumulation, are mediated through nicotinic ACh receptors because they can be attenuated by d-tubocurarine. They occur when muscle membrane AChE is inhibited to about 32% of the control levels. The fact that pretreatment with 2-PAM prevented paraoxon-induced depolarization indicates that AChE on the outer membrane was essential for the effect because 2-PAM, which is a quaternary ammonium compound, does not cross the cellular membrane. These findings support the proposal that both the pre- and the postsynaptic actions of paraoxon are consequences of AChE inhibition and are not caused by any direct effects of paraoxon (Laskowski and Dettbam, 1979). 50.10.2 NEUROMUSCULAR TRANSMISSION
Van Dongen et al. (1988) explored the role of de novo synthesis of AChE in the spontaneous recovery of neuromuscular transmission in isolated diaphragms from soman-intoxicated rats. Within 10 min after soman administration, neuromuscular transmission was completely blocked, but AChE activity in endplate and endplate-free regions recovered linearly during the next 3 h, and neuromuscular transmission also improved. This recovery could not be attributed to synthesis of AChE because de novo synthesis of AChE had been blocked by cycloheximide and reinhibition of AChE was ensured with an additional dose of soman. Thus, this recovery in muscular activity seemed to be independent of AChE activity. Smith and Wolthuis (1983) exposed rhesus monkeys to soman and then treated the animals with the oxime HI6. Soman exposure induced neuromuscular blockade in intercostal and diaphragm muscle strips that could be partially relieved by their exposure to the AChE activating oxime.
50.10.3 EFFECTS OF ACETYLCHOLINESTERASE REACTIVATORS
Rats poisoned with sarin, soman, or VX and then administered two bispyridium oximes, BDB-27 and HGG-12, were tested for their ability to restore neuromuscular function in isolated phrenic nerve-diaphragm preparations. These results were compared with the traditional oximes HI-6 and TMB-4. BDB-27 was equal or superior to HI-6 in sarin, soman, and VX poisoning and better than TMB-4 in tabun poisoned animals. However, recovery of function after HGG-12 was equal to HI-6 only in soman poisoning, but much less pronounced against neuromuscular blockade induced by the three other OPs (Jovanovic, 1983). Several investigators (Bhattacharyya et aI., 1990; Ozkutlu et al., 1995) also showed that hemicholinium-3 and its analogs provide protection against OP-induced toxicity. These drugs, which block choline uptake, were able to provide significant antagonism against OPs and also to produced a profound nicotinic receptor desensitization as reflected in reductions to indirect or directly evoked EPC (endplate current). Whereas these effects were voltage-dependent, they may be due to reduction of endplate permeability, rather than to blockage of ACh synthesis, because it has been claimed that hemicholinium can also block nicotinic receptor associated ion channels in their open form (Bhattacharyya et al., 1990). 50.10.4 MUSCARINIC RECEPTOR DOWNREGULATION
Single and short-term inhalation of soman in rats reduced subsequent sensitivity of smooth muscle to contraction induced by cholinergic stimulation. A single exposure to 8.51 mg/m 3 of soman for 45 min inhibited bronchial smooth muscle AChE by 85% and reduced the contraction induced by carbachol and/or ACh by 70-80%. This exposure did not induce alterations in eH]QNB binding. However, consecutive exposures to low doses of soman (0.45-0.63 mg/m3) markedly reduced [3H]QNB binding in bronchial smooth muscle, in addition to evoking AChE inhibition and inhibiting the ACh- and carbachol-induced concentration of smooth muscle (Aas et al., 1987). These results indicate that short-term inhalation of relatively low concentrations of soman reduces the number of cholinergic muscarinic receptors in the peripheral cholinergic system.
50.11 EFFECTS ON CARDIOVASCULAR AND RESPIRATORY SYSTEMS AND TEMPERATURE CONTROL Anticholinesterase OP compounds have widespread effects on both the cardiovascular and respiratory systems because both are under the control of the parasympathetic and sympathetic
50.12 Summary and Conclusions
divisions of the autonomic nervous system. In autonomic ganglia, ACh serves as the primary neurotransmitter. Furthermore, ACh also serves as a transmitter at the neuromuscular endplate in both skeletal and smooth muscles. Administration of 32, 80, or 160 I-l-g/kg of soman to guinea pigs resulted in respiratory arrest followed by circulatory failure and death. Atropine treatment restored the circulatory parameters and improved respiration. However, atropine was ineffective after very high doses of soman (10 x LDso). These results indicate that soman-induced respiratory depression is mainly due to CNS affects and that a significant neuromuscular block develops only at very high doses of soman. The circulatory disturbances are mainly caused by bradycardia that results from peripheral muscarinic stimulation in the heart (Worek and Szinicz, 1993). ACh reduced atrial contraction by 82.5, 50.8, and 41.5% in rats, guinea pigs, and rabbits, respectively. The ECso values for the negative inotropic effect of ACh were 3.3 x 10-7 M in rat and guinea pig atria, and 4.1 x 10-6 M in rabbit atria. However, there was no correlation between the species differences in the negative inotropic effect of ACh in the atria and the density or affinity of AChE or the characteristics of muscarinic receptors. Inhibition of atrial AChE with soman reduced the ECso of ACh by threefold in all species, but did not change the maximal inotropic effect of ACh. The authors hypothesized that species differences in the negative inotropic effect of ACh may be due to differences in the coupling between myocardial muscarinic receptors and the ion channels that mediate negative inotropy (Maxwell et al., 1991). Worek et al. (1994) observed that atropine treatment was very effective in improving the respiratory function after a dose of 60 I-l-g/kg of tabun, but ineffective when tabun was given at a dose of 300 I-l-g/kg. However, the circulatory parameters were restored almost completely in all atropine treated groups. However, when atropine was combined with an oxime, such as obidoxime, Hl6 7, or HI 6, the antidotal efficacy of the combination was markedly improved over atropine alone. Atropine was especially effective in restoring circulation, but respiration could be improved only in cases of intoxication with low doses of tabun. The results of this study demonstrated the remarkable synergistic protective effect of atropine and an AChE reactivator against tabun-induced circulatory and respiratory failure, both of which seem to be strictly mediated via the cholinergic system. Chiou and Li (1994) studied the effects of cholinolytic agents (i.e., ganglion blocking drugs) on cardiovascular effects of DFP intoxication in rats. The lethal action of DFP (8 mg/kg) was partially or completely prevented by pretreatment with hexamethonium (10 mg/kg), trimethaphan (80 mg/kg), and mecamylamine (30 mg/kg). The combined effects of these drugs with an AChE reactivator 2-PAM (100 mg/kg) greatly improved the prophylaxis of presynaptic cholinolytic drugs against DFP intoxication. Even though the exact mechanism of action of these drugs remains to be elucidated, DFP induced cardiovascular suppression before neuromuscular blockade, indicating that the cardiovascular system is more sensitive to the effects of OPs than the neuromuscular junction. Kubinec et al. (1987) found that when rats were treated with paraoxon, signif-
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icant decreases in heart cholinergic muscarinic receptor densities occurred both in the atria and ventricles after several daily injections; these changes were associated with decreased levels of cardiac AChE activity. Thus, we cannot exclude the possibility that downregulation of cholinergic muscarinic receptors may be involved in the effects of OPs on the cardiovascular system. Bartholomew et al. (1985) explored the mechanism of OPinduced depression of the respiratory center in the brain stem, bronchoconstriction, increased secretions in the airways, and paralysis of the respiratory musculature in rats. After 14 daily doses of malathion (400 mg/kg), the activities of AChE were 26-28% in the striatum (ST), hippocampus (HI), and cortex (CX), but 41 % in the brain stem of the corresponding control values. Furthermore, there was a marked reduction in the numbers of mAChRs in ST, HI, and CX, but not in the brain stem. The authors concluded that the mechanisms by which the respiratory center in the brain stem adapts to the effects of OP exposure differ from responses to OPs in the rest of the CNS. Clement (1991, 1993a, b) studied the effects of OP anticholinesterase agents on temperature control in experimental animals. Several studies have shown that subchronic administration of a cholinergic agonist or an AChE inhibitor frequently results in tolerance to its behavioral effects. Tolerance to the OP compound DFP was characterized by a decrease in the symptoms of poisoning, such as salivation, lacrimation, and hypothermia (Gupta and Dettbam, 1986; Lomax et al., 1986; Overstreet et aI., 1973), and a decrease in the number but no change in the affinity of muscarinic receptors in various regions in the brain (Churchill et aI., 1984a, b; Yamada et aI., 1983). Likewise, following acute (Aronstam et al., 1987) or subchronic (Churchill et aI., 1984a, b) administration of soman, there was a decrease in the number of muscarinic receptors in the brain (i.e., receptor downregulation). Repeated administration of soman and DFP induced tolerance toward OPinduced hypothermia after a few doses, and this effect showed cross-tolerance with a direct cholinergic agonist, oxotremorine (Clement, 1991). In later studies, Clement (1993a) suggested that soman-induced hypothermia may be due to the recovery of AChE, perhaps from the assembly of previously synthesized precursors. Soman hypothermia appears to be due to muscarinic receptor activation and can be partially, but not completely, antagonized by atropine. Thus, soman-induced hypothermia is primarily a muscarinic receptor-related event (Clement, 1993b).
50.12 SUMMARY AND CONCLUSIONS Insecticidal OPs and nerve gases cause their effects by inhibiting AChE, thereby leading to ACh accumulation. The rapidity of this accumulation depends on the inhibition rate of the AChE by the OP; thus, different OPs can have different effects. The signs and symptoms of OP intoxication are due to activation of cholinergic muscarinic and nicotinic receptors. Muscarinic receptor-mediated effects include convulsions, smooth muscle
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CHAPTER 50
Toxic Actions of Organophosphates
activation, and increased secretions, whereas nicotinic receptor- necessary if effective therapeutic agents and measures are to be induced effects are due to activation of autonomic ganglia developed. and neuromuscular endplates, leading to muscular dysfunctions and cardiovascular effects such as bradycardia. Muscarinic reREFERENCES ceptors are membrane integral proteins, so-called transmembrane G-protein-coupled receptors, most of which mediate their effects through generation of calcium-mobilizing second Aas, P., Veiteberg, T. A., and Fonnum, F. (1987). Acute and sub-acute inhalation of an organophosphate induced alteration of cholinergic muscarinic recepmessengers, inositol phosphates. Some of the muscarinic retors. Bioehem. Pharmaeol. 36(8), 1261-1266. ceptors, however, mediate their effects by inhibiting adenylate Abdallah, E. A. M., Jett, D. A., Eldefrawi, M. E., and Eldefrawi, A. T. (1992). Differential effects of paraoxon on the M 3 muscarinic receptor and its efcyclase, thereby decreasing cellular cAMP levels. Nicotinic refector system in rat submaxillary gland cells. l. Bioehem. Toxieol. 7(2), ceptors are ion channels that open to permit preferential in125-132. flux of sodium into the cell. Muscarinic receptor activation Ahlbom, J., Fredriksson, A., and Eriksson, P. (1995). Exposure to an can be blocked by muscarinic antagonists such as atropine, but organophosphate (DFP) during a defined period in neonatallife induces peratropine is not effective in inhibiting nicotinic receptor actimanent changes in brain muscarinic receptors and behaviour in adult mice. Brain Res. 677, 13-19. vation. d-Tubocurarine and its analogs can be used to block nicotinic receptors. Both receptor systems contribute to OP- Aldridge, W. N., ed. (1996). "Mechanisms and Concepts in Toxicology." Taylor & Francis, London. induced lethalities. The muscarinic receptors are involved in Aldridge, W. N., and Reiner, E. (1972). "Enzyme Inhibitors as Substrates. convulsions and their associated neuronal damage-depression Interaction of Esterases of Organophosphorus and Carbamic Esters." NorthHolland, Amsterdam. of the respiratory center, and increased bronchoconstriction and bronchial secretions. Nicotinic receptors cause muscle fascicu- Alha, A. (1967). Forensic-chemically detected poisonings in Finland in 1966. Acta Neurol. Scand. 43(Suppl. 31), 133. lations, flaccid paralysis, and hypertension. Cholinergic musAntonelli, T., Beani, L., Bianchi, L., Pedata, F., and Pepeu, G. (1981). Changes carinic receptors have widespread effects on cellular function in synaptosomal high affinity choline uptake following electrical stimulain the CNS and periphery by altering cell signalling, gene tion of guinea-pig cortical slices: Effect of atropine and physostigmine. Br. l. Pharmaeol. 74,525-531. expression, cellular metabolism, levels of brain neurotransmitters, and brain electrophysiology. Muscarinic receptor activa- Anzueto, A., Berdine, G. G., Moore, G. T., Gleiser, c., Johnson, D., White, C. D., and Johanson, W. G., Jr. (1986). Pathophysiology of soman tion can be modulated by a number of antagonists, including intoxication in primates. Toxieol. Appl. Pharmacol. 86, 56-68. muscarinic antagonists, which may be especially effective when Arenander, A. T., de Vellis, J., and Hershiman, H. R. (1989). Induction of e-fos and TIS genes in cultured rat astrocytes by neurotransmitters. l. Neurosci. combined with oximes, so-called AChE reactivators. Whereas Res. 24, 107-114. brain cholinergic systems seem to be at least under partial Aronstam, R. S., Smith, M. D., and Buccafusco, J. L. (1987). Clonidine preGABAergic control, benzodiazepines can also be used as anvents the short-tenn down regulation of muscarinic receptors in the mouse tidotes for OP-induced convulsions. Recent observations indibrain induced by the acetylcholinesterase inhibitor soman. Neurosei. Lett. cate that cholinergic muscarinic stimulation may be the trigger 78, 107-112. that evokes convulsions, but that subsequent glutamate release Ashkenazi, A., Ramachandron, J., and Capon, D. J. (1989). Acetylcholine analogue estimates DNA synthesis in brain-derived cells via specific musmay be necessary for their maintenance and propagation. Ancarinic receptor subtypes. Nature 340, 146--150. tagonists of both NMDA and non-NMDA glutamate receptor Bakry, N. M. S., El-Rashidy, A. H., Eldefrawi, A. T., and Eldefrawi, M. E. antagonists have been proven to be effective antagonists of mus(1988). Direct actions of organophosphate anticholinesterases on nicotinic and muscarinic acetylcholine receptors. l. Bioehem. Toxieo!. 3,235-259. carinic receptor-mediated intoxication, including OP-induced intoxication. Glutamate also may be important in cholinergic- Baraban, J. M., Snyder, S. H., and Alger, B. E. (1985). Protein kinase C regulates ionic conductance in hippocampal pyramidal neurons: Electroinduced toxicity because this amino acid transmitter is involved physiological effects of phorbol esters. Proe. Natl. Aead. Sci. U.S.A. 82, in neuronal programmed cell death, or apoptosis (see Nicotera 2538-2542. et al., 1999). This OP-induced neurotoxicity can be prevented Bartholomew, P. M., Gianutsos, G., and Cohen, S. D. (1985). Differential cholinesterase inhibition and muscarinic receptor changes in CD-l mice at least partially by using glutamatergic antagonists. This inmade tolerant to malathion. Toxieol. Appl. Pharmacol. 81, 147-155. formation provides new insights into cholinergic toxicity and Berridge, M. J. (1987). Inositol triphosphate and diacylglycerol: Two interactalso opens new vistas for research to explore mechanisms of OP ing second messengers. Ann. Rev. Biochem. 56, 159-193. toxicity. The essential elements of apoptosis, such as DNA frag- Berridge, M. J. (1989). Inositol trisphosphate, calcium, lithium, and cell signalling. lAMA 262, 1834-1841. mentation, mitochondrial failure, activation of caspases, and cytochrome c release, need to be clarified in this context. The role Berridge, M. J., and Irvine, R. F. (1989). Inositol phosphates and cell signalling. Nature 341, 97-205. of nicotinic receptors, which are also important in OP-induced Besser, R., Gutmann, L., Dillmann, U., Weilemann, L. S., and Hopf, H. C. intoxication, has been better clarified, but it also requires further (1989). End-plate dysfunction in acute organophosphate intoxication. Neuresearch. Nicotinic receptor activation in autonomic ganglia and rology 39, 561-567. neuromuscular endplates can be blocked by nicotinic antago- Bhattacharyya, B., Sokoll, M. D., Flynn, S. J. R., Nyanda, A. M., Lee, T., Cannon, J. G., and Long, J. P. (1990). Mechanism for antagonism of paraoxon nists, although this is more difficult to achieve than blockade of by hernicholinium-3 analogues. Arch. Int. Pharmaeodyn. 308, 149-167. muscarinic receptors. Because many OP compounds, especially Blanchet, G., Baubichon, D., Mavet, S., Morelis, P., and Lemercier, G. (1986). the insecticides and nerve agents, are often extremely toxic, a Modulation of the number of muscarinic receptors in mouse neuroblastoma cells by soman. Bioehem. Pharmaeol. 35(22), 4077-4081. thorough understanding of their basic mechanisms of action is
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Sterri, S, H" Berge, G., and Fonnum, F. (1985). Esterase activities and soman toxicity in developing rat. Acta Pharmacol. Toxicol. 57, 136-140. Storm, 1. E., Rozman, K. K., and Doull, 1. (2000). Occupational exposure limits for 30 organophosphate pesticides based on inhibition of red blood cell acetylcholinesterase. Toxicology 150, 1-29. Taylor, P., and Brown, 1. H. (1989). Acetylcholine. In "Basic Neurochemistry" (G. Sieger, B. Agranoff, R. W. Albers, and P. Molinoff, eds.), 4th ed., pp. 203-232. Raven Press, New York. Thomsen, R. H., and Wilson, D. F. (1986). Chronic effects of paraoxon on transmitter release and the synaptic contribution to tolerance. J. Pharmacol. Exp. Ther. 237(3), 689-694. Unwin, N., Toyoshima, C., and Kubalek, E. (1988). Arrangement of the acetylcholine receptor subunits in the resting and desensitized states, determined by cryoelectron microscopy of crystallized Torpedo postsynaptic membranes. J. Cell BioI. 107(3), 1123-1138. Van Dongen, C. 1., Valkenburg, P. w., and van Helden, H, P. M. (1988). Contribution of de novo synthesis of acetylcholinesterase to spontaneous recovery of neuromuscular transmission following soman intoxication. Eur. J. Pharmacol. 149,381-384. Van Dongen, C. 1., and Wolthuis, O. L. (1989). On the development of behavioral tolerance to organophosphates I: Behavioral and biochemical aspects. Pharmacol. Biochem. Behav. 34,473-481. Viana, G. B., Davis, L. H., and Kauffman, F. C. (1988). Effects of organophosphates and nerve growth factor on muscarinic receptor binding number in rat pheochromocytoma PCI2 cells. Toxicol. Appl. Pharmacol. 93,257-266. Wade, 1. v., Samson, F. E., Nelson, S. R., and Pazdemik, T. L. (1987). Changes in extracellular amino acids during soman- and kainic acidinduced seizures. Journal oJ Neurochemistry 49(2), 645-650. Ward, T. R., Ferris, D. 1., Tilson, H. A., and Mundy, W. R. (1993). Correlation of the anticholinesterase activity of a series of organophosphates with their ability to compete with agonist binding to muscarinic receptors. Toxicol. Appl. Pharmacol. 122,300-307.
1041
Wecker, L., Laskowski, M. B., and Dettbam, W.-D. (1978). Neuromuscular dysfunction induced by acetylcholinesterase inhibition. Federat. Proc. 37(14), 2818-2822. Whalley, C. E., and Shih, T.-M. (1989). Effects of soman and sarin on high affinity choline uptake by rat brain synaptosomes. Brain Res. Bull. 22, 853858. Worek, F., and Szinicz, L. (1993). Analysis of cardiovascular and respiratory effects of various doses of soman in guinea-pigs: Efficacy of atropine treatment. Arch. Int. Pharmacodyn. 325, 96-Il2. Worek, F., Kirchner, T., and Szinicz, L. (1994). Treatment of tabun poisoned guinea-pigs with atropine, HL6 7 or HI 6: Effect on respiratory and circulatory function. Arch. Toxicol. 68,231-239. World Health Organization (1986). "Organophosphorus insecticides: A general introduction." Environmental Health Criteria 63, World Health Organization, Geneva. Yagle, K., and Costa, L. G. (1996). Effects of organophosphate exposure on muscarinic acetylcholine receptor sUbtype mRNA levels in the adult rat. NeuroToxicology 17(2), 523-530. Yamada, S., Isogai, M., Okudaira, H., and Hayashi, E. (1983). Correlation between cholinesterase inhibition and reduction in muscarinic receptors and choline uptake by repeated diisopropylfluorophosphate administration: Antagonism by physostigmine and atropine. J. Pharmacol. Exp. Ther. 226, 519-525. Zimmer, L. A., Ennis, M., EI-Etri, M., and Shipley, M. T. (I 997a). Anatomical localization and time course ofJos expression following soman-induced seizures. J. Comp. Neurol. 378,468-481. Zimmer, L. A., Ennis, M., and Shipley, M. T. (l997b). Soman-induced seizures rapidly activate astrocytes and microglia in discrete brain regions. J. Comp. Neurol. 378, 482-492.
CHAPTER
51 Clinical Toxicology of Anticholinesterase Agents in Humans Marcello Lotti Universita degli Studi di Padova Several chemicals display anticholinesterase activity among which organophosphorus esters (OPs) represent the vast majority because they are widely used and easily available. Since their introduction after World War 11 as insecticides (Khurana and Prabhakar, 2000), countless publications have described their clinical toxicology in humans. Initially, most case reports dealt with accidental and occupational poisoning, whereas later the majority of poisonings were the result of suicide attempts. Most recently, attention has been directed to possible subtle effects caused by low-level long-term exposures such as those encountered in the workplace. Concurrently, an even larger amount of experimental studies have contributed to our understanding of anticholinesterase toxicology. Therefore, anticholinesterase agents, and OPs in particular, undoubtedly represent the most extensively studied class of chemicals in toxicology and it is not surprising that a very large number of reviews, textbooks, book chapters, and other general publications, in addition to research papers, have appeared over the years on their clinical toxicology (the most recent reviews include: Bardin et aI., 1994; Brown and Brix, 1998; De Bleecker et aI., 1992a; ECETOC, 1998; Eyer, 1995; Gunderson et aI., 1992; Jamal, 1997; Karalliede, 1999; Karalliedde and Senanayake, 1999; Marrs, 1993; Millard and Broomfield, 1995; Ray, 1998a, b; Steenland, 1996). Nevertheless, some controversies still exist on some aspects of both the clinical toxicology and the treatment of OP poisoning. In particular, conflicting results have been published on long-term sequelae of acute exposure and on possible effects of long-term exposures that do not cause overt cholinergic toxicity, and different opinions exist on the uses of antidotes in the treatment of acute poisonings. This chapter deals with the clinical aspects of OP toxicology in humans, mentioning, where appropriate, those of other anticholinesterase agents and supporting experimental evidence in animals as well. Moreover, this chapter does not replace the specific chapters of the previous editions of this book (Hayes, Handbook of Pesticide Toxicology Volume 2. Agents
1982; Hayes and Laws, 1991), where additional details on OP toxicology, as derived from the older literature, can be found.
51.1 THE CHOLINERGIC SYNDROME The cholinergic syndrome is characterized by overstimulation of cholinergic receptors throughout the body. It may be the consequence of single or repeated exposures to a variety of chemicals such as OPs and carbamates, which inhibit acetylcholinesterase (AChE) at the synaptic level. As a consequence, the level of acetylcholine increases, although the amount and time course of neurotransmitter accumulation may vary widely among various districts of the cholinergic system (Stavinoha et aI., 1976). Thus, the clinical picture that results from this excess of neurotransmitter may be quite variable in presentation of signs and symptoms, at the onset, in time course, and in outcome. This clinical polymorphism largely depends on the chemical involved and on the dose. 51.1.1 ETIOLOGY 51.1.1.1 AntichoIinesterases as Pesticides and Warfare Agents Although the use of anticholinesterase OP insecticides has declined during the last two decades, particularly in agriculture, they still represent an important class of pesticides, which account for about 10% of all active ingredients currently used as pesticides. However, several OPs that caused poisoning in humans in the past are believed to be no longer manufactured or marketed (Tomlin, 1997). The long inventory of acute intoxications in humans involves a variety of OPs. Table 51.1 lists examples of OPs, ranked according to their chemical structure, that caused acute toxicity in humans. Most compounds that belong to group A are chemical warfare agents and are highly toxic. Most pesticides belong
1043
Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
1044
CHAPTER 51
Clinical Toxicology of Anticholinesterase Agents
Table 51.1 Classification of Organophosphorus Esters;a General Formula: RI,,-
p-"'" 0
R2/
--- X
A. Compounds where X = halogen or CN, CNS, etc.
\. RI = alkoxy, R2 = alkyl Example: sarin (isopropyl methylphosphonofluoridate), nerve gas (Okumura et aI., 1996) 2. RI and R2 = alkoxy Example: DFP (diisopropyl phosphorofluoridate), laboratory chemical and discontinued drug (Moore, 1956) 3. RI
= alkylamide, R2 = alkoxy
Example: tabun (ethyl N-dimethylphosphoroamido cyanidate), nerve gas (Compton, 1987) 4. RI and Rz = mono- or dialky1amido Example: mipafox (N NI-diisopropyl phosphorodiamido fluoridate), laboratory chemical (Bidstrup et aI., 1953). B. Compounds where X
= alkyl, alkoxy, or aryloxy
\. Alkoxydialkyl or dialkoxyalkyl compounds Example: trichlorfon (dimethyl (2,2,2-trichloro-I-hydroxyethyl) phosphonate), insecticide and drug (Vasilescu and Florescu, 1980) 2. Trialkyl compounds and dialkoxy, aryloxy compounds Example: dichlorvos (dimethyl 2,2 dichlorovinyl phosphate), insecticide (Wadia et aI., 1987) C. Thiol- and thiono-phosphorus compounds
1. Thiol compounds Example: demeton-S-methyl (S-[2-(ethylthio)ethyl] dimethyl phosphorothioate), insecticide (Weir et aI., 1992) 2. Thiono compounds Example: parathion (diethyl O-(4-nitrophenyl)phosphorothioate), insecticide (Namba et aI., 1971) 3. Thiol - thiono compounds Example: malathion (dimethyl S-(l,2-dicarboxylethyl) phosphorodithioate), insecticide (Baker et aI., 1978) D. Derivatives of pyrophosphorous acid Example: sulfotepp (tetraethyl dithionopyrophosphate), insecticide (Namba et aI., 1971) E. Compounds containing a quaternary nitrogen Example: ecothiophate (diethyl-S-2-trimethyl ammonium-ethyl phosphorothioate iodide), drug (Gesztes, 1966). aNerve gases belong to group A, whereas most common pesticides belong to groups B and C. Examples of OPs which caused poisoning in humans are given. Source: Holmstedt (1963).
to groups Band C, among which dimethoxy OPs are the most common. In contrast to other OPs, dimethoxy OPs are less toxic because of certain biochemical characteristics (i.e., relative high speed of reactivation of dimethoxyphosphorylated AChE; see Section 51.1.2.2). In Table 51.2, the World Health Organization (WHO) recommended classification of OP pesticides by hazard is reported. The classification is based on acute oral and dermal toxicity in rats and the physical state of product or formulation. Confirmation of the hazard severity of these chemicals has often been obtained from clinical observations. OPs have also been used as warfare agents and large amounts are thought to be stockpiled in arsenals worldwide. They were developed during the 1950s and were used as recently as the early 1990s in a terrorist attack in a Tokyo subway (Suzuki et aI., 1995). Moreover, during the Persian Gulf War, soldiers were thought to have been exposed to low levels of nerve gases (NIH, 1994). 51.1.1.2 Anticholinesterases as Drugs Anticholinesterase drugs have a variety of indications in clinical medicine, and toxicity may arise from their therapeutic
uses. However, although used in the past, nowadays drugs do not include OPs, but use reversible inhibitors of AChE such as physostigmine, pyridostigmine, neostigmine, and edrophonium. Their acceptability has been established in four areas: atony of the smooth muscle of the intestinal tract and urinary bladder, myasthenia gravis, glaucoma, and termination of effects of competitive neuromuscular blocking drugs (Taylor, 1996a). Moreover, reversible anticholinesterases are also used for treatment of overdoses of atropine and other anticholinergic drugs (such as phenothiazines and tricyclic antidepressant; Nilsson, 1982). Pyridostigmine was administered on a large scale during the Persian Gulf War to soldiers as a prophylaxis for nerve gas attacks to obtain reversible carbamylation of AChE, thereby preventing irreversible phosphorylation (Keeler et aI., 1991). Several AChE inhibitors have been used or are on clinical trial for the treatment of Alzheimer's disease. The rationale is to ameliorate cognitive performance in these patients by increasing synaptic levels of acetylcholine (Summers et aI., 1986). Tacrine (tetrahydroaminoacridine) is the most extensively studied drug; however, controversies arose from initial
51.1 The Cholinergic Syndrome
1045
Table 51.2 Organophosphorus Pesticides and Risk of Cholinergic Syndromea Classification
Common name
Extremely hazardous
Chlorethoxyfos, chlormephos, coumaphos, disulfoton, EPN, ethoprophos, fenamiphos, fonofos, mevinphos, parathion, parathion-methyl, phorate, phosphamidon, sulfotepp, tebupirimfos, terbufos
Highly hazardous
Azinphos-ethyl, azinphos-methyl, cadusafos, chlorfenvinphos, demeton-S-methyl, dichlorvos, dicrotophos, edifenphos, famphur, heptenophos, isazofos, isofenphos, isoxathion, mecarbam, methamidophos, methidation, monocrotophos, omethoate, oxydemeton-methyl, pirimiphos-ethyl, propaphos, propetamphos, thiometon, triazophos, vamidothion
Moderately hazardous
Anilofos, bilanafos, butamifos, chlorpyrifos, cyanophos, diazinon, dimethoate, ethion, etrimfos, fenitrothion, fenthion, formothion, methacrifos, naled, phenthoate, phosalone, phosmet, phoxim, piperophos, profenofos, prothiofos, pyraclofos, pyrazophos, quinalphos, sulprofos
Slightly hazardous
Acephate, azamethiphos, iprobenfos, malathion, pirimiphos-methyl, pyridaphenthion, trichlorfon
aWHO recommended classification by hazard (WHO, 1998). Based on acute oral and dermal toxicity in rats and the physical state of product or formulation, pesticides are ranked as follows: Extremely hazardous: 5 mg/kg or less if solid; 20 mg/kg or less if liquid Highly hazardous: 5-50 mg/kg if solid; 20-200 mg/kg if liquid Moderately hazardous: 50-500 mg/kg is solid; 200-2000 mglkg ifIiquid Slightly hazardous: over 500 mglkg if solid; over 2000 mglkg if liquid Only compounds in use or being developed are reported (Tomlin, 1997).
reports (Relman, 1991) and clinical results have so far not been convincing (Davis et aI., 1992; Maltby et aI., 1994). Finally, pyridostigmine salicylate was shown to rapidly reverse clinical signs of central and peripheral anticholinergic toxicity caused by a variety of drugs, such as antidepressants, antiparkinsonians, antihistamines, and antispasmodics, and toxic plants, such as mushrooms, potato sprouts, and bittersweet (Granacher and Baldessarini, 1975).
OPs exert their toxic action by interfering with cholinergic transmission. The molecular mechanism of cholinergic toxicity involves the interaction of OPs with AChE, which almost completely explains all the signs and symptoms of acute OP poisoning. Symptoms and signs are related to excess acetylcholine and the consequent overstimulation in all districts of the central and peripheral nervous systems (CNS and PNS) where acetylcholine acts as a neurotransmitter.
residues that line about 40% of the gorge, providing an array of low-affinity binding sites (Sussman et al., 1991). The synaptic function of AChE is to remove the neurotransmitter acetylcholine. The hydrolysis of acetylcholine involves acetylation of the serine residue followed by restoration of the active center of AChE. The time required for hydrolysis of acetylcholine at the neuromuscular junction is less than a millisecond (Taylor, 1996a). However, because the cholinergic transmission is involved in a variety of functions that require specific features at different sites, AChE distribution and hydrolysis time may vary, depending on the type of response needed in a given district. Moreover, differences in the cholinergic transmission also depend on several other factors, including synthesis storage and release of acetylcholine, distribution of different receptors, and types of signal transduction (Brown and Taylor, 1996; Taylor, 1996a, b). Cholinergic transmission is involved in the stimulation of skeletal muscle and autonomic ganglia (nicotinic), autonomic effector cells in the white muscles, cardiac conduction system, and secretory glands (muscarinic), and CNS neurons.
51.1.2.1 Acetylcholinesterase and Cholinergic Functions
51.1.2.2 Chemistry and Biochemistry of Anticholinesterases
AChE is an elongated molecular structure formed by heterologous subunits that is localized mainly in the outer basal lamina of the synapse. The enzyme is highly concentrated at the neuromuscular junction, and it is synthesized in both nerve and muscle. A single gene encodes the catalytic subunits of AChE (Taylor, 1996a). The atomic structure, determined by x-ray analysis, reveals that the active site catalytic triad (serine, histidine, and glutamate) lies near the bottom of a deep and narrow gorge that reaches halfway into the protein. Substrates and inhibitors are drawn to the active site of the enzyme by an aromatic guidance mechanism that is formed by 14 aromatic
OPs that display anticholinesterase activity are triesters of phosphoric acid with the general structure shown in Table 51.1 and in Figs. 51.1 and 51.2. RI and R2 vary (e.g., alcohols, amides) as does the X moiety (e.g., fluorine, phenoxy). The nomenclature of OPs and other anticholinesterases, and their classification may follow various schemes (Chambers, 1992; Edmundson, 1988; Holmstedt, 1963). Figure 51.3 and Table 51.3 illustrate the biochemical interactions between OPs and AChE, and Table 51.3 offers some examples of them. OPs react covalently with AChE by phosphorylating the serine residue at the catalytic center. This occurs essentially in the
51.1.2 PATHOGENESIS
1046
CHAPTER 51
Clinical Toxicology of Anticholinesterase Agents
same manner that acetylcholine acetylates AChE (Aldridge and Reiner, 1972; reactions 1 and 2 in Fig. 51.3). Affinity constants vary, depending on the OP involved. However, in contrast to the acetylated enzyme, which rapidly yields acetic acid and restores the catalytic center, the phosphorylated enzyme is stable and catalytic activity recovers very slowly (reaction 3 in Fig. 51.3). The rate of spontaneous reactivation depends on the chemistry of the attached phosphoryl residue. In the case of dimethoxyphosphorylated AChE, rates are much higher compared with those of phosphoryl residues with longer carbon chains (diethoxy, dipropoxy, etc.), where reactivation might not occur at all. Rates of spontaneous reactivation are even highcr for carbamylated AChE, which restores its catalytic activity more rapidly than the phosphorylated enzyme, although still much more slowly than acetylated AChE. For this reason, carbamates are defined as reversible AChE inhibitors. Rates of spontaneous reactivation should be considered in conjunction with those of a further nonenzymatic reaction that occurs on phosphorylated AChE. This reaction, called aging, involves the loss of one alkyl group and leads to stabilization of the phosphorylated
enzyme (reaction 4 in Fig. 51.3). The degree of AChE inhibition and its duration in vivo largely depend on these rates: when the rate of spontaneous reactivation is higher than that of the aging reaction, almost complete recovery of activity is expected. On the contrary, if the rate of aging is higher than that of spontaneous reactivation, then irreversible inhibition takes place (Table 51.3). Rates of spontaneous reactivation and aging may have clinical relevance in the case of poisoning by dimethoxy OPs (and by carbamates, where carbamylated AChE does not undergo aging) because diagnostic and therapeutic attitudes may differ from those in cases of poisoning by other OPs (see Section 51.1.3.8).
51.1.3 CLINICAL MANIFESTATIONS 51.1.3.1 General Features The clinical pictures of acute OP poisoning reflect the degree of accumulation of neurotransmitter that causes cholinergic overstimulation in various organs. Early effects are characterized by stimulation or facilitation at various sites, which are followed, at higher concentrations of anticholinesterases, by inhibition or paralysis (Taylor, 1996a). The relationship between OP toxicity and nervous tissue AChE inhibition is influenced by many factors. In general, 50-80% of AChE must be inactivated before symptoms are noted. Brain AChE activity around 10-15% of normal is associated with severe toxicity, and below 10% with coma, seizures, respiratory failure, and death. Lethal exposures in the absence of treatment have been estimated to correspond to approximately 30-50 times the minimal symptomatic exposure for most OPs (Holmstedt, 1959). A pharmacological description of acute poisoning is reported in Table 51.4, where signs and symptoms are ranked as muscarinic, nicotinic, and CNS. Signs and symptoms of acute poisoning usually appear within minutes or a few hours of exposure, depending on the chemical involved, route of exposure, and dose. Unusual cases of suicide attempts by intravenous injections of OPs have been reported, where symptoms and signs appeared quite early even if doses were (probably) small (Gtiven et aI., 1997; Lyon et aI., 1987).
0-0 s
- F; - CN: - CH 2CH 2 C2H ;
@- 10
2
: - CH=CCI 2
Figure 51.1 Chemical structure of organophosphorus esters. Oxygen can be replaced with sulfur at all sites. Oxygen atoms bound to RI and R2 can also be replaced with nitrogen or may be absent on one (phosphonates) or both groups (phosphinates). R groups can be the same or different. Examples of the chemistry of R groups and of leaving groups (X) are also given (also see Table 51.1).
RO _ _
~O
P -;:::::::::-RO ~ ---------X Phosphates R __
~O
S
.;:::70
RNH --.........
RNH/
R _____
P '"-X
RO.../"
~O
P
O/~X
R/~X
Sulphonates
Phosphinates
---...X
Phosphonates
Phosphoroamidates R __
~O
P
R
'-.N-C~
R/
0
-----X
Carbamates
Figure 51.2 Chemical structure of anticholinesterase agents. Some are pesticides (phosphates, phosphoroamidates, phosphonates and carbamates); others are not (phosphinates and sulfonates). Carbamates are reversible inhibitors of AChE.
51.1 The Cholinergic Syndrome Cases have been reported of delayed onset of symptoms and signs often associated with prolonged toxicity and relapses. Five patients displayed mild cholinergic toxicity within a few hours of ingestion of dichlofenthion, but severe toxicity did not appear until 40-48 h afterward (Davies et aI., 1975). Two patients died and cholinergic symptomatology persisted up to 48 days in the survivors. In one patient, blood dichlofenthion was detected for 75 days after poisoning, whereas in another patient, inhibition of both plasma and red blood cell (RBC) cholinesterases was detected for 66 days. In another case of combined dermal and inhalation exposure to fenitrothion, mild symptoms appeared after 2 days and more severe symptoms appeared over the next 3 days. Relapses of cholinergic signs occurred on days 11 and 17 (Ecobichon et aI., 1977). In a case of ingestion of fenitrothion, delayed onset of poisoning was confirmed (Sakamoto et aI., 1984).
1047
A case of suicide attempt was described where the patient started to complain of symptoms 5 days after ingestion of fenthion and relapse of cholinergic toxicity occurred on day 24 (Merrill and Mihm, 1982). In another suicide attempt with fenthion, the patient had initial symptoms a few minutes after ingestion that became severe 31 h later and lasted for 18 days (Mahieu et aI., 1982). In a further case of poisoning by fenthion, blood levels of the chemical measured 11 days after intoxication were 1000 times higher than those measured upon admission to the hospital (Martinez-Chuecos et aI., 1992). In a case of suicide attempt with isofenphos, severe symptoms occurred 24 h after intramuscular injection and lasted for 10 days (Zoppellari et aI., 1997). All these cases involved OPs with slow disposal and long persistence in the fat; both these pharmacokinetic characteristics are justified by the high partition coefficient of these
Table 51.3 Examples of Interactions of Some OPs with Human AChEa Spontaneous Organophosphates
Inhibition
Phosphorylated AChE
(AChE ISO 10- 6 M) 0 CH,O, 11 '" P-O-CH=CCI, CH,O
<;5-
C,H,O
'" p-o·
Cl
Dichlorvos
E-0 -P, 11 /OCH,
0.95
(t1/2 hours)
(t1/2 hours)
0.85
3.9
OCR,
0
Clorpyrifos-oxon
E-0 -P, 11 /OC,H,
0.007
58
41
OC,H,
Cl 0 i-C,H,O, 11 /P-F i-C,H,Q
a Data
Aging
0
Cl
0 C,H,O, 11
reactivation
DFP
~
0.83
0 II/OC,Hri -P, OC,Hri
No reactivation at 6
4.6
from Lotti and Johnson (1978), Capodicasa et al. (1991), and EPA (1992).
o
o 11
11
O-R
ACh E-OH+X-O-P::: O-R
'1
~hE-OH-X-O-~ ::: O-RJ ~ 0
O-R
O-R
ACh E-OH + HO-P::: O-R 2
--------+~
r ~ ACh 3
E-O-p:::
o 11
O-R
O-R
+X
0-
ACh E-O-p::: O-R Figure 51.3 General representation of biochemical interactions between OPs and AChE. Reaction 1 leads to the formation of Michaelis complex and reaction 2 leads to phosphorylated AChE. Rates of these reactions indicate the affinity of enzyme for a given OP. Reaction 3 is spontaneous reactivation of AChE, which is usually very slow, although in the case of dimethoxy phosphorylated AChE the speed of reactivation is higher. Reaction 4 leads to a stable, negatively charged phosphorylated AChE (aging of phosphorylated AChE).
+R
1048
CHAPTER 51
Clinical Toxicology of Anticholinesterase Agents
Table 51.4 Signs and Symptoms of Organophosphate Poisoning Manifestations
Signs and symptoms
Muscarinic Respiratory system
Wheezing, dyspnea, cyanosis, bronchorrhea, bronchospasm, pulmonary edema
Gastrointestinal system
Anorexia, nausea, vomiting, diarrhea, abdominal pain, fecal incontinence
Cardiovascular system
Bradycardia, hypertension
Urinary system
Urinary incontinence
Glands
Hypersalivation, hyperlacrimation, increased sweating
Pupils
Miosis, unreactive to light
Nicotinic Red muscles (including respiratory muscles) Central nervous system
Weakness, fasciculations, twitching, tachycardia, hypertension Headache, drowsiness, dizziness, confusion, blurred vision, slurred speech, ataxia, coma, convulsions, depression and block of respiratory center
compounds (Tomlin, 1997). These chemical and toxicological characteristics of certain OPs should be kept in mind either when patients are first observed or when antidotal treatment is discontinued. The first signs to appear are usually muscarinic, which may or may not be in combination with nicotinic signs. The incidence of signs and symptoms is variable, depending on the dose, the chemical involved, and the time after exposure when detected. Respiratory failure is the hallmark of the clinical picture of severe OP poisoning, whereas mild poisoning and/or early stages of an otherwise severe poisoning may display no clear-cut signs and symptoms. Therefore, diagnosis is made through symptom recognition, followed by grading of poisoning severity, although the latter is only a guide for immediate treatment and has no prognostic value (Bardin et aI., 1987; Bardin and van Eeden, 1990; Lotti, 1991; Minton and Murray, 1988). Miosis is observable in more than 80% of patients and in the case of mild poisoning, may represent the only sign (Rengstorff, 1985, 1994). Anorexia, nausea, and vomiting are reported in 40-80% of patients and may also be the only and earliest signs of poisoning (Bardin et aI., 1987; Grob and Harvey, 1958; Hayes et aI., 1978; Namba et al., 1971; Ohbu et aI., 1997; Okumura et aI., 1996; Saadeh et aI., 1996; Tafuri and Roberts, 1987; Tsao et aI., 1990). Diarrhea and abdominal pain are also reported in 20-60% of patients. Hypersalivation is reported in more than 60% and excessive sweating in more than 30% of patients. Hyperlacrymation is less frequent (lO-30%). Bronchial hypersecretion and respiratory distress usually follow other muscarinic signs, but not always. Urinary and fecal incontinence can be observed in the most severe cases. Weakness is the only nicotinic symptom that appears at early stages of poisoning. Muscle fasciculations appear later when the clinical picture becomes more severe. CNS signs such as coma and convulsions appear after muscarinic and nico-
tinic symptoms, although early CNS symptoms may include headache, dizziness, and blurred vision. Based on the recording and evaluation of this constellation of signs and symptoms, and on the circumstantial evidence of exposure, diagnosis is relatively easy. The course of the illness depends on the toxicological characteristics of the OP, on the severity of the clinical picture, and on the promptness and efficacy of treatment. 51.1.3.2 Respiratory Failure
As result of combined nicotinic, muscarinic, and CNS cholinergic overstimulation, severe OP poisoning results in respiratory failure. Thus, bronchoconstriction, bronchorrhea, pulmonary edema, fasciculations and paralysis of respiratory muscles (the diaphragm in particular), and depression of the brain respiratory center all contribute to respiratory insufficiency. This clinical condition usually develops within 24 h of exposure and within a shorter period from the onset of signs and symptoms. It should be distinguished from a somewhat delayed respiratory failure (one to several days after poisoning) that characterizes the intermediate syndrome, which has a different pathophysiology (see Sections 51.2.2 and 51.2.3.1). The severity of respiratory failure in 52 patients with acute OP poisoning was graded according to the presence/absence of acidosis, which in turn predicted the survival rate, which is much higher in patients with hypoxemia only (Goswamy et aI., 1994). In the same study some markers at the time of physical examination were identified as indicative of the need for artificial ventilation. These markers include miosis, hypotension, fasciculations, unconsciousness, and low plasma cholinesterases. As a single sign, fasciculation was the most significant prognostic marker for ventilator requirement and final outcome. This observation indirectly suggests that respiratory muscle failure is more important than any other factor in causing respiratory insufficiency. However, some authors indicated the failure of central respiratory drive as the important factor (Grob, 1956; Tsao et aI., 1990; see Section 51.1.3). The use of plasma cholinesterase (ChE) levels to assess poisoning severity and to predict the development of respiratory failure, as also suggested by others (Tsao et aI., 1990), should be discouraged because low plasma ChE do not necessarily correlate with the severity of cholinergic overstimulation (see Section 51.1.3.5 and 51.4.3.2). 51.1.3.3 Cardiac Manifestations
Given the cholinergic innervation of the heart (Lefkowitz et aI., 1996), several types of cardiac alterations are included in the clinical picture of acute poisoning. In a review of 168 cases of acute poisoning that involved a variety of OPs (including dimethoate, methylparathion, trichlorphon, sevin, mevinphos, dichlorvos, and malathion), 134 patients showed electrocardiographic abnormalities, including prolonged QT interval and ST and T abnormalities. Fifty-six patients had arrhythmias 3-15 days after poisoning: ventricular extrasystoles and ventricular tachycardia with "torsade de pointes" characteristics have been observed. Bradyarrhythmias were less
51.1 The Cholinergic Syndrome frequent, although two cases of AV block occurred (Kiss and Fazekas, 1979). QT prolongation and/or polymorphous ventricular arrhythmias also were observed in 14 patients of another series of 15 cases of OP poisoning that involved phosdrin, parathion, and phosphamidon (Ludomirsky et al., 1982). A case in which sinus bradycardia, AV dissociation, idioventricular rhythm, multiform ventricular extrasystoles, and prolongation of the PR, QRS, and QT intervals was observed. Polymorphic ventricular tachycardia, characterized by extreme variability of QRS morphology and changes in the RR interval, was also present (Brill et al., 1984). QT prolongation was recorded in 97 patients in a series of 223 OP poisonings, when electrocardiograms (ECGs) were examined in retrospect (Chuang et al., 1996). Mevinphos, parathion, methamidophos, ethyl 4-nitrophenyl phenylphosphonothioate (EPN), and other unspecified OPs were involved. The authors concluded that patients who presented with such ECG changes have higher mortality and a higher incidence of respiratory failure compared with those without QT prolongation. These patients were also among those who had the highest plasma cholinesterase inhibition. However, whereas patient ranking was based on plasma cholinesterase levels, and QT changes and respiratory failure also were detected in patients graded as having very mild to moderate poisoning, it is not possible to appreciate the prognostic value, if any, of these ECG changes. In another review of 46 cases of OP and carbamate poisoning, 31 patients developed QT prolongation with or without ST and T changes, among which, 4 had ventricular tachycardia and 2 had ventricular fibrillation. Sinus tachycardia and bradycardia were equally represented (Saadeh et al., 1997). In 50 fatal cases of OP intoxication, focal myocardial damage was observed at autopsy, including pericapillar hemorrhage, micronecrosis, and patchy fibrosis (Kiss and Fazekas, 1982). Post mortem ultrastructural examination of the heart was also performed on 10 patients who died of acute poisoning by azinphos-ethyl (9) or dimethoate (Pimentel and Carrington da Costa, 1992). Patients died 3-17 days after poisoning, all having shown ECG changes during the illness. Lysis of myofibrils, swollen and fragmented mitocondria, disorganization of nuclear chromatin, and Z band abnormalities were observed. Whether these changes are primarily related to cholinergic overstimulation of the heart and subsequent arrhythmias or are secondary to the general condition of the patients cannot be ascertained. In conclusion, several types of arrhythmias may develop during acute OP poisoning. Therefore, patients must be carefully monitored and promptly treated. 51.1.3.4 Central Nervous System Manifestations Various manifestations of CNS involvement that grossly reflect the severity of poisoning have been described. Thus symptoms and signs range from headache, anxiety, confusion, sleep disturbances, and blurred vision to tremor and convulsions to coma, hypothermia, and central respiratory depression.
1049
Some authors believe the failure of the central respiratory drive is an important factor in causing respiratory failure (Grob, 1956; Tsao et al., 1990), although clinical discrimination of the selective effects of anticholinesterases on the CNS is difficult, particularly in assessing their relative importance in causing death (Lotti, 1992a), because toxicity to the CNS, in general (Norton, 1986), and of anticholinesterases, in particular (Glow and Rose, 1965), always holds a strong peripheral component (see also Section 51.1.3.2). Early CNS signs and symptoms may last for several days, in which only moderate additional cholinergic symptomatology may be detectable (Grob and Harvey, 1958). Electroencephalogram (EEG) abnormalities can be detected at the onset of symptoms and are characterized by irregularities in rhythm, variation and increase in potential, and intermittent bursts of abnormally slow waves of elevated voltage similar to those seen in epilepsy; these symptoms usually persist for about a week or longer (Grob et aI., 1947). Coma is usually due to direct CNS depression by OPs, although hypoxia that derives from respiratory failure may contribute. In this condition, EEGs show profound depression of cortical activity (Lotti and Becker, 1982a). OPs vary in their potency to induce seizures (Hoskins et aI., 1986) and perhaps this manifestation is not entirely due to AChE inhibition (Van Meter et aI., 1978), considering that it is blocked by benzodiazepines, which is known to act via y-aminobutyric acidergic (GABAergic) mechanisms (Lipp, 1973; see also Section 51.1.3.8). Occasionally, during the acute phase of OP intoxication, patients displayed opsoclonus (De Bleecker, 1992; Hata et aI., 1986; Pullicino and Aquilina, 1989). Opsoclonus is an abnormal, rapid, involuntary, repetitive, chaotic, and conjugated ocular movement. Different OPs were involved in reported cases; the onset was from 3 to 48 h and there no was correlation with the typical symptomatology, although all patients were severely poisoned. Opsoclonus is most likely another sign of cholinergic overstimulation, which lasts for several hours to a few days and recovers spontaneously. One case of poisoning with chlorpyrifos and two others with unknown OPs presented with choreo-athetosis (Joubert et al., 1984; Joubert and Joubert, 1988). Extrapyramidal signs lasted a few days during recovery from acute intoxication. Similar signs also have been reported several days after the onset of cholinergic toxicity (see Section 51.1.3.6). 51.1.3.5 Laboratory Findings Red Blood Cell and Plasma Cholinesterases In addition to synapses, AChE is also present in the outer membrane of red blood cells and, to a lesser extent, in plasma. Its physiological functions in blood are unknown. Plasma butyrylcholinesterase (BuChE), also known as pseudocholinesterase, has a different substrate specificity because it hydrolyzes butyrylcholine. Its physiological functions also are not known in plasma or elsewhere. BuChE inhibition by OPs is, therefore, not necessarily indicative of exposures high enough to cause poisoning. Moreover, certain diseases and genetic conditions are characterized
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CHAPTER 51
Clinical Toxicology of Anticholinesterase Agents
Table 51.5 Correlation between Severity of Poisoning, Inhibition of RBC AChE, and Symptoms and Signs of OP Poisoning" Severity of poisoning (approximate activity
Symptoms and signs
ofRBC AChE)
Muscarinic
Nicotinic
Central nervous system
Mild (RBC AChE> 40%)
Nausea, vomiting, diarrhea, salivation/lachrymation, miosis bronchoconstriction, increased bronchial secretions, bradycardia Same as above plus pupils unreactive to light, urinary/fecal incontinence Same as above
Usually none
Headache, dizziness
Muscle fasciculation (fine muscles)
Same as above plus dysarthria, ataxia
Same as above plus muscle fasciculation (diaphragm and respiratory muscles)
Same as above plus coma, convulsions
Moderate (20% < RBCAChE < 40%) Severe (RBC AChE < 20%)
aModified from Lotti (1991).
by low levels of plasma BuChE (see Section 51.4.3.2). Several methods are available to measure these blood enzymes (among the many are Doctor et aI., 1987; Ellman et al., 1961; GarciaLopez and Monteoliva, 1988; Lewis et aI., 1981; London et aI., 1995; St. Omer and Rottinghaus, 1992; Wilson et aI., 1996). However, because hospital laboratories rarely measure RBC AChE activity, in most circumstances one should rely on measurements of plasma BuChE, for which kits are easily available. RBC AChE inhibition confirms the diagnosis of acute OP poisoning. Whole blood AChE also may be measured, considering that only about 10% of the activity is due to the plasma enzyme (Worek et aI., 1999b). Usually there is a good correlation between the severity of signs and symptoms of poisoning and the degree of inhibition of RBC AChE (Table 51.5). Nevertheless, because acute poisoning usually requires prompt treatment, treatment should not be delayed while laboratory confirmation is sought. Therefore, measurement of RBC AChE has limited value in an emergency because diagnosis is exclusively clinical and severe poisoning is inevitably associated with high RBC AChE inhibition. It is more difficult to interpret the relatively low levels of RBC and plasma cholinesterases such as those observed in cases of poisoning that present with equivocal symptoms. This may be the case in mild poisoning or in the initial phase of poisoning caused by OPs that are slowly disposed. Reasons for these difficulties are manyfold and include the following: • Large inter- and intraindividual variability of both RBC AChE and plasma BuChE, making the distinction between physiologically low and inhibited activities impossible. Methods to address these difficulties have recently been developed but they have not been applied yet in clinical settings (Polhuijs et aI., 1997; see Section 51.4.3.2). • Different sensitivity of AChE and BuChE to the same inhibitor. In many cases, the identity of the chemical involved and knowledge of its biochemical characteristics are unknown, thus hampering the interpretation of a significant inhibition of plasma BuChE associated with little
or no inhibition of RBC AChE. This may be the case in both a mild poisoning or the early phase of a more severe poisoning when poisonings are caused by OPs that preferentially inhibit BuChE (see Section 51.4.3.2). • The ratio between the inhibition of RBC AChE and that in the synapses may vary according to the compound. Inhibition of the RBC enzyme may be detected without clinical signs of toxicity in cases of exposures to OPs that do not easily cross the blood-brain barrier (see Section 51.4.3.2). • If reactivators (oximes) are administered, the pharmacological effect depends on the ratio between inhibited and aged enzyme. This ratio may be different in blood and in the nervous tissue, and the pharmacological reactivation of inhibited blood enzymes will be more effective than that of enzymes in the nervous system. Under these circumstances relatively small inhibition in blood enzymes would not correlate with symptomatology (see Section 51.4.3.2). • The method of cholinesterase measurement involves dilutions, which in the case of inhibitors such as carbamates and dimethyl phosphates, may favor the spontanous reactivation of inhibited enzymes (see Section 51.1.2.2). These assay related problems would understimate the actual in vivo inhibition. In conclusion, measurements of blood cholinesterases may confirm the diagnosis, but are not essential: clinical observation remains the cornerstone for diagnosis. For the same reasons, repeated measurements of blood cholinesterase during poisoning have no prognostic value (Nouira et aI., 1994) and they cannot be used to assess the efficacy of treatment. When AChE is irreversibly inhibited by OPs, the reappearance of RBC activity depends on new erythrocytes entering the blood stream. Whereas the average lifespan of RBCs is 120 days, in most cases, observed reappearance of RBC AChE occurs at a rate of about 1% per day. The corresponding rate for plasma BuChE, which derives from liver synthesis, is about 5% per day (see Section 51.4.3.2).
51.1 The Cholinergic Syndrome
1051
100
10 0;
El
SCo
~
0
I>Il
::!.
0.1
O.oI 0
2
4
6
8
10
DAYS
Figure 51.4
Time course of blood concentration of several OPs in acutely poisoned patients. Data for
(+) methamidophos, (_) fenitrothion, (J.) methylparathion, and (e) parathion from (Lotti, 2000); (<» chlorpyrifos from (Lotti et al., 1986); (0) methidation, (L'.) dimethoate, and (x) mecarbam from (Tsatsakis et aI., 1996); (0) malathion from (Lyon et al., 1987); (*) trichlorfon from (Nordgren et aI., 1980).
Measurements of OPs and Metabolites in the Blood and Urine Several analytical methods based on chromatographic techniques are available for quantitative and qualitative measurements of OPs and their metabolites in body fluids (Tomlin, 1997). These measurements have confirmatory uses in clinical toxicology although they are rarely performed because they are not easily available. Nevertheless, serial measurements of the parent compound in blood (Fig. 51.4) identifies the compound(s) and provides information about its pharmacokinetics in humans. Metabolites of OPs can be measured in the urine. They are products of hydrolytic reactions, and include the alkylphosphate and the alcoholic moieties. Because the kinetics of urinary excretion of metabolites differ, quantitative extrapolations are almost impossible unless complete urine collection is performed over several days and both metabolites are measured. Moreover, if only one of these metabolites is measured, the results may be less specific because different OPs can share identical acidic or alcoholic moieties. Thus, two parent OP compounds that display quite different acute toxicity may produce the same amount of the same metabolite, thereby further hampering quantitative extrapolations (see Section 51.4.3.1). In conclusion, measurements of OPs and their metabolites have a limited value for diagnosis of acute OP poisoning. They may be useful for confirmatory purposes and research. Routine Hematological and Biochemical Tests Abnormalities of almost all common laboratory tests have been reported during the course of acute OP poisoning. None of them is specific and they may reflect either somewhat typical short-term complications (pancreatitis and myopathy; see Section 51.1.3.6) or the general clinical conditions of the patient (degree of respiratory failure, changes in organ perfusion, concurrent infections, iatrogenic consequences, etc.). Isolated reports indicate that activation of blood coagulation was ob-
served in cases of parathion and dimethoate poisoning, and in one case the patient was treated with eparine (Jastrzebski et aI., 1994; Kaulla and Holmes, 1961). However, when studies were performed on a series of 31 patients moderately poisoned with either parathion or sarin, both hyper- and hypocoagulability were observed (Kaulla and Holmes, 1961). The clinical and toxicological relevance of these isolated observations remains unclear. In conclusion, changes in common laboratory tests are not specific for the diagnosis of OP poisoning. Electrophysiology The electrophysiological consequences of AChE inhibition at the neuromuscular junction have been reviewed (Singh et al., 1998b), and are characterized by single electrical stimulus-induced repetitive compound muscle action potentials (CMAPs) and, in response to repetitive nerve stimulation, by a decrement of the CMAP (Besser et aI., 1989a, b; Maselli et al., 1986). These events reflect the excess of acetylcholine at synaptic levels and subsequent alteration of nicotinic receptor responses. The repetitive muscle response to a single nerve stimulus is thought to result from reexcitation of the muscle by the prolonged endplate potentials brought about by the excess of acetylcholine. At relatively low AChE inhibition, repetitive stimulation causes a decrement followed by complete recuperation of CMAP amplitudes. However, at higher AChE inhibition, repetitive stimulation results in an unimodal pattern of progressive decrements of CMAP amplitudes. Experimental studies demonstrated that in the former case, impaired neuromuscular transmission is caused by transient depolarization of the endplate region, whereas in the latter, direct blockade is due to desensitization of postsynaptic nicotinic receptors (Maselli and Leung, 1993; Maselli and Soliven, 1991). Decremental response to repetitive nerve stimulation dramatically worsened after injection of edrophonium (MaseUi et aI., 1986), indicating the effects of a further excess of acetylcholine, whereas the administration of pancuronium improved
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CHAPTER 51
Clinical Toxicology of Anticholinesterase Agents
neuromuscular transmission, indicating a blockade of nicotinic receptors (Besser et aI., 1990, 1991). Distinction between these findings and those detected during the intermediate syndrome may be difficult (see Section 5l.2.3.2). Other Tests On a small series of patients poisoned with OPs, cerebral perfusion was investigated by brain single photon emission computerized tomography (Yilmazalar and bzyurt, 1997). The authors concluded that in severe cases of poisoning, patients showed perfusion defects, especially in the parietal lobe. This study is difficult to assess because some patients who had normal plasma cholinesterase and some mild symptoms probably were not poisoned. Moreover, the patients who had major perfusion deficits were the oldest and almost no improvement was detected 3 months after poisoning. Pathology Gross pathology and histological examination performed at autopsy after fatal poisoning were either unremarkable or nonspecific (Maresch, 1957). In particular, given the mechanism by which OPs cause death, neuropathology is not observed unless severe hypoxia or convulsions occurred (McLeod, 1985). A common nonspecific feature of CNS histopathology is vascular damage associated with increased permeability of the vessel walls, suggesting major changes in the blood-brain barrier that might occur during OP poisoning (see Section 5l.l.3.8). 51.1.3.6 Short-Term Complications Neurological Fatal encephalopathy was reported in two cases of acute OP poisoning (de Reuck et aI., 1979). These patients, after an initial recovery from cholinergic toxicity, developed a severe encephalopathy within 3-4 days and died 9-20 days after admission. Necroscopy showed severe hemorrhagic necrosis of the ventricles, resembling lesions observed in Wernicke's encephalopathy. There is no explanation for such findings, although it was reported that one patient had a history of alcoholism and the other of recurrent depression. Neuroleptic malignant-like syndromes following organophosphate poisoning have been reported in the Japanese literature (Ochi et aI., 1995). In one case, a 60-year-old schizophrenic patient who had undergone a frontal lobotomy at the age of 20, but was free of medications since then, attempted suicide by ingesting a large amount of methidathion. About a week after full recovery from symptoms and signs of OP poisoning, the patient developed a high fever, extrapyramidal rigidity, and coma. Serum CPK and LDH were increased, and high urinary myoglobin was detected, whereas plasma cholinesterase was back to normal. The patient was treated successfully with dandrolene. A correlation with OP poisoning is difficult to ascertain because this clinical picture is also consistent with lethal catatonia, a rare and potentially lethal syndrome that occurs in schizophrenic patients (Castillo et aI., 1989). Extrapyramidal manifestations that complicate OP poisoning were described in six patients poisoned with fenthion (Senanayake and Sanmuganathan, 1995). Patients had moderate to severe cholinergic signs and all developed intermediate
syndrome subsequently. Dystonia was observed in all patients, and the most common signs were tremor, choreo-athetosis, and cog-wheel rigidity. Onset of these extrapyramidal signs was variable (4-40 days) and spontaneous recovery was observed within 4 weeks. Although the development of extrapyramidal signs is rarely described in OP poisoned patients (and in these cases, it may be compound specific), a possible relationship with inadequate oxime therapy has been inferred. Transient bilateral vocal cord paralysis has been reported after OP poisoning, although the reports are inconsistent and describe different clinical characteristics. A 3-year-old boy severely poisoned with chlorpyrifos was intubated and artificially ventilated (Aiuto et aI., 1993). On day 3, immediately after extubation, the boy developed stridor and was reintubated. He was reextubated on day 6 and specific treatment for OP poisoning was halted. Occasional stridor was treated conservatively until day 11, when he needed intubation again. Tracheostomy was performed on day 19 and maintained for more than 50 days. Generalized areflexia was observed on day 18 that rapidly recovered within a week. Follow-up was not reported. Although the authors suggested that organophosphate-induced delayed polyneuropathy (OPIDP) developed, the distribution of weakness and the time course of events is not characteristic (see Section 51.3.3.1). Three cases of delayed recurrent laryngeal nerve paralysis (25-35 days from the onset poisoning) were observed in subjects severely poisoned with three different OPs (chlorpyrifos, methamidophos, and parathion) who required intubation and artificial ventilation (de Silva et aI., 1994). At the onset of stridor, no other nerve was involved and the patients had been extubated 14-26 days earlier. One patient required tracheostomy, whereas the others were treated conservatively. Recovery occurred within 4-15 weeks. Although OPIDP was suggested as the cause of laryngeal nerve paralysis, the isolated involvement of this nerve speaks against it. Moreover, parathion is not known to cause OPIDP (see Section 5l.3.1). Another case of isolated bilateral vocal cord paralysis was associated with OP exposure (Thompson and Stocks, 1997). However, exposure was anecdotal and there was no evidence of OP poisoning. In further case of poisoning with unspecified OPs, bilateral vocal fold palsy was reported (Indudharan et aI., 1998). Palsy was observed at extubation, about 10 days after poisoning, and required tracheostomy that was maintained for 2 months. The authors suggested intermediate syndrome, but it was unknown when, after poisoning, this palsy occurred (because the patient was intubated) and if other signs of proximal weakness were detected (see Section 5l.2.3.1). Cortical visual loss was observed in two patients poisoned with OPs (Wang et aI., 1999). One patient had cortical visual loss as shown with positron emission tomography (PET) scan 1 year after poisoning with EPN. During this period, blurred vision was the main symptom and was not clear how long she was hypoxic. Another patient had similar effects that were detected 2 months after acute poisoning with mevinphos. She was reported to have had apnea, but the duration was unspecified. Therefore, it is likely that these two cases represent conse-
51.1 The Cholinergic Syndrome quences of secondary hypoxia, although selectivity for this cortical area is unusual. A case of moderate poisoning with a mixture of dimethoate, diazinon, and methoxychlor was reported where the patient was markedly hypothermic 1 h after ingestion in addition to having miosis and diffuse fasciculations. She had a rectal temperature of 33°C, which normalized within 1 h after passive rewarming (Hantson et aI., 1996). In conclusion, there is no strong evidence for any neurological short-term complication directly linked to OP toxicity, except, perhaps, hypothermia, which has been consistently reported in animal studies as a compound-specific effect of certain OPs (Coudray-Lucas et aI., 1983). Pancreatitis Although it is uncommon, pancreatitis has been consistently associated with acute OP poisoning. Nevertheless, because the diagnosis of pancreatitis is usually established by detection of an increased level of serum amylase, quite often the roles of salivary gland overstimulation and/or of acidosis, both present in OP poisoning, have been overlooked as a cause of increased serum amylase. Additionally, in most reports there is no information on possible alternative causes of pancreatitis, including anoxia, hypoperfusion, infections, and concurrent treatment. The latter is important for a differential diagnosis of toxic pancreatitis, because about 5% of causes of acute pancreatitis are related to commonly used drugs (Greenberger et aI., 1998). Transient hyperglycemia and glycosuria, which are often found in severe OP poisoning (Namba et al., 1971; Zadik et aI., 1983), are also common in acute pancreatitis. Hypotheses have been formulated that link the pathophysiology of the triad of pancreatitis, hyperamylasemia, and hyperglycemia in acute and severe OP poisoning (Haubenstock et aI., 1983). The first case was reported on a patient who was anoxic for an unspecified period, was misdiagnosed as acute kerosene inhalation, and was later found to be intoxicated with OPs. High serum amylase levels were almost immediately detected and after 9 days, a pancreatic pseudocyst was drained (Dressel et aI., 1979). A case of OP poisoning by coumaphos was reported where the patient with severe respiratory failure had elevated serum amylase on admission that returned to normal 20 h later (Moore and James, 1981). Another case of pancreatitis was less convincingly attributed to cutaneous exposure to an organophosphate insecticide. Symptomatology was mild though consistent with OP poisoning but there was no evidence of poisoning and medical history suggested a very mild exposure, if any. Symptoms of pancreatitis persisted for 6 months (Marsh et al., 1988). Acute painless pancreatitis developed in a patient soon after severe intoxication with mevinphos (Hsiao et aI., 1996). RBC AChE was inhibited, serum amylase and lipase were elevated, and a computed axial tomography (CT) scan indicated pancreatitis. Persistent hyperglycemia developed. Two cases of acute severe pancreatitis, complicated by pancreatic necrosis and retroperitoneal sepsis were described in patients with severe poisoning by unspecified OPs (Panieri et aI., 1997). In both
1053
cases, diagnosis of pancreatitis was made several weeks after admission (2 and 5, respectively) and confirmed at surgery. On a series of 75 patients admitted to the hospital for malathion poisoning, serum amylase was serially measured (Dagli and Shaikh, 1983): 47 patients had a mildly raised amylase that reversed within 2 days. Mild symptoms compatible with OP poisoning were reported in most patients, but it is not clear if they were the same patients with elevated amylase. Apparently none had severe poisoning and no toxicological evidence of OP poisoning was provided. In another series of nine patients poisoned with parathion, painless acute hemorrhagic pancreatitis was manifested by ileus in two cases (Lankisch et aI., 1990). Patients were severely ill, requiring artificial ventilation, and had depressed plasma cholinesterase levels. One subject developed persistent ileus 1 week after admission and hemorrhagic pancreatitis was observed at surgery. The other patient presented with severe shock and paralytic ileus. Blood lipase and amylase were elevated and the CT scan indicated pancreatitis. A series of 17 children poisoned with OPs was compared with a matched control group with similar abdominal symptoms (Weizman and Sofer, 1992). Compounds were identified in nine cases and included parathion, malathion, and diazinon. Five poisoned children were diagnosed with acute pancreatitis based on elevated serum amylase trypsin and glucose. Ultrasonography was not performed. None of the controls developed pancreatitis. It is not clear when pancreatitis developed after exposure, if adominal pain in these patients was, due to either pancreatitis or cholinergic overstimulation, how severe the poisoning was and what disease was diagnosed in controls. In a retrospective study on 32 cases of OP poisoning (unspecified chemicals), 16 patients developed respiratory failure (Matsumiya et aI., 1996). The average levels of plasma amylase of these patients on admission was higher than that of patients who did not developed respiratory failure. The authors concluded that high amylase was predictive of subsequent respiratory failure. However, several patients with respiratory failure had normal plasma amylase and several patients without respiratory failure had elevated plasma amylase. In a retrospective study on 159 patients admitted to the hospital for OP poisoning, Lee et al. (1998) found hyperamylasemia in 44 patients (out of 121 for which data were available) and hyperlypasemia in 9 (out of 28). Nine patients were diagnosed as having pancreatitis and two of them died. The incidence ofhyperamylasemia was related to the clinical severity of poisoning. It is difficult to conclude from the described clinical data whether OPs are directly toxic to the pancreas. Different OPs have been involved in these cases, the onset of pancreatitis seems to be quite variable after the beginning of cholinergic symptomatology, and important details are missing in many reports. Moreover, there are indications that in most circumstances, shock and subsequent hypoperfusion preceded the development of pancreatitis. In conclusion, although pancreatitis is consistently reported, there is no strong evidence that it is a characteristic feature of acute OP toxicity.
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CHAPTER 51
Clinical Toxicology of Anticholinesterase Agents
Myopathy On several occasions, junctional myopathy, that is, muscle damage that originates in the endplate region, has been observed in poisoned patients. The first description of muscle fiber necrosis was obtained from a case of severe parathion poisoning (de Reuck and Willems, 1975). The patient was admitted with severe respiratory failure and was artificially ventilated. Myoclonic jerks and fasciculations were present and blood cholinesterases were profoundly depressed. Episodes of irregular tachycardia developed on day 4 and the patient died of cardiac arrest on day 9. Post mortem examination revealed patchy and focal areas of necrosis of muscle fibers in the diaphragm. Nerve endings in the segmental necrotic zones of the muscle fibers were degenerated, whereas normal motor endplates were observed in the nonnecrotic muscle fibers. Similar lesions involving muscles other than the diaphragm were reported after acute poisoning with trichlornate (de Reuck et aI., 1979). The patient had moderate signs of acute OP poisoning including muscle fasciculations. After an initial improvement, the patient became comatose and required artificial ventilation on day 6, and eventually died 20 days after admission because of Gram-negative sepsis. Post mortem examination revealed waxy degeneration and lysis of individual fibers in the quadriceps femori muscle and, to a lesser degree, in the deltoid, diaphragm, and intercostal muscles. Nerve fibers were normal. In another study, samples of intercostal muscles were obtained at autopsy from a subject who died after an unquantified exposure to malathion and diazinon. Toxicological evidence of exposure and information on the severity of cholinergic overstimulation were scanty. Moreover, the patient died of brain hemorrhage. Histopathology of muscles showed basophilic inclusions and scattered necrotic fibers. It is difficult to attribute these changes to OP exposures and even more difficult to attributed them to prolonged AChE inhibition. In fact, muscular AChE was inhibited to about 50% and plasma cholinesterases were inhibited perhaps a little more (Wecker et aI., 1985). Another case report associated myopathy with OP exposure, although it was much less convincing (Ahlgren et aI., 1979). The patient, who was working as an exterminator, had never had acute poisoning. Sometimes he exhibited symptoms compatible with cholinergic overstimulation, but they rapidly subsided. He presented with a 2-year history of muscle weakness, involving mainly the trunk, the shoulders and the pelvic girdle. Several muscle biopsies performed during the course of the disease revealed necrotic fibers, inflammatory reactions, and fibrotic changes in the deltoid muscle and quadriceps femoris muscles. Five years after the onset of symptoms, this patient died of progressive respiratory failure. At autopsy generalized atrophy and fibrosis of striated muscles was observed; and in particular, the diaphragm was almost entirely fibrotic. It is difficult to judge whether chronic exposure to OP caused the progressive myopathy. Clinical expression of any toxic disease is usually complete after cessation of exposure, and this patient reduced and eventually halted the exposure because of the clinical condition. No other reports indicate the progression of acute myopathy beyond the initial period of cholinergic overstimulation. There is
no quantitative evidence of exposure, which was presumably low and comprised several pesticides, although the patient reported that he most frequently sprayed diazinon. In conclusion, there is limited evidence that acute OP poisoning causes myopathy in humans and it may be the consequence, in the cases described, of either fasciculations or sepsis. Muscle necrosis, the human equivalent of a well known effect observed in experimental animals (Dettbarn, 1992), was thought to be a possible cause of intermediate syndrome (Karalliedde and Henry, 1993; Senanayake and Karalliedde, 1987; see Section 51.2.2). However, although clinical signs suggestive of red muscle damage (such as elevated CPK) have been observed in some patients who displayed signs of intermediate syndrome (He et aI., 1998), muscle biopsies suggest that the lesions were too sparse to justify clinically detectable muscle weakness (De Bleecker, 1993; De Bleecker et aI., 1993). Others Severe intoxications carry obvious risks of several secondary manifestations. Examples include a case of diazinon poisoning characterized by loss of fluids complicated with acute renal failure (Abend et aI., 1994), a case of acute dimethoate poisoning complicated with Gram-negative pneumonia, acute respiratory distress syndrome, and acute renal failure (Betrosian et aI., 1995), and two fatal cases of poisoning with diazinon presenting with severe hyperglycemia, metabolic acidosis, and hypoalkalemia (Hui, 1983). Preexisting diseases may also be worsened by OP poisoning as described in cases of cardiovascular disorders, which have precipitated cerebral infarction and gangrene (Buckley et aI., 1994). All these and other complications should be expected in any severe case of poisoning treated in an intensive care unit. The low levels of plasma cholinesterase caused by OP exposure change the pharmacokinetics of drugs that are substrates for these enzymes. Thus, neuromuscular blocking agents, which are widely used in anesthesia, will be slowly metabolized, thereby prolonging their pharmacological action (0stergaard et aI., 1992). The first case was described in a patient severely poisoned with parathion who had undetectable plasma cholinesterase activity. In an effort to control the patient's convulsions, small doses of succinylcholine were administered (Quinby et aI., 1963). The patient became suddenly apneic and completely flaccid, and the marked neuromuscular block lasted for about 2 h. A patient suspected of having a partial obstruction of the small intestine underwent emergency laparotomy and the anesthesia was induced with thiopentone and suxamethonium (Gesztes, 1966). At inspection, no obstruction was observed and the patient had a prolonged apnea when treatment was discontinued. It was later recognized that the patient had been treated with eye drops that contained ecothiopate iodide-an anticholinesterase drug-that was systemically absorbed. This clinical case is fully justified by the toxicity of ecothiopate. Thus, the pseudo-obstruction associated with the diarrhea was likely a sign of cholinergic overstimulation by ecothiopate and the prolonged apnea was the consequence of plasma cholinesterase inhibition, eventually leading to reduced metabolism of the suxamethonium.
51.1 The Cholinergic Syndrome A similar case was reported in a girl severely poisoned with chlorpyrifos following administration of succinylcholine for airway management (Selden and Curry, 1987). A patient exposed to malathion underwent surgery for acute appendicitis. Because he had no signs of cholinergic overstimulation, suxamethonium was given to facilitate tracheal intubation (Guillerrno et aI., 1988). He suffered prolonged apnea due to low plasma cholinesterase activity, which most likely was due to relatively low exposure to malathion because he had a normal plasma cholinesterase phenotype. Another case of a boy who suffered relatively mild exposure to chlorpyrifos and propetamphos, and exhibited no symptoms of cholinergic overstimulation, showed prolonged apnea following suxamethonium treatment (Weeks and Ford, 1989). These cases represent one of several possible conditions in which low plasma cholinesterase activity prolongs the pharmacological effects of neuromuscular blocking agents (Davies and Landy, 1998; Hart et aI., 1995; Kopman et aI., 1978; see Sections 51.4.3.2 and 51.4.3.3). OP intoxication seems to have no effect on pregnancy if patients are properly treated. Pregnant women, between 9 and 36 weeks of gestation, intoxicated with OPs (sarin, methamidophos and fenthion) who received appropriate management completely recovered from the poisoning, allowing the pregnancies to continue to term unaffected and the delivery of healthy babies (Karalliedde et al., 1988; Ohbu et aI., 1997). 51.1.3.7 Differential Diagnosis Needless to say, it is vital to distinguish OP poisoning from other diseases, particularly when the differential diagnosis may be hampered by unusual scenarios. Thus, when observing a severely ill patient, an exposure to OPs may not be suspected at first glance (Bjornsd6ttir and Smith, 1999). Conversely, patients known to have been exposed to OPs may present with illnesses that are unrelated to exposure. Although diagnosis of acute OP poisoning is straightforward, difficulties have been reported. In 20 children transferred from other hospitals, correct diagnosis of OP poisoning was made in only 4, whereas the others had diagnoses of pneumonia, various infectious diseases, encephalopathy and head trauma (Zwiener and Ginsburg, 1988). Conversely, in another series of 78 patients admitted to the hospital with diagnoses of pesticide poisoning (among which 34 were thought to be caused by OPs), only 36 (among which 18 were due to OPs) were confirmed as clinical poisoning (Lamminpaa and Riihimaki, 1992). Reasons for misdiagnosis included other illnesses, no evidence of exposure, or proven limited absorption. Although the signs of cholinergic overstimulation are characteristic, they must be actively sought when OP poisoning is suspected, particularly in mild cases, because they are rarely isolated. Detection of concurrent signs would help, for instance, to ascertain the toxic nature of respiratory failure, which should be distinguished from other respiratory and circulatory causes and, similarly, for differential diagnosis of the most serious CNS manifestation such as seizures and coma (Greenaway and Orr, 1996; Hollis, 1999).
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51.1.3.8 Treatment Prompt treatment of OP poisoning is lifesaving. Several procedures are available, the sequence of which depends solely on the severity of poisoning. Special attention should be exercised by medical personnel caring for these patients, because passive contamination may occur (Ohbu et aI., 1997). Minimizing Further Exposure Procedures aimed at decontamination and/or at minimizing absorption depend on the route of exposure. Thus, in the case of dermal exposure, contaminated clothing should be removed and the skin should be washed with alkaline soap. When the skin appears to be clear, the patient should be bathed or swabbed, because most OPs are more soluble in alcohol than in water (Durham and Hayes, 1962). In the case of eye contamination, extensive irrigation with water or saline should be performed for several minutes. In case of ingestion, various procedures have been recommended to reduce absorption from the gastrointestinal tract, although there is limited evidence of their efficacy (John son and Vale, 1992; Lotti, 1991). In conscious patients, vomiting is usually induced with syrup of ipecacuanha (10-30 ml followed by 2-300 ml of water). This treatment is contraindicated if the patient is semiconscious, has difficulty swallowing, or if the pesticide is dissolved in hydrocarbon solvents or is corrosive, given the high incidence of pneumonitis/atelectasias due to insecticides that contain petroleum distillate (Zwiener and Ginsburg, 1988). In these circumstances, the increased probability of aspiration pneumonia largely exceeds that of potential benefits. Gastric lavage with instillation of activated charcoal should be performed after the patient's airway has been protected with an endotracheal tube. It has been suggested that this procedure be repeated every few hours, as long as the chemical is detectable in the lavage fluid, because it may persist for several days (Futagami et aI., 1995; Willems, 1981). Although the value of this continuous lavage is not proven, empirical evidence and the slow absorption characteristic of some OPs (see Section 51.1.3.1) suggest this option be considered. Moreover, complications due to administration of activated charcoal, in addition to aspiration, may also include intestinal ulceration and massive bleeding due to constipation (Mizutani et aI., 1991). In fact, when these patients are treated with atropine concurrently, they display decreased bowel peristalsis. Thus, although the use of vigorous cathartic treatment has been suggested as an adjunct to charcoal to hasten the elimination of the charcoal-poison complex, studies in humans have failed to demonstrate any substantial benefit from this combination (Neuvonen and Olkkola, 1988). Other procedures aimed at removing the absorbed OP, for example, hemoperfusion or hemodialysis, have been suggested (Luzhnikov et aI., 1977; Verpooten and De Broe, 1984; Yokoyama et aI., 1995), but it is doubtful that these methods achieve higher clearance of OPs from the blood than that which occurs spontaneously (Nagler et aI., 1981). Moreover, because of the high reactivity and large volume of distribution of OPs, even a high blood clearance will not significantly reduce the total amount of the compound. In a retrospective study
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CHAPTER 51
Clinical Toxicology of Anticholinesterase Agents
on the effects of hemoperfusion, it was shown that this procedure did not remove significant amounts of various insecticides (fenitrothion, methylparathion, fenthion, chlorpyrifos, trichlorphon, diazinon, omethoate and fenamiphos) from the body of 10 patients (Martinez-Chuecos et aI., 1992). High plasma concentration of OPs reappeared after the last hemoperfusion, symptoms and signs did not ameliorate, and prolonged clinical course and complications were not avoided. Similar results were reported in a fatal case of fenitrothion poisoning treated with combined hemoperfusion and hemodialysis (Yoshida et aI., 1987). Atropine Atropine represents the cornerstone of the treatment for poisoning by anticholinesterase. Other available drugs, such as reactivators of inhibited AChE, which are very effective, should be considered, in principle, as valid adjuncts to atropine administration. Atropine is a muscarinic receptor antagonist that prevents the effects of acetylcholine by blocking its binding to muscarinic cholinergic receptors (Brown and Taylor, 1996). It is a racemic mixture of active l-hysocyamine and inactive dhysocyamine that should be stored at 15-30°C and protected from light. Freezing should be avoided. The shelf life is 24 months from the date of manufacture if it is kept under the recommended conditions (Heath and Meredith, 1992). Pharmacokinetic data about atropine are limited. The kinetics of distribution of atropine seem to be dose dependent: about 20% of the drug is bound in plasma and two phases, with apparent half-lives of 1 and 140 min, respectively, have been identified after intravenous injection (Hinderling et aI., 1985). However, for practical purposes, the reported plasma half-lives after both intravenous and intramuscular injections, varying between 1.3 and 4.3 h, should be considered. Differences are due to assay methods and to a considerable intraand interindividual variability (Adams et aI., 1982; Kanto and Klotz, 1988; Kenta1a et aI., 1990). In children and in the elderly, the plasma half-life may be longer. The reported apparent volume of distribution is quite large (2-3.5 literkg- l ), implying significant intra- and extracellular binding and partitioning of the drug. In children, higher volumes of distribution than in adults have been reported. About 50% of atropine is eliminated unchanged in the urine. There is no correlation between plasma levels and maximal pharmacological effects after intravenous injections (Adams et aI., 1982); therefore, the dose of atropine cannot be titrated by means of plasma concentration. For practical purposes, however, one should consider that the effects of intravenous atropine begin within 3-4 min and are maximal about 12-16 min after injection. Atropine is less effective in blocking certain muscarinic effects (for example, effects on the gastrointestinal and urinary tracts) than others effects (for example, effects on the heart and the salivary glands). Atropine has no effect on nicotinic symptoms, and central muscarinic effects may be undetectable, perhaps reflecting the difficulty of penetration of atropine into the CNS, which can be achieved only by large doses (Taylor, 1996a).
Caution should be exercised in the use of atropine in hypoxic patients because it may cause ventricular fibrillation due to the increased myocardial oxygen demand brought about by the increased heart rate produced by atropine (Massumi et aI., 1972). Therefore, in severe cases of OP poisoning, anoxia should be corrected before atropine is administered (Durham and Hayes, 1962). However, when arterial oxygen has been normalized, there is no reason to avoid the use of atropine because of the suggested risk of ventricular fibrillation (Kecik et aI., 1993). Atropine is preferably given intravenously, although the intramuscular route is also effective. Satisfactory absorption also can be achieved by inhalation, and the pharmacokinetic characteristics are similar to those that occur after intramuscular injection (Harrison et aI., 1986). However, atropine administration by inhalation has not been tested in cases of OP poisoning. Although several dosage regimens have been proposed and some caution was suggested in the dosage of atropine (de Kort et aI., 1988), the best clinical approach is to administer doses of atropine large enough to achieve clinical evidence of atropinization, that is, flushing, dry mouth, changes in pupil size, bronchodilation and increased heart rate. If such signs are undetected, the dose of atropine is assumed to be insufficient and it must be increased (Barr, 1966). A mild degree of atropinization should be maintained for at least 48 hand withdrawal of atropine should be very carefully monitored because relapse can occur, particularly when OPs are stored in fat (see Section 51.1.3.1). In case of relapse, atropinization should be immediately reestablished. In patients with mild cholinergic signs, it is appropriate to start with a test dose of atropine (1 mg in the adult and 0.01 mg kg- l in children, intravenously). If signs of atropinization occur rapidly, severe poisoning is unlikely, although observation of the patient for at least 24 h is mandatory. In moderately to severely poisoned adult patients, 2-5 mg of atropine should be given intravenously and repeated every 10-20 min (0.02-0.05 mg kg- l in children at the same intervals). Continuous intravenous infusion may be required in severe cases. Because patients poisoned with OPs are tolerant to the effects of atropine, quite large doses of the drug have been used in cases of severe and prolonged poisoning (Golsousidis and Kokkas, 1985; Lotti et al., 1986). In a case of dimethoate poisoning, 30 g of atropine were given over 35 days with maximum daily dosage of 3.5 g (Le Blanc et aI., 1986). Indicated dosages of atropine in accord with the severity of the clinical picture are summarized in Table 51.6. Overdosage with Table 51.6 Indicative Dosage of Atropine in OP Poisoning According to Severitya Adults
Children
Poisoning
(mg)
(mgkg- l i.v.)
Mild Moderate Severe
1.0 2.0-5.0 20h- 1
0.01 0.02-0.05 0.2h- 1
(infusion)
(infusion)
a See
Table 51.5 for grading of poisoning severity.
51.1 The Cholinergic Syndrome
atropine is rarely serious in OP poisoned patients. On the contrary, patients frequently die because of insufficient atropine. When massive tachycardia is produced by atropine, it may be corrected by propanolol (Valero and Golan, 1967), thus avoiding the need to reduce the amount of atropine. It has also been suggested that using a combination of atropine and glycopyrrolate might offer an advantage over atropine alone, inasmuch as tachycardia could be avoided and adequate antimuscarinic effects still could be provided (Tracey and Gallagher, 1990), and that glucopyrrolate may better alleviate some signs of cholinergic overstimulation (Choi et aI., 1998). Oximes Oximes are nucleophilic chemicals that remove the phosphoryl group from the inhibited enzyme, thus restoring the catalytic site of AChE and its function (Bismuth et aI., 1992). However, this chemical reaction occurs only when the phosphorylated AChE has not undergone the intramolecular rearrangement known as aging (Fig. 51.3 and Table 51.3; Aldridge and Reiner, 1972; Holmstedt, 1959; Taylor, 1996a). Whereas this reaction is fast (usually within a few minutes), oximes should be available at the synaptic cleft as long as there is newly inhibited AChE. Therefore, oximes should be administered over the first several hours after poisoning, although treatment may be prolonged in cases of massive poisoning and in poisonings by OPs with slow pharmacokinetics. Several oximes have been synthesized and are available (Bismuth et aI., 1992; Dawson, 1994): pralidoxime (2-pyridine aldoxime or 2-PAM) is the most commonly used. Oximes currently in use are pralidoxime chloride (Protopam) in the United States, pralidoxime methylsulfate (Contrathion) in France and Italy, obidoxime chloride (Toxogonin) in Germany and Sweden, and pralidoxime methanesulfonate (P2S) in the United Kingdom. The availability of these oximes, including others such pralidoxime iodide, varies according to national pharmacopeas. Other oximes have been designed for the treatment of nerve gases, but they are not available for civilian uses (Kusic et aI., 1985). The pharmacokinetics of oximes has been studied mostly in normal volunteers and a few studies have compared these results with those observed in poisoned patients. In studies that involved 15 volunteers given pralidoxime (single intravenous doses 2.5, 5.0, 7.5, and 10.0 mg), the plasma half-life was about 1.3 h and the apparent volume of distribution was 0.8 liter kg- l (Sidell and Groff, 1971; Sidell et aI., 1972). Pharmacokinetic parameters have been derived after intravenous injection of obidoxime in five healthy volunteers. The half-life (mean ± SD) was 1.2 ± 0.16 h and the volume of distribution (steady state) was 0.17 literkg- l (Sidell et aI., 1972). Pharmacokinetic data are also available for P2S (Holland and Parkes, 1976; Sundwall, 1960, 1961). A short infusion regimen of pralidoxime chloride was compared with administration of a loading dose followed by continuous infusion in healthy volunteers (Medicis et al., 1996). Plasma levels above 4 mg liter- l were maintained with the latter regime for twice as long as with the former (257 vs. 118 min). The pharmacokinetics of pralidoxime chloride was
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compared in nine healthy volunteers and six severely poisoned patients (Jovanovic, 1989). Compounds involved in the poisonings were malathion, quinalfos, and dimethoate, and pharmacokinetic data were obtained after a single intramuscular dose of 1 g 2-PAM. The pharmacokinetic parameters (means ± SD) in volunteers were different from those previously reported in the literature: plasma half-lives were 148 ± 65 min, volume of distribution was 2.7 literkg- l , and total body clearance was 9±4 mlmin- l kg-I. Nevertheless, when these parameters were compared with those in poisoned patients, significant differences were found. Thus, mean plasma concentrations of 2-PAM were one and a half times higher in patients than in volunteers at each time point and remained above 4 mg liter- l for 239 and 137 min, respectively. Elimination of oxime was greatly reduced in poisoned patients as compared with controls. Different pharmacokinetics of pralidoxime chloride also were reported in a series of children poisoned to a different degrees of severity with parathion, dichlorvos, diazinon, chlorfenvinfos, dicrotofos, and other OPs (Schexnayder et aI., 1998). Continuous intravenous infusion of 10-20 mg kg- l h- l pralidoxime following a loading dose of 15-50 mg kg- l gave the following results: mean (±SD) steady state plasma concentration of pralidoxime was 22.2 mg litec l ± 12.3 (range 6.9-47.4 mgliter- l ), mean (±SD) plasma half-life was 3.6 ± 0.8 h (range 2.4-5.3 h), mean clearance (±SD) was 0.88 ± 0.55 literkg- l h- l (range 0.28-2.2 literkg- l h- l ), and mean volume of distribution (±SD) was 5.5±4liter kg- l (range 1.713.8 liter kg-I). The large clearance and the high variability of volume of distribution found in these children should be noted. These differences of 2-PAM pharmacokinetics as compared with that in normal subjects are likely to be due to changes in hemodynamics during OP poisoning. A study that involved nine patients poisoned with various OPs (dimethoate, ethyl parathion, methyl parathion, and bromophos) indicated that pharmacokinetic data after intravenous pralidoxime methylsulfate were similar to those of pralidoxime chloride (Willems et aI., 1992). Thus, after a loading dose of 4.42 mg kg- l pralidoxime methylsulfate, followed by a maintenance dose of 2.14 mg kg- l h- l , plasma levels in these patients ranged from 2.12 to 9 mg liter- l . Calculated pharmacokinetic data (means ± SD) were total body clearance of 0.57 ±0.27literkg- 1 h-l, elimination half-life of 3.44 ± 0.9 h, and volume of distribution of 2.77 ± 1.45 liter kg-I. Pharmacokinetic differences between healthy volunteers (see previous discussion) and poisoned patients also have been reported for obidoxime. Pharmacokinetic parameters were calculated under a steady state condition in the case of a patient severely poisoned with methamidophos and complicated by renal failure. This patient was given 4 mg kg- l obidoxime intravenously over 30 min every 6 h (Bentur et aI., 1993). The obidoxime half-life was 6.9 h, the volume of distribution was 0.845literkg- l , and the total body clearance 85.4 mlmin-l. Obidoxime plasma concentrations ranged from a preinfusion valueof5.6 ).Lgml- l to 20.8 ).Lgml- l during the infusion. These values are comparable to the reported values in healthy volunteers. On the contrary, the plasma half-life was longer, the
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Clinical Toxicology of Anticholinesterase Agents
clearance was lower, and the volume of distribution was larger. The most likely explanation for such differences is the renal insufficiency that affected this patient. In conclusion, pharmacokinetic parameters should be assessed cautiously because of changes in hemodynamics, particularly because of the reduction of renal blood flow produced by severe OP intoxication (Green et aI., 1985; Jovanovic, 1989). Each oxime has different reactivation power on a given phosphorylated AChE and different phosphoryl residues attached to AChE are not equally susceptible to the same oxime (Aldridge and Reiner, 1972; Durham and Hayes, 1962; Kassa and Cabal, 1999; Worek et aI., 1996, 1998a, b, 1999a). Several authors reported limited or no efficacy of oximes in the treatment of OP poisoning (Besser et aI., 1995; Bismuth et aI., 1992; de Silva et aI., 1992; Erdmann et aI., 1966; Singh et aI., 1995; Tafuri and Roberts, 1987; Willems et aI., 1993). Poisonings thought to be resistant to reactivation therapy involved chrotoxyphos, demeton, dimethoate, dimefox, methyl-phenkapton, shradan, prothoate, and triamiphos (summarized in Bismuth et aI., 1992). Whereas these compounds form phosphoryl-AChE complexes that are identical to those formed by other OPs for which oximes have been found effective (Worek et aI., 1999a), it seems to be inappropriate to consider resistance as the cause of lack of effects; other reasons should be sought, including the dose of oxime and duration of treatment, which are often insufficient (John son et aI., 1992; Willems et aI., 1993), the pharmacokinetics of the oxime, which may be quite variable depending on the clinical conditions of the patient, and the rates of oxime-induced reactivation of AChE (Worek et aI., 1996). In this respect, the plasma concentration of the OP is relevant. In cases of ethyl and methyl parathion poisonings, oximes were shown to be ineffective as long as OP concentrations remained above 30 J.l.g liter- I (Willems et aI., 1993; also see subsequent text). In one study, 21 severe to moderately poisoned patients treated with atropine alone were compared with 24 patients treated with atropine plus pralidoxime (de Silva et aI., 1992). Patients were poisoned by a variety of OPs, including malathion, methamidophos, fenthion, dimethoate, phoxim, phentoate, and trichlorfon. Because the clinical outcomes were similar in the two groups, the authors cast doubt on the necessity of cholinesterase reactivators for the treatment of acute OP poisoning. This report was criticized because of the high mortality in both groups and because low doses of pralidoxime were used (John son et al., 1992). Moreover, given the specific characteristics of the phosphoryl residue attached to the enzyme in the rates of aging and of reactivatibility of AChE, the lack of measurement of plasma concentrations of OPs, and the different dosages of atropine, the study and the control groups are not comparable. Finally, there is no precise endpoint to assess the efficacy for a given oxime treatment when key variables that influence the clinical outcome are involved, such as mechanical ventilation. In conclusion, data on the inefficacy of oximes are not convincing and this potentially lifesaving therapy cannot be dismissed. On the contrary, such a course is highly recommended
in any case of severe OP poisoning and treatment should continue as long as there is circulating OP. However, because OP plasma concentrations cannot be obtained easily, some empirical approaches have been suggested to assess the need for oxime treatment. For instance, measurements of RBC AChE before and after a bolus of oximes, or in vitro reactivatibility of AChE from RBC sampled from the patient may indicate whether newly inhibited AChE can be reactivated (Lotti, 1995). Other methods have been described based on in vitro inhibition of cholinesterases by the plasma of the patient (Dawson et aI., 1997; Mahieu et aI., 1982). Nevertheless, because reactivation of blood enzymes may not strictly reflect that in the synapses and because even minimal reactivation in the nervous system is likely to be beneficial, these approaches and their results should be regarded as a guide and not as a rule. Dosing regimes for various oximes that depend on the severity of poisoning have been suggested. The following regimes are recommended by the manufacturers. • Pralidoxime chloride: Start with 1 g intravenously, followed by another 1 g after 15-30 min if no improvement. If still no improvement, start an infusion of 0.5 g h -I. Slow intravenous administration is preferable, but intramuscular injection also is an option. In healthy volunteers, the pharmacokinetics of pralidoxime chloride was similar with either route of administration (Sidell and Groff, 1971). • Pralidoxime methylsulfate: Start with 400 mg, followed by 200 mg after 0.5,4,6, and 12 h. In severe cases, start with 500 mg, repeat the same dose after 30 min, and then give 200 mg in repeated doses up to 2 g in 24 h. Continuous infusions also may be used (Willems et aI., 1993), up to 500 mg/h in cases of slowly disposed OPs (Tush and Amstead, 1997). • Obidoxime chloride: Start with 250 mg and repeat the same dose within 2 h, or 3-6 mg kg-I once or twice after poisoning. • In children, a loading dose of 25-50 mg kg-I pralidoxime chloride is recommended followed by a continuous infusion of 10-20 mg kg- I h- I .
All these recommended dosage schedules are aimed at achieving a plasma oxime concentration of 4 mg liter- 1, which was shown to be effective when using pralidoxime methanesulfonate (Sundwall, 1961). This concentration subsequently has been used a target reference value for all oximes. Nevertheless, this reference plasma level cannot be generalized because the molar concentrations of 2-PAM, obidoxime, and other oximes are different. Moreover, when the effects of various salts are compared, their different water solubilities should be taken into account because the amount of free base may vary considerably in the administered dose (Durham and Hayes, 1962). In some patients with ethyl and methyl parathion poisoning, enzyme reactivation could be obtained with pralidoxime methyl sulfate concentrations as low as 2.88 mg liter- I, whereas in other patients, oxime concentrations as high as 14.6 mg liter- 1 had no effect. In such cases, the therapeutic effect of the oxime depends
51.1 The Cholinergic Syndrome
on the plasma concentration of ethyl and methyl parathion (Willems et aI., 1993). Insufficient oxime therapy also has been considered as a possible cause of the intermediate syndrome (Benson et aI., 1992). However, such a hypothesis does not explain why the intermediate syndrome is not influenced by atropine or why this condition selectivity affects certain neuromuscular junctions (see Section 51.2.3.1). Treatment with oximes is not reccommended in carbamate poisoning and in moderate poisoning by dimethylphosphate OPs because the rates of spontanous reactivation of dimethoxyphosphorylated and carbamylated AChE are fast. Moreover, with certain carbamates, more toxic complexes with oximes may be formed (Sterri et aI., 1979). Pralidoxime is thought to be effective only on peripherally inhibited AChE because the quaternary nitrogen atom does not allow the drug to cross the blood-brain barrier. However, clinical observation in humans and some experimental data on animals point to the contrary (Lotti and Becker, 1982a; Namba et aI., 1971). A possible explanation is an alteration of the blood-brain barrier brought about by OP poisoning. The toxicity of oximes was studied in human volunteers and iatrogenic effects were reported after treatment in some cases of OP poisoned patients (Marrs, 1991). In a study on volunteers, the acute and chronic toxicities of 2-PAM, P2S and 1, 1'-trimethylenebis(4-formylpyridinium chloride) (TMB4CI2) were compared. TMB4Cl2 and, to a lesser extent, P2S (doses higher than 2.5 g a day) were found to cause marked gastrointestinal disturbances. Minor reversible cardiovascular effects also were noted (Calesnick et aI., 1967). In a patient severely poisoned with coumaphos, sudden cardiac arrest was observed 2 min after beginning an infusion of pralidoxime methyliodide, initiated about 4 h after poisoning. Trifluperazine and chloropromazine were also found in the blood of the patient. After restoring cardiac activity oxime infusion was started again but asystole recurred (Scott, 1986). The presence of phenothiazines might have potentiated the OP toxicity. However, this effect was shown in humans only in a single case report (Arterberry et aI., 1962) and results from research on experimental animals are conflicting (Fernandez et aI., 1975; Michaleck and Stavinoha, 1978). Cardiac arrest was probably coincidental. Another case of worsening symptomatology coincident with the beginning of 2-PAM treatment was reported, although there were some notable differences (Good et aI., 1993). A 51-yearold man presented at the hospital with limited signs consistent with moderate acute OP toxicity after several weeks of exposure to phosmet. RBC AChE was normal. Within several minutes of 2-PAM infusion, the patient had systemic and ventilatory weakness that required intubation. Electrophysiology revealed a subacute postsynaptic neuromuscular syndrome associated with some CNS dysfunctions that lasted for several weeks. One suggestive explanation offered by authors is that desensitization cholinergic receptors was produced by prolonged excess cholinergic stimulation and calcium influxes damaged the neuromuscular junction. The recommendation
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was made not to use oximes in the presence of postsynaptic dysfunction because of the possible direct effects of oximes on the receptor itself (Alkondon and Albuquerque, 1989). It remains to be explained why this patient had normal RBC AChE at the onset and during the entire course of the illness. There is some indication that further inhibition of AChE might occur during reactivation with 2-PAM (de Jong and Ceulen, 1978). Two phosphoryl oximes that formed during reactivation of the ethoxy methylphosphonyl-AChEconjugate by two oximes (LiiH6 and TMB4) have been detected; they have not been detected during the reactivation of diethylphosphorylAChE conjugates. These phosphoryl oximes were found to be potent inhibitors of AChE, although usually they are likely to be hydrolyzed by paraoxonase (Luo et aI., 1999). The worsening of symptomatology after treatment of massive poisonings with certain oximes may be due to accumulation of phosphoryloximes that occurs fast enough to saturate paraoxonase activity. It has been suggested that these events might also explain the pathogenesis of intermediate syndrome (Luo et aI., 1999), although intermediate syndrome has been observed in many patients poisoned with diethyl OPs (see Section 51.2.3). Mild biochemical signs of liver toxicity have been related to the use of oximes. The symptoms disappeared when treatment was discontinued and seemed more frequent with obidoxime (Balali-Mood and Shariat, 1998; de Kort et aI., 1988). Diazepam Diazepam must be included in the treatment of acute OP poisoning in all but the mildest cases (WHO, 1986). Diazepam relieves anxiety in mild cases, and reduces muscle fasciculations and antagonizes convulsions in the more severe cases (John son and Vale, 1992; Namba et aI., 1971; Willems and Belpaire, 1992), although OP-induced convulsions are usually reduced by large doses of atropine (Vale and Scott, 1974). Moreover, animal data indicate that benzodiazepines improve morbidity and mortality in OP poisoning (Boskovic et aI., 1984). Doses of diazepam (10-20 mg) given subcutaneously or intravenously are recommended and may be repeated as needed (Minton and Murray, 1988). Rarely, other anticonvulsants such as phenytoin have been used successfully in OP poisoning cases (Sellstrom, 1992). Supportive Treatment The cornerstone of supportive treatment for severe poisoning is artificial ventilation, which must be started at the first signs of respiratory insufficiency. In such cases, admission to intensive care facilities is mandatory. Supplemental oxygen may be required to correct hypoxemia, and adjustments of fluid intake and electrolyte balance should be made as necessary. In severely ill patients, it may be necessary to maintain cardiac and urinary output pharmacologically. Prophylaxis of infections and ad hoc treatment of cardiac arrhythrnias are also necessary. Sequence of Treatment The sequence of first aid maneuvers depends on the circumstances and on the severity of poisoning. In cases of poisoning in the field, where antidotes are rarely
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CHAPTER 51
Clinical Toxicology of Anticholinesterase Agents
Table 51.7 Suggested Sequence of Therapeutic Approaches to Acute OP Poisoning According to Severitya What to do Clinical conditions
First
Second
Third
Fourth
Mild
Decontamination
Atropineb
Diazepamc
Observation
Fifth
(bolus) Moderate Severe
Atropineb
Decontamination
Diazepamc
Pralidoximed
Observation
Artificial ventilation
Diazepamc
Atropineb
Pralidoximed
Decontamination
(infusion) aModified from Lotti (1991). See Table 51.5 for grading severity. bS ee Table 51.6 for dose. c]Omgs.c. d See dose in the text.
available, the patient should be rushed to the hospital: very cautious decontamination may be tried in the case of dermal exposure, but in the case of ingestion, vomiting should not be induced. After the patient is evaluated in the hospital, the sequence of treatment is dictated by the severity of the clinical picture. Suggestions are given in Table 51.7. 51.1.3.9 Late Complications Neurological Although it is known that the recovery time for some effects exceeds, to a limited extent, the time to replace AChE (Bowers et aI., 1964; Namba et al., 1971; Whorton and Obrinsky, 1983), these cholinergic signs and symptoms that last up to several weeks after peak effects will not be considered. Moreover, OPIDP, a well established toxicity of some OPs, will be discussed separately (see Section 51.3). Several neurologic, psychiatic, and neurobehavioral abnormalities have been observed in patients who suffered previous acute poisoned with OPs, although prospective studies are not available. These symptoms and signs recently were conceptualized as a syndrome called chronic OP-induced neuropsychiatric disorders (COPIND; lamal, 1997), together with similar symptoms also observed after long-term low-level exposures. COPIND after acute poisoning has been labeled phenomenon 1, whereas that after long-term exposure has been labeled phenomenon 2. Similarities between the two phenomena are few and superficial, many findings are contradictory and inconsistent, and given the different types of exposures, there is no reason to believe that they belong to the same entity. Moreover, possible neurological, psychiatric, and behavioral effects either after acute or after low-level long-term exposures would be better appreciated if a distinction is maintained (see Section 51.4.1). Like any condition associated with prolonged hypoxia, severe OP poisoning obviously could lead to various persistent neurological disorders of the CNS. Therefore, in cases of late CNS disturbances, assessments of the severity of the poisoning and of time elapsed between onset of symptoms and the beginning of treatment are required to distinguish between primary and secondary effects of OPs. For instance, a patient severely poisoned with sarin during the terrorist attack in Tokyo pre-
sented 6 months later with retrograde amnesia (Hatta et aI., 1996). Because the patient was hypoxic for several minutes, it is impossible to ascertain whether the cause of amnesia was a direct biochemical effect of sarin. An unusual syndrome was reported in a patient 10 weeks after discharge from the hospital following acute poisoning with dimethoate (Sahin et aI., 1994). The patient presented with erythema edema and hyperesthesia in the hands associated with pain and limited movements. Upper arm electrophysiological studies revealed bilateral neuropathy and a bone scan detected increased osteoblastic activity of the hand bones. A diagnosis was made of reflex sympathetic dystrophy, which was associated with the previous poisoning. This was probably a coincidental association because the report was isolated, dimethoate does not cause OPIDP, and when OPIDP does occurs it does not exclusively affect the upper limbs or the bones (see Section 51.3). After ingestion of bromophos, a patient presented with no signs of cholinergic overstimulation and the only evidence of exposure was a reduction of plasma cholinesterases to about 10% of normal values. Oxime was given, but no atropine (Michotte et aI., 1989). The patient was under treatment with maprotiline for a recurrent unipolar depressive disorder. Five weeks later the patient developed a cerebellar ataxia that subsided after 5 weeks. This case was likely a casual association, perhaps due to uncontrolled dosing with maprotiline. EEG changes in industrial workers with past repeated accidental exposures to the warfare agent sarin have been reported (Duffy et aI., 1979). These exposures were not quantified, occurred at least 1 year prior to examination, and caused symptoms as well as significant RBC AChE inhibition. However, it is not clear whether cases of frank poisoning occurred. Some individuals had up to six such episodes. A number of differences, derived from complex analysis of EEG spectra, were observed between 77 exposed workers and 38 controls from the same factory but not exposed to sarin. These changes include increased f3 activity, increased 8 and e slowing, decreased et activity, and increased amounts of rapid eye movement sleep. Most of these changes were detected in the temporal and occipital lobes. Some controversies exist concerning the value
51.1 The Cholinergic Syndrome
of computerized analysis of brain wave topography (American Electroencephalographic Society, 1987; Duffy et aI., 1986; Oken and Chiappa, 1986). In a commentary on these and similar results observed in animals (Burchfiel et aI., 1976) the toxicological significance of these findings was questioned (Duffy and Burchfiel, 1980). The EEGs of 100 individuals with previous acute OP poisoning (one or more episodes occurring from 3 months up to 25 years before the survey) were compared with those of matched controls (Savage et aI., 1988). Several OPs were involved in the poisonings, including methylparathion, parathion, malathion, disulfoton, mevinphos, dicrotophos, TEPP, dioxathion, DEF, and phorate. Poisoned cases had slightly more abnormal EEGs, but results were not significantly different between the matched cases and the control cohort. Although the authors stated that poisoning documentation was screened for completeness, some information was missing, such as the clinical severity of poisoning, the toxicological evidence of poisoning, and the nature of intercurrent diseases. For instance, one exclusion criterion was head trauma with period of unconsciousness totaling more than 15 min, but the clinical conditions of cases with unconsciousness less than that were unreported. A retrospective study examined the vibrotactile thresholds in three groups of subjects: (1) previously poisoned with a variety of OPs (15 subjects), (2) poisoned with methamidophos (21 subjects), and (3) a matched control (35 subjects; McConnell et aI., 1994). The results indicated that over one-fourth of the subjects previously poisoned with methamidophos, known to cause OPIDP (Senanayake and Johnson, 1982), had higher vibrotactile thresholds, but similar though less pronounced effects were seen in subjects poisoned with other OPs not known to cause OPIDP. The authors concluded that classical OPIDP is only the worst disease caused by methamidophos in a spectrum of peripheral nervous system impairments that represent the sequelae of poisoning. However, toxicological and clinical assessment of the poisoning episodes were not reported, and the elevated vibrotactile threshold was not symmetrical and also detected in the fingers. In mild toxic axonopathies, lesions are confined to the lower limbs and are characteristically symmetrical; hence, the described findings are not consistent with a toxic neuropathy. Moreover, subjects were examined 1-3 years after the acute poisoning and a toxic peripheral neuropathy is likely to recover over this period of time if exposure ceases. In a retrospective study (Steenland et aI., 1994),83 subjects exposed to a variety of OPs (among which only chlorpyrifos, which accounted for 10 cases, is known to cause OPIDP), who had one or more symptoms compatible with poisoning and documented inhibition of either RBC AChE or plasma BuChE (more than 20% of baseline or below normal range), showed significant alterations of vibrotactile sensitivity of fingers and toes compared with a control group (90 subjects). It is not clear when testing was performed, although it appears that it was done several years later. Because actual electrophysiological and vibration sensitivity data were not reported, and the clinical and toxicological data are not comprehensive, it is difficult to assess the biological significance of such changes. Moreover, as
1061
stated before, involvement of the arms is not expected in OPIDP unless it is extremely severe and a toxic peripheral neuropathy is expected to have recovered years after cessation of exposure. A similar retrospective study was reported by the same group of investigators (Ames et aI., 1995). In 45 asymptomatic subjects who had a history of cholinesterase inhibition short of frank poisoning (RBC ::; 70% of baseline or plasma cholinesterase ::;60% of baseline), some electrophysiological parameters were measured and no differences were found between cases and 90 controls. The OPs involved were not identified. The difference between cases in this study and those of the preceding study seems to be due to the presence of at least one symptom of cholinergic overstimulation in the subjects of the former study group, whereas the latter group had none. Cholinesterase inhibitions were overlapping in the two studies. The authors concluded that preventing acute organophosphate poisoning also prevents neurological sequelae. Knowing the limitations of measurements of cholinesterase inhibition, of reporting symptoms common to both OP toxicity and to a variety of other conditions, and of electrophysiological studies on the upper limbs, and in the absence of detailed clinical data in one of the studies, it is difficult to compare the two studies or to agree with the authors' conclusions about the existence of chronic sequelae of acute OP poisoning, which would not occur unless substantial AChE inhibition had occurred. In a study carried out in 1983-1984 and published in 1996, volunteers exposed to sarin concentrations that caused 3040% RBC AChE inhibition showed mild electrophysiological changes up to 15 months after a single exposure. The small number of subjects, the high variability of changes, and the fact that only three out of eight subjects displayed persistent changes make this study difficult to interprete (Baker and Sedgwick, 1996). In a cross-sectional study on 164 pesticide workers, a correlation was found between past OP poisoning and increased incidence of symptoms such as dizziness, sleepiness, and headache (London et aI., 1998; see also Section 51.4.1.2). No evidence of correlation with past poisoning was found when vibration sense and tremor were evaluated and found to be unaltered. Moreover, clinical details of OP acute poisoning were not given. Psychiatric Follow up studies based on interview, physical examination, and blood chemistry on long-term sequelae of acute OP poisoning revealed no significant serious neuropsychiatric effect in a group of 114 individuals, 6 of whom had severe poisoning and the others mild to moderate poisoning (Tabershaw and Cooper, 1966). An array of different symptoms reported by subjects was not considered to be associated with the poisoning. However, the authors conceded that their study would not reveal minor after effects or those of low incidence. Preexisting psychiatric symptomatology has been reported to have worsened over 2 years after OP poisoning and whenever small further exposures to OPs occurred (Rosenthal and Cameron, 1991). Details of the poisoning of this patient were not given.
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CHAPTER 51
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Neurobehavioral In the study described previously, combined clinical and neuropsychological evaluations were used to detect changes in the cognitive functions in a group of 100 subjects with previous acute OP poisonings compared with a matched control group (Savage et aI., 1988). In this study, however, the limitations already outlined raise the question whether these changes represent a consequence of brain hypoxia or of other intercurrent factors, given the very large variability in the time elapsed from poisoning to assessment. Most differences between the two groups, detected on a number of tests, were within normal variability. Certainly other factors such as educational differences might account for differences in comprehension, arithmetic, vocabulary, etc. Moreover, toxicological analysis showed that blood levels of organochlorine pesticides in the study group were about twice those of the controls. Whereas statistical analysis failed to show any association between such blood levels and the results of neuropsychological tests, the authors ruled out organochlorine as the causative agent of such impairments. However, from a toxicological viewpoint, organochlorine exposure might have been more relevant, given the pharmacokinetic differences between these pesticides and OPs, especially because organochlorine is more easily stored in the body. In a retrospective cohort study (Rosenstock et aI., 1991), a group of 36 subjects previously poisoned with OPs were tested on average about 2 years after the episode of poisoning and compared with a matched control group. The poisoned group did much worse than the control group on several neuropsychological tests (visual and verbal attention, visual memory, visuomotor activities, and dexterity). The type of OPs involved and the severity of poisoning were not reported, and the design and statistical significance of the study were criticized (Schuman and Wagner, 1991). Moreover, subjects in the control group (a close friend or sibling in the same community who had never been treated for pesticide poisoning) were also occupationally exposed to OPs. Given the endpoints used (for instance, visuomotor performance) and the lack of follow-up studies, it is possible that the neuropsychological deficits were a cause rather than a consequence of OP poisoning. A neuropsychological test battery was administered to 21 migrant farm workers who had been acutely exposed to phosdrin and other pesticides (lannate and maneb) and to matched controls (Reidy et aI., 1992). Two acute exposures occurred 3 years apart and subjects were examined 2 years after the second exposure on the occasion of a worker's litigation. The exposed group was significantly more impaired than controls on tests of psychomotor speed, dexterity, and visuospatial memory. Although symptomatic, RBC AChE and plasma BuChE were normal on both occasions. Therefore, if related these changes were to pesticide exposures, they cannot be a consequence of OP poisoning, but perhaps of the other involved pesticides. In the previously quoted study (Steenland et aI., 1994), several behavioral parameters were tested, but only sustained activity was found to be worse in the case group than in the controls. In another previously mentioned study (Ames et aI., 1995), a number of neurobehavioral tests were performed on
subjects who had a history of cholinesterase inhibition "short of frank poisoning." Only one test (serial digit performance) was statistically significant, but it was opposite to the hypothesized direction. In conclusion, there is little evidence to support the notion that acute OP poisoning may result in late permanent toxic effects other than OPIDP if hypoxia and/or severe uncontrolled convulsions did not occur or did not last for sufficient time. Similar conclusions have been reached by others (Ray, 1998a, b). Moreover, ranking all these effects under one syndrome (COPIND phenomenon 1) is inappropriate and may be misleading.
51.2 THE INTERMEDIATE SYNDROME The intermediate syndrome is characterized by weakness of respiratory, neck, and proximal limb muscles. It is not a direct effect of AChE inhibition and appears several hours after the beginning of signs and symptoms of severe cholinergic overstimulation. It is caused by a variety of OPs and seems to be related to postsynaptic effects. This form of OP toxicity was first conceptualized by Senanayake and Karalliedde (1987), although the first accurate description of the syndrome was given by (Wadia et aI., 1974), who categorized the neurological manifestations of acute OP poisoning into two groups. Type 1 signs are the classical signs of cholinergic overstimulation, whereas type 2 signs, which appear later and while undergoing atropine treatment, are characterized by proximal weakness and cranial nerve palsies. New case reports and prospective studies or cases derived from retrospective analysis of medical records have been reported. 51.2.1 ETIOLOGY
Intermediate syndrome seems to occur in 20-50% of acute OP poisoning cases (Sedgwick and Senanayake, 1997). It has been observed after exposures to several OPs, including fenthion (De Wilde et aI., 1991; Karademir et aI., 1990; Senanayake and Karalliedde, 1987), omethoate (He et aI., 1998), dimethoate (De Bleecker et aI., 1993; He et aI., 1998; Senanayake and Karalliedde, 1987), methamidophos, monocrotophos (Senanayake and Karalliedde, 1987), diazinon (Wadia et aI., 1974), demeton S-methylsulfone (Besser et aI., 1989a), trichlorfon (Karademir et aI., 1990), parathion (De Bleecker et aI., 1993; He et aI., 1998), methylparathion (De Bleecker et aI., 1993) dichlorvos (He et aI., 1998), phosmet (Good et aI., 1993), and malathion (Gadoth and Fisher, 1978), and to various mixtures of OPs (De Bleecker et aI., 1993; He et aI., 1998). 51.2.2 PATHOGENESIS
The mechanism by which the intermediate syndrome develops is unknown. The first characterization of the syndrome suggested a postsynaptic effect based on e1ectromyographic
51.2 The Intermediate Syndrome
evidence of fade on tetanic stimulation, absence of fade on lowfrequency stimulation, and absence of posttetanic facilitation (Senanayake and Karalliedde, 1987). This concept was further supported by morphological and electrophysiological studies, both in humans (De Bleecker et al., 1992b, 1993; Sedgwick and Senanayake, 1997; Singh et al., 1998a, b) and animals (Engel et al., 1973), suggesting that muscle weakness may result from cholinergic receptor desensitization due to prolonged cholinergic stimulation. Therefore, the hypothesis was put forward that the pathophysiology of intermediate syndrome is the result of a time-confined phenomenon that includes both changes in the postsynaptic structures by desensitization and restoring the ratio of acetylcholine to AChE (De Wilde et al., 1991). This process may explain the observation of an unusual case of respiratory failure precipitated by 2-PAM in a patient thought to have had prolonged cholinergic overstimulation by phosmet (Good et al., 1993). Neuromuscular block may have been increased because of a sudden reduction of acetylcholine levels that had caused the postsynaptic dysfunction. Similarly, a patient poisoned with oxydemeton-S-methyl was comatose shortly after poisoning, but responded to noxious stimuli. However, such response decreased suddenly 42 h after intoxication and the electrophysiological investigation performed 66 h after poisoning detected severe neuromuscular dysfunction characterized by the decrement phenomenon after repetitive nerve stimulation. At this point in time, obidoxime was given and the neuromuscular block worsened (Besser et al., 1995). With this interpretation, the hypothesis that the intermediate syndrome may result from excessive and persistent acetylcholine levels due to insufficient oxime therapy during the early acute phase of cholinergic toxicity might have some validity (Gadoth and Fisher, 1978; Benson et al., 1992). Another hypothesis relates the development of intermediate syndrome to the formation of oxime-phosphoryl complexes, which greatly inhibit AChE and not efficiently cleaved by paraoxonase (Luo et al., 1999; see Section 51.1.3.8). Muscle necrosis, the human equivalent of a well known effect observed in experimental animals (Dettbarn, 1992), was also thought to be a possible cause of intermediate syndrome. Myopathy was described by de de Reuck and Willems (1975), Ahlgren et al. (1979), de Reuck et al. (1979), and Wecker et a!. (1985; see Section 51.1.3.6). However, the clinical evidence of muscle necrosis was not consistent in a 21 case series because only about one-half of the patients had elevated CPK and LDH (He et a!., 1998). Moreover, the histopathological lesions were too limited (De Bleecker et aI., 1993) to support the notion that intermediate syndrome is due to muscle necrosis. In conclusion, in support of the explanation that intermediate syndrome is a consequence of desensitization of nicotinic receptors is the observation that several OPs, quite different chemically and toxicologically, did produce intermediate syndrome and that all patients had a prolonged AChE inhibition and consequent high levels of acetylcholine. However, the reasons for the selectivity for some nicotinic receptors only, as shown by the clinical features of intermediate syndrome, remain unexplained.
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51.2.3 CLINICAL MANIFESTATIONS 51.2.3.1 Clinical Signs and Course The intermediate syndrome develops during recovery from cholinergic manifestations, one to several days after the poisoning. Distinction should be made between this syndrome and the recurrence of cholinergic toxicity, which may occur with OPs that display prolonged disposal (Davies et aI., 1975; Ecobichon et aI., 1977; Gadoth and Fisher, 1978; Molphy and Rathus, 1964; Perron and Johnson, 1969; see Section 51.1.3.1). In some cases, the sudden onset of intermediate syndrome occurs in patients when they are completely recovered from the initial cholinergic crisis (Senanayake and Karalliedde, 1987; Wadia et aI., 1974), whereas in others, it is concurrent with muscarinic signs of toxicity or with superimposed muscarinic relapses (De Bleecker et aI., 1993). Concurrent and recurrent cholinergic signs are controlled by atropine, whereas those of the intermediate syndrome are not. A constant feature in all patients is a marked weakness of neck flexion and of proximal limb muscles. These patients are unable to raise their head off the pillow, to abduct their shoulders' or to flex their hips. A sudden respiratory failure due to weakness of respiratory muscles also characteristically occurs in 70-100% of the cases, often drawing attention to the onset of the syndrome. Upper and lower limb reflexes are often reduced or absent, and in several patients there was evidence of weakness of muscles innervated by cranial nerves. One or more nerves may be involved, including the Ill, IV, VII, IX, X, and XI. There is no distinct pattern in the development of these signs. Mortality due to respiratory paralysis and complications ranges from 15 to 40%. The clinical course in surviving patients lasts up to 30-40 days. Regression of cranial nerve palsies appears first, followed by improvement of respiratory insufficiency and recovery of strength in the proximal limb muscles. Neck flexion is the last function to recover. 51.2.3.2 Electrophysiology Electrophysiological studies have been performed on a few patients with intermediate syndrome (De Bleecker et aI., 1993; Sedgwick and Senanayake, 1997; Senanayake and Karalliedde, 1987; Singh et al., 2000; Wadia et aI., 1987). Nerve conduction velocity was normal and distallatencies were either normal or slightly reduced. No signs of spontaneous activity, such as fibrillation potentials or positive sharp waves, were observed. Repetitive stimulation showed decrements of CMAP at low and/or intermediate frequencies in most patients (l0-50 Hz), lasting for several days, but always disappearing before clinical normalization, indicating that the neuromuscular junctional dysfunction in the intermediate syndrome is likely postsynaptic. Electrophysiological changes that occur after repetitive stimulation have been detected both in poisoned patients with clinical evidence of intermediate syndrome and in patients with acute poisoning without subsequent development of intermediate syndrome (De Bleecker et aI., 1993). In the latter group,
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CHAPTER 51
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these changes followed recovery from initial depolarization block due to acetylcholine excess (see Section 51.1.3.5). It is, therefore, not clear whether this postsynaptic block is always present in all cases of acute OP poisoning, beginning a few days from the initial cholinergic syndrome. In such a case, the reasons for the switch from subclinical to a clinically evident intennediate syndrome are unknown.
51.2.3.3 Pathology Muscle histopathology was perfonned in very few patients (De Bleecker et aI., 1992a, b, 1993). Few and scattered necrotic muscle fibers were observed. Histochemical endplate AChE staining was variable in intensity. Ultrastructural examination showed neuromuscular junctions that contained numerous synaptic vesicles, swollen mitochondria, and synaptic clefts with well established basal lamina. Some endplates showed vesiculations and phagocytic lysosomal activity suggestive of degeneration. Some of the synaptic clefts were widened and filled with debris, junctional folds were simplified, and postsynaptic areas were denuded and degenerated.
51.2.3.4 Treatment Treatment is exclusively supportive because there is no specific treatment for the intennediate syndrome and atropine is not effective. Endotracheal intubation and mechanical ventilation are lifesaving.
51.3 DELAYED POLYNEUROPATHY Organophosphate-induced delayed polyneuropathy is a rare toxic effect in humans, although some epidemics have occurred in the past such as the famous Ginger-Jake paralysis when thousands of patients were intoxicated with triorthocresyl phosphate (TOCP; Morgan, 1982; Inoue et aI., 1988). Neuropathy is characterized by a symmetric, distal sensory-motor, centralperipheral axonopathy that affects the legs and, in the most severe cases, also the arms. OPIDP is mechanistically unrelated to cholinergic and intennediate syndromes, and, therefore, it is not necessarily associated with the anticholinesterase activity of OPs. In fact, several OPs (such as triarylphosphates) are devoid of this activity, but may cause OPIDP. A large body of experimental data indicate that this axonopathy is likely to be correlated with effects on a neural esterase known as neuropathy target esterase (NTE; Johnson, 1990; Lotti, 1992b). Clinical onset is delayed for up to 10-20 days after a single exposure and for an unspecified period after continuous exposures.
51.3.1 ETIOLOGY Not all OPs are capable of causing OPIDP and in the case of OP insecticides currently in use, polyneuropathy develops exclusively after a severe episode of cholinergic toxicity. Because OPIDP is not a consequence of acute cholinergic toxicity,
Table 51.8 a Comparative Sensitivities of Human AChE and NTE for Various Inhibitors
Compoundb Dichlorvos Methamidophos (L isomer) Chlorpyrifos-oxon Mipafox Phenylsaligenin phosphate
AChE 150
NTE 150
(I-tM)
(I-tM)
0.95 100 0.01 >100 0.12
AChE 150 NTEIso
16
0.06
400
0.2
0.2 12 0.003
0.05 >1 39
aData from Lotti and Johnson (1978), Bertolazzi et al. (1991), and Capodicasa et al. (1991). bDirect acting OPs or metabolites (chlorpyrifos oxon from chlorpyrifos and phenylsaligenin phosphate from TOCP).
but just another toxic effect of OPs, in the case of insecticides it will develop at a much higher dose than that which causes cholinergic overstimulation. This is indirectly shown in Table 51.8, where sensitivity to various inhibitors of AChE (target of cholinergic toxicity) and of NTE (target of OPIDP) derived from human tissues are compared. OP insecticides that caused OPIDP in humans (Table 51.9) produced OPIDP only after cholinergic toxicity. Mipafox was never developed as an insecticide and caused OPIDP after mild cholinergic toxicity. Phenyl saligenin phosphate is the active metabolite of TOCP (not an insecticide), which caused several cases of OPIDP without cholinergic toxicity. Similar differences were seen with hen enzymes, which in turn correlate with the capability of a given OP to cause OPIDP, relative to that of causing death (Lotti and Johnson, 1978). Therefore, compounds with AChE 150INTE 150 ratios> 1 may cause OPIDP without cholinergic toxicity, whereas those with a ratio < 1 will cause OPIDP only after cholinergic toxicity and appropriate antidotal treatment. OPIDP displays a characteristic age-related sensitivity in both experimental animals and humans. Children are resistant to OPIDP (Goldstein et aI., 1988) and when they are affected, they recover much quicker than adults, usually within a few months (Senanayake, 1981). In addition to the case reports listed in Table 51.9, other reports can be found where development of OPIDP was associated with single or short-term exposures to certain OPs, although clinical and toxicological evidence was not convincing. EPN and leptophos cause OPIDP in hens and are NTE inhibitors, but only at doses that cause severe cholinergic toxicity (John son, 1975; Ohkawa et aI., 1980). However, a report of OPIDP that involved several workers who had long-tenn exposure to EPN indicated little or no evidence of cholinergic overstimulation and most clinical details were missing. Moreover, during the release of EPN, these workers were also exposed to other chemicals derived from an explosion and fire in a manufacturing facility (Xintaras and Burg, 1980). An outbreak of neurological disorders occurred in a plant that manufactured leptophos (Xintaras et aI., 1978), which is
51.3 Delayed Polyneuropathy
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Table 51.9 OP Insecticides that Cause Delayed Polyneuropathy in Humans a Compound
Number of cases
Circumstances
References
Chlorpyrifos
2
Suicide
Lotti et aI., 1986; Martinez-Chuecos et al., 1992
Dichlorvos
3
Suicide
Vasilescu and Florescu, 1980; Wadia et aI., 1985
Isofenphos b
3
Suicide
Catz et aI., 1988; Tracey and Gallagher, 1990; Moretto and Lotti, 1998
Methamidophos
Several
Suicide/occupational
Senanayake and Johnson, 1982; Moretto and Lotti, 1998; McConnell et aI., 1999
Mipafox
2
Occupational
Bidstrup et aI., 1953
Trichlorfon
Several
Suicide
Hierons and Johnson, 1978; Johnson, 1981; Shiraishi et aI., 1983; Niedziella et aI., 1985;
Trichlomate
2
Suicide
Jedrzejowska et aI., 1980; Willems, 1981
Csik et aI., 1986
a Modified
from Lotti (2000). All cases displayed preceding cholinergic toxicity. bOne of these cases was a combined exposure with phoxim.
known to cause OPIDP in experimental animals (Hollingshaus et aI., 1981). Three subjects had signs compatible with OPIDP at medical examination and six had symptoms in a retrospective study. Several subjects, however, had neurological signs unrelated to OPIDP and all were exposed to a variety of neurotoxic chemicals, including n-hexane. Another group of case reports suggested OPIDP development after exposures to omethoate, parathion, mecarbam, fenthion, and mevinphos. However, these pesticides are not NTE inhibitors and negative results have always been reported in the hen test (FAOIWHO, 1987, 1996, 1997; Johnson, 1975; Lotti, 1992b). A polyneuropathy compatible with OPIDP developed after a suicide attempt with omethoate (Curtes et aI., 1981), but in a subsequent report of a man who died shortly after an acute omethoate poisoning, no NTE inhibition was detected in post mortem nerve tissues (Lotti et aI., 1981). Parathion was associated with OPIDP after massive poisoning together with methanol (de Jager et aI., 1981). Toxicological evidence of parathion in body fluids was missing and the clinical description does not support evidence ofRBC AChE inhibition or methanol poisoning (Lotti and Becker, 1982b). Moreover, several cases of severe parathion poisoning resulted in no OPIDP (Namba et aI., 1971). Mecarbam was reported to cause neuropathy, but nerve biopsy revealed segmental demyelination without axonal degeneration (Stamboulis et aI., 1991), a morphological lesion not expected in OPIDP (see Section 51.3.3.3). A case of delayed neuropathy apparently developed 1 month after acute poisoning by fenthion, which was followed by intermediate syndrome (Karademir et aI., 1990). It is not clear whether diagnosis was made on clinical grounds or exclusively on electromyography (EMG). Because the results ofEMG studies were not reported, interpretation of this case is difficult. One further case of delayed polyneuropathy by fenthion was reported (Martinez-Chuecos et aI., 1992), but no clinical details were given. Moreover, another patient poisoned with fenthion, and belonging to the same series, did not develop neuropathy. A female patient who attempted suicide with methylparathion, fell into a deep coma that lasted 4 weeks (Nisse et aI.,
1998). Electromyography was normal 3 weeks after poisoning, whereas 1 week later it showed signs of mild distal sensory motor polyneuropathy. Diffuse myogenic alterations were also detected, but electrophysiological data were not reported. Because the neuropathy disappeared within 4 weeks, this was unlikely a case of OPIDP and was probably a consequence of prolonged coma. A case of severe poisoning by mevinphos was reported to have been complicated by polyneuropathy (Hsiao et aI., 1996). No clinical or electrophysiological data were given. It is said that nerve conduction studies confirmed the neuropathy. In such a case, OPIDP would be unlikely, because conduction is usually, at the most, slightly affected in axonopathy. No followup was reported. An isolated case of Guillan-Barre-like syndrome was described in a patient after exposure to merphos, a defoliant with little anticholinesterase activity (Fisher, 1977). Exposure was dermal and likely very low. Four days later the patient started complaining of upper and lower limb weakness; 14 days after exposure, he developed facial diplegia. EMG and clinical features were suggestive of Guillan-Barre syndrome. This is the only case in the literature of Guillan-Barre-like signs after acute exposure to OP, but the clinical and electrophysiological characteristics of OPIDP are quite different. In conclusion, combined clinical and experimental evidence allows firm conclusions on OPIDP only for a few chemicals (Table 51.9). Nevertheless, it should be pointed out that neuropathic impurities may be present in commercial formulations, which perhaps accounts for some of these cases (Johnson, 1984). 51.3.2 PATHOGENESIS
The initial event in OPIDP is the inhibition of NTE, followed by aging of the phosphoryl NTE complex. These molecular changes occur within a few hours of exposure, but almost nothing is known about what happens between these events and the clinical, morphological, and electrophysiological onset of
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OPIDP 2-3 weeks later (Johnson, 1990; Lotti, 1992b). Limited evidence suggests that development of OPIDP in humans also involves inhibition of NTE. In two fatal cases of OP poisoning, NTE activity was measured post mortem in the peripheral nerve. As expected from animal studies, NTE was found to be inhibited in a case of chlorpyrifos poisoning (Osterloh et aI., 1983), whereas it was not in a case of omethoate poisoning (Lotti et aI., 1981). Lymphocytic NTE (see Section 51.3.3.2) was found to be inhibited soon after exposure in one case of poisoning with chlorpyrifos (Lotti et aI., 1986) and in two cases with methamidophos that all developed OPIDP weeks later (Moretto and Lotti, 1998; McConnell et aI., 1999). Lymphocytic NTE inhibition was also found in a patient poisoned with isofenphos who died on day 32; OPIDP might have developed after this time. In severe poisoning by compounds not known to cause OPIDP, lymphocytic NTE inhibition was not detected (Moretto and Lotti, 1998). After occupational exposures to DEF, substantial NTE inhibition in lymphocytes was measured, but found not to be associated with the development of OPIDP or with electrophysiological changes. In this case, lymphocytic NTE did not represent a good mirror of peripheral nerve NTE, probably because of the particular pharmacokinetics of this compound (Lotti et aI., 1983). 51.3.3 CLINICAL MANIFESTATIONS 51.3.3.1 General Features Symptoms of OPIDP begin 2-3 weeks after single doses when, as in the case of insecticides, cholinergic symptoms have subsided (Lotti et aI., 1984). The lag time between single or short-term exposure and the clinical onset of OPIDP depends on both the chemical involved and the dose (Bidstrup et aI., 1953; Lotti et aI., 1986; Senanayake and Johnson, 1982). OPs with slow pharmacokinetics may cause OPIDP after a prolonged period following exposure (up to 4 weeks), whereas higher doses of OPs that are powerful in causing OPIDP may shorten this period to about 10 days. Clinical features of OPIDP are usually fully expressed within a few days of the onset of symptoms and signs, and no progression has been observed in the absence of further exposure. After repeated exposures, such as those to nonanticholinesterase OPs, the onset of symptoms and their full development is more variable and less definible (Vasilescu, 1982). The usual initial complaint is cramping muscle pain in the legs (Susser and Stein, 1957), followed by distal numbness and paresthesia (Senanayake and Jeyaratnam, 1981; Vasilescu et aI., 1984). Progressive leg weakness occurs, together with depression of tendon reflexes. Symptoms and signs may also appear in the arms and forearms following those in the legs, but always after severe exposures (Bidstrup et aI., 1953; Moretto and Lotti, 1998; Senanayake and Jeyaratnam, 1981; Vasilescu, 1982; Vasilescu et aI., 1984). Physical examination reveals distal symmetrical predominantly motor polyneuropathy, with wasting and flaccid weakness of distal limbs muscles, especially in the legs. Signs
include a characteristic high-stepping gait associated with bilateral foot drop (Senanayake and Jeyaratnam, 1981). Severe OPIDP may result in quadriplegia with foot and wrist drop as well as mild pyramidal signs. In time, there is complete functional recovery if spinal cord axons have been spared by smaller doses (Senanayake, 1981); otherwise, pyramidal and other signs of central neurological involvement may become more evident. The degree of pyramidal involvement determines the prognosis for functional recovery, and spastic ataxia may represent a permanent outcome of severe OPIDP (Morgan and Penovich, 1978; Tosi et aI., 1994; Vasilescu, 1982). Objective evidence of sensory loss is usually slight or absent. In one group of patients poisoned with methamidophos, some sensory symptoms, but no objective signs, were recorded (Senanayake and Johnson, 1982). In two patients who developed OPIDP after exposure to chlorpyrifos and isofenphos, slight sensory alterations were detected during both physical and electrophysiological examination (Moretto and Lotti, 1998). However, in a series of patients, cases were reported where purely sensory peripheral neuropathy was displayed after repeated low exposures to chlorpyrifos that caused some symptomatology, and no signs or mild signs of cholinergic overstimulation (Kaplan et aI., 1993). This contrasts with the known toxicological characteristics of chlorpyrifos, which is a better inhibitor of AChE than NTE (Capodicasa et aI., 1991; Richardson, 1995), and the clinical features observed in two cases of OPIDP induced by chlorpyrifos, where OPIDP was always preceded by severe cholinergic overstimulation (Lotti et aI., 1986; Martinez-Chuecos et aI., 1992). Whereas the exposure assessment in the Kaplan series was limited and based almost exclusively on medical history, interpretation of these discrepancies is difficult. 51.3.3.2 Laboratory Findings They are no specific changes in common laboratory tests, including chemical and morphological analysis of spinal fluid. Increased serum levels of immunoglobulin G autoantibodies to glial fibrillary acidic protein and to neurofilament 200 have been detected in a case of methamidophos poisoning after the development of OPIDP (McConnell et aI., 1999), probably reflecting peripheral nerve damage. Lymphocytic NTE NTE activity was found in lymphatic tissues in humans (Moretto and Lotti, 1988) and its characteristics in circulating lymphocytes led to the conclusion that the level of this blood enzyme is similar to that in the nervous system (Bertoncin et aI., 1985). On this basis, suggestions were made to measure and use lymphocytic NTE like RBC AChE activity is used in the clinical setting and in the biomonitoring of occupational exposures (Lotti, 1987). NTE activity has also been detected in humans platelets (Maroni and Bleecker, 1986). Only on one occasion has the ratio between NTE inhibition in lymphocytes and peripheral nerves been measured, and it was found to be about 1 (Osterloh et aI., 1983), although it is anticipated that it will not be always so, given the different pharmacokinetics of OPs. Inhibition of lymphocytic NTE soon after poisoning
51.4 Long-Term Exposures
was predictive of OPIDP development when measured several days before the onset of OPIDP in two cases (McConnell et aI., 1999; Moretto and Lotti, 1998). Given the relatively high turnover of blood lymphocytes and the usually rapid disappearance of OPs from the blood, measurement of lymphocytic NTE should be made in the early days after poisoning because the activity may be back to almost normal at the onset of OPIDP. However, because no treatment for OPIDP is available, the detection of lymphocytic NTE soon after poisoning has limited clinical value. Similarly, measurements of lymphocytic NTE to monitor occupational exposures to OP insecticides have no practical value because such exposures preferentially inhibit blood cholinesterases and, therefore, lymphocytic NTE rarely would be affected (see Section 51.3.1). Electrophysiology Electrophysiological changes are usually detected concurrently with the onset of clinical symptoms and signs of OPIDP. When performed during the symptom-free period between the disappearance of cholinergic toxicity and the clinical onset of OPIDP, the electrophysiological examination is normal (Lotti et aI., 1986). The electrophysiological picture accords well with the histopathological findings of distal axonopathy (Jedrzejowska et aI., 1980; Lotti et aI., 1986; Moretto and Lotti, 1998; Vasilescu and Florescu, 1980; Vasilescu et aI., 1984; see Section 51.3.3.3). The evaluation reveals partial denervation of affected muscles, with increased insertional activity, abnormal spontaneous activity (fibrillation potentials and positive sharp waves), and a reduced interference pattern; large polyphasic motor unit potentials also may be found after a few weeks. The compound muscle action potentials to supramaximal stimulation of motor nerves are reduced in amplitude, and terminal motor latencies are delayed; maximal conduction velocity is usually normal or slightly reduced. Minimal electrophysiological abnormalities of sensory function are occasionally detected. About 1 year after poisoning, normalization of electrophysiological parameters parallels that of clinical signs unless the pyramidal tract is involved. In such a case, findings may resemble those of amyotrophic lateral sclerosis (Vasilescu, 1982). 51.3.3.3 Pathology The histopathology of OPIDP has rarely been described in humans (Aring, 1942; Jedrzejowska et aI., 1980; Lotti et aI., 1986; Vasilescu et aI., 1984), although there are no major differences from what has been extensively observed in experimental animals (Abou-Donia and Lapadula, 1990; Cavanagh, 1973; Tanaka and Bursian, 1989). The central peripheral distribution of lesions is similar to that of toxic neuropathies of other origins. The vulnerability of nerve fibers is directly related to axonal length and diameter; large-diameter and long fibers are more susceptible than small and short ones. Spinal cord changes involve mainly the anterior horn cells and the pyramidal tracts. Lesions in the tract of Goll were less constant and no lesions were seen in the tract of Burdach (Aring, 1942). Sural nerve biopsies indicated axonal-type lesions with an even degree of involvement of myelinated fibers of differ-
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ent sizes and a lesser degree of involvement of unmyelinated fibers. Dark and swollen axoplasm due to the accumulation ofaxoplasmic organelles can be observed association with aspects of axonal degeneration. On teased fiber preparations, some ovoids arranged in linear rows were identified. Electron microscopy found myelin debris in the Schwannian profiles. Depending on the time of biopsy after poisoning, various stages of regeneration and remyelination can be observed. Segmental demyelination is not observed (Jedrzejowska et aI., 1980; Lotti et aI., 1986). These changes indicate a process that is a primary distal axonopathy with moderate, secondary, and distalloss of myelin. 51.3.3.4 Differential Diagnosis The unequivocal suggestion for diagnosis of OPIDP caused by insecticides is the presence of an episode of acute cholinergic toxicity in the recent past medical history. More difficult is differential diagnosis when substantial exposures to nonanticholinesterase OPs that cause OPIDP is overlooked. Symmetricalleg involvement with additional involvement of upper limbs only in severe cases, lack of involvement of cranial nerves and the autonomic system, and electromyographic changes consistent with distal axonal neuropathy are all indicative of OPIDP. Medical history aimed at identification of possible sources of OP exposure remains, in these cases, the only way to etiologically attribute neuropathy. 51.3.3.5 Treatment There is no specific treatment for OPIDP. Intensive programs of physical therapy are indicated to ameliorate muscle trophism during the recovery from peripheral nerve lesions. In later stages, if spasticity develops, GABA antagonists may be used.
51.4 LONG-TERM EXPOSURES Long-term exposures to OPs may cause cholinergic syndrome if both the size of repeated doses and the intervals between them overcome AChE resynthesis in the nervous tissues. In such a case, a buildup of AChE inhibition may occur and when threshold is reached, symptoms that are indistinguishable from those observed after single or short-term exposures are produced. Therefore, only signs and symptoms unrelated to overt cholinergic toxicity will be considered in this section. Many reports on the effects of long-term exposures to OPs lack follow up studies, particularly after cessation of exposures. Therefore, most of the described effects may not be chronic (i.e., longstanding or irreversible) and probably reflect the effects of current exposures. Therefore, it is advisable not to talk about chronic effects, but rather of effects of low-level exposures, either during exposure or shortly afterward, and keep them distinct from effects detectable several months or years after cessation of exposure. Moreover, the major problem of these studies is often the insufficient assessment of exposures, that obviously hampers the interpretation of results.
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51.4.1 NEUROLOGICAL, PSYCHIATRIC, AND BEHAVIORAL EFFECTS The large amount of literature that describes neurological, psychiatric, and behavioral effects has been reviewed in several articles (Brown and Brix, 1998; ECETOC, 1998; Eyer, 1995; Steenland, 1996; Ray, 1998a, b). In one review, these effects were all ranked under the heading of chronic OP-induced neuropsychiatric disorders (COPIND, phenomenon 2; Jamal, 1997). As previously discussed (see Section 51.1.3.9), the various effects will be discussed separately for better comprehension and because there is no evidence that they collectively represent a single nosological entity. 51.4.1.1 Neurological Effects on Central Nervous System Clinical reports A case of parkinsonism was described in a subject with reported past and prolonged exposure to OPs. Apparently he also had several episodes of acute poisoning with parathion and malathion said to have required treatment with oral doses of atropine to control symptoms and signs (Davis et aI., 1978). The past history was not fully reported and it is doubtful that oral atropine would have been effective because it is known to be poorly and unreliably absorbed. Consequently, this case remains an isolated and anecdotal report. A visual syndrome, known as Saku disease, which is characterized by reduced visual field, myopia, astigmatism, lesions of the optic nerve, and abnormal retinal functions, was associated with the extensive use of OPs during the 1960s in one area of Japan (Saku). These effects, exclusively reported by Japanese investigators, have been summarized by PleStina and Piukovic-Plestina (1978) and Dementi (1994). However, the symptoms were not consistant among various OPs to which patients were allegedly exposed and often, but not always, associated with AChE inhibition. This inconsistency raises the question whether the effects are compound-specific and related to RBC AChE inhibition. These results have been criticized and the etiologicallink between OPs and Saku disease remains, for the time being, speCUlative (Erikson-Lamy and Grant, 1992). Veterans who took part in the Persian Gulf War reported higher rates of many symptoms, including neurological ones, and had a decreased perception of well-being (Ismail et aI., 1999; NIH, 1994). Although a single consistent pattern of symptoms and signs is far from being defined (Gray et aI., 1996), this mysterious ailment is now known as the Gulf War syndrome or illness, and several hypotheses have been made concerning causes (Lotti, 1999). One theory states that wartime exposure to a combination of OPs and other cholinesterase inhibiting chemicals synergistically produced the syndrome and the neurological signs in particular, (Haley and Kurt, 1997). Among these chemicals, pyridostigmine bromide was the only defined risk factor (Shen, 1998) because it was given to soldiers who served in the Gulf, apparently for several weeks, as prophylaxis for possible nerve gas attacks. The dosing regime was a 30 mg tablet every 8 hours and it aimed to cause reversible
inhibition of AChE at nerve endings, thereby preventing irreversible inhibition of the enzyme by OP weapons. None of the soldiers ever experienced acute cholinergic symptoms and signs compatible with OP or pyridostigmine poisoning. In a study based on a questionnaire submitted by 249 Gulf War veterans from a single battalion of 606 soldiers, factor analysis of symptoms yielded several syndrome factors (possibly variants of a single syndrome) that suggested various neurological dysfunctions (Haley et aI., 1997a). Subjects with the highest factor scores on syndrome 1 (impaired cognition), syndrome 2 (confusion-ataxia), and syndrome 3 (arthro-myo-neuropathy), for a total of 23 cases, were evaluated for neurological functions and compared with 20 controls from the same battalion, 10 of whom had been deployed in the war region but had no complaints and 10 of whom had not been deployed (Haley et aI., 1997b). Brain dysfunction was evidenced by changes in auditory evoked potentials, interocular asymmetry of nystagmic velocity, asymmetry of saccadic velocity, and somatosensory evoked potentials. However, no clinical differences between cases and controls were detected on neurological examination. Exposures to anticholinesterase chemicals of either cases or controls were not reported. Therefore, whether the Gulf War syndrome exists, whether it affects the nervous system, and whether the clinical findings are due to anticholinesterases cannot be ascertained from these studies. Occupational Exposure Studies Minimal EEG disturbances were reported in a study on 50 workers engaged in the manufacture of a range of unspecified OPs (Metcalf and Holmes, 1969). These changes were not seen in 22 controls and mirrored, to a lesser degree, the more severe disturbances usually seen after acute exposures. Work history and exposure data were insufficient, although it was claimed that the workers were also exposed to chlorinated hydrocarbons. Certain neuropsychological changes were also reported, but it is not clear whether they were associated with such persistent EEG changes. In another study, 32 workers exposed to both OPs and organochlorine (dieldrin) pesticides were subdivided into two equal groups, low and high exposure, based on occupational history. Plasma cholinesterase levels were the same in both groups. EEG and neuropsychological changes were found in the high exposure group (Korsak and Sato, 1977). Quantitative exposure data were not given and EEG changes were different from those reported in the above-mentioned study. A selective effect on the left frontal hemisphere, as derived from EEG and neuropsychological results, was detected. This seems inconsistent with a toxic effect. In a seven country biomonitoring and cross-sectional epidemiological survey of low-dose occupational exposure to OPs, changes in EEG were reported only from some countries. Data from one country showed postseason slow wave activity, whereas data from another country reported different changes (Richter, 1993). These results are difficult to assess given the lack of information on exposure and other confounding factors at the time of testing.
51.4 Long-Term Exposures
51.4.1.2 Neurological Effects on Peripheral Nervous System Clinical Reports In a study on volunteers, mevinphos (25 !1-gkg-1) was administered daily for 28 days to eight subjects, whereas placebo was given to eight controls (Verberk and Salle, 1977). RBC AChE was depressed by 19%, but no correlation was found with the detected changes. At the end of exposure, a 7% decrease in slow fiber motor nerve conduction velocity and a 38% increase in Achilles tendon reflex force were found (as percentages of preexposure values). No effects on neuromuscular transmission were detected. The authors concluded that the significance of such effects with regard to health is unclear. Sensory neuropathies on a series of patients with low-level exposures to chlorpyrifos have been consistently associated with mild or no cholinergic symptoms (Kaplan et aI., 1993) although there is little evidence, if any, for a causal relationship between sensory neuropathy and low-level exposures to chlorpyrifos (see Section 51.3.3.1). In a pilot study, 14 Gulf War veterans were examined for peripheral nerve dysfunction and compared with a control group. Although differences were detected in some parameters (cold threshold, sural nerve latencies, and median nerve sensory action potential), the authors' conclusion was that there may be dysfunction in veterans, but more studies are required (Jamal et aI., 1996). Moreover, the hypothesis that anticholinesterases, other than pyridostigmine, represented a risk factor for veterans of the Gulf War remains to be demonstrated. A clinical study was performed on 72 selected subjects with long-term exposure to sheep dip OPs identified in pilot field studies (Pilkington et aI., 1999). According to defined criteria and neuropathy scores, 23 workers were ranked as having probable/definite neuropathy, 34 workers had possible neuropathy, and 15 workers had no neuropathy. Clinical evaluation, quantitative sensation testing, nerve conduction studies, and electromyography were performed. Results showed neurological signs in 10% of subjects and some small fiber abnormalities in 65% of electrophysiological tests. None of subjects with symptoms was in the no neuropathy subgroup. This neuropathy was thought to be different from classical OPIDP (see Section 51.2), because sensory fiber almost exclusively and small fiber more than large fiber populations were affected. Clinical and electrophysiological assessments were apparently performed at variable times after cessation of peak exposures (up to 1.5 years). Unlike other toxic neuropathies, this apparently new entity also affected upper limbs at early stages, but unlike OPIDP caused by commercial OP pesticides, there was no preceding cholinergic toxicity. Results from this study are difficult to evaluate because they are not presented analytically. Electrophysiological changes were overwhelmingly more frequent than neurological ones, but it is impossible to derive what electrophysiological abnormalities were found in patients with clinically detectable neuropathy. Moreover, the relevance of electrophysiological alterations that are not associated with signs and symptoms is unclear. According to the criteria of
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sample selection, a causal relationship with exposure cannot be inferred; there is also a lack of relationship with the estimated cumulative dose. Finally, we should ask why, after such a long period since cessation of exposure, a very mild neuropathy did not recover. It seems that perhaps the study better represents a validation of a screening system to detect minor electrophysiological signs to be used in the field than a demonstration of a causal relationship between low-level exposure to OPs during sheep dipping and the development of a new form of toxic peripheral sensory neuropathy. Occupational Exposure Studies Neuromuscular function was assessed with surface electrodes on the upper limbs of workers exposed to OPs and organochlorine pesticides (Jager et aI., 1970). A higher incidence of electromyographic changes (repetitive activity and reduced amplitude) was detected in workers exposed to both chemical classes (n = 36) as compared to those exposed to organochlorine only (n = 24) and controls (n = 28). The biological significance of these small changes is unclear, in part because of the use of surface electrodes. There was insufficient information, only statements, concerning exposures. Changes were thought to be related to synaptic dysfunction because they were similar to changes found in myastenic patients who were overtreated with anticholinesterase drugs. However, when observed in such patients, these changes are associated with substantial inhibition of AChE, whereas changes in workers were not associated with whole blood AChE inhibition. Fifty-three workers exposed to both OPs and organochlorine were examined shortly after the start and toward the end of a spraying season (Drenth et aI., 1972). Surface electrode electromyography records of 12 subjects changed from normal to abnormal, whereas those of 13 subjects changed from abnormal to normal. No differences in blood AChE were detected. No evidence of exposure was given. Therefore, the conclusion of the authors that electromyography abnormalities represent only an indication of the need for more protection of workers and no evidence of immediate health problems is not substantiated. Minimal electromyographic changes were detected by surface electrodes in 102 workers exposed to OPs when compared to an unmatched control group of 75 subjects (Roberts, 1976). Fifty-six workers were examined before and after a holiday period. Subjects who displayed these changes somewhat improved after the holidays, whereas some unspecified variability was observed in exposed subjects with normal electromyography. No exposure data were available. In a longitudinal study on six workers exposed to OPs over a 7-9 month period, surface electrode electromyography indicated that voltages varied in a manner that reflected a vague assessment of the pattern of exposure (Roberts, 1977). It is difficult to evaluate these results given the lack of exposure data and the methods used. Neuromuscular function was assessed with surface electrodes in a group of 11 spraymen exposed to OPs (including
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bromophos, diazinon, chlorpyrifos, and malathion) on a recurrent basis (Stalberg et aI., 1978). Plasma cholinesterases were significantly reduced after work, whereas RBC AChE was not. A slight reduction in sensory conduction velocity and increased fiber density was detected in some workers, but was unrelated to lowered plasma cholinesterase activity. Although exposure was not assessed in this study, plasma enzyme inhibition indirectly suggests a lack of correlation between degree of exposure and detected electrophysiological changes. Another study was conducted on four groups of subjects in which approximate exposure was assessed: 42 highly exposed to OP pesticides, 14 seasonal workers exposed to OPs and reexamined after exposure, 129 agricultural workers with low exposure to OPs, and 26 agricultural workers not exposed (Jusic et aI., 1980). The authors concluded that synapse testing with needle electromyography and clinical examination did not detect latent OP intoxication. A study was conducted on workers exposed to the defoliant DEF, where needle electromyography and biochemical studies (lymphocytic NTE and blood cholinesterases) were performed before and after the spraying season (Lotti et aI., 1983). Air and dermal exposure were measured on a typical working day. No electrophysiological changes were detected, although lymphocytic NTE was about 60% inhibited. NTE inhibition after exposure without correlation with electrophysiological changes was explained by the pharmacokinetics of DEF, which requires metabolic activation to be an esterase inhibitor and occurs mainly in the liver; the active metabolite formed is extremely reactive and unless large amounts are formed, it will not reach the nervous system, but will reach the blood, where it inhibits lymphocytic NTE. Twenty-four workers exposed to fenthion were examined with surface electromyography before and after exposure and compared with 19 unexposed controls (Misra et aI., 1988). Serum AChE was also measured. Electrophysiological findings after exposure were no different from controls. However, mean values of some electrophysiological parameters were altered in the exposed group when results obtained during exposure were compared with follow-up data collected 3 weeks after withdrawal from exposure. Also, mean cholinesterase values increased after the end of exposure, but remained within the normal range. Although each exposed individual was his or her own control, results are difficult to interpret because the intraindividual variability of measured parameters of controls was not reported. Two-hundred twenty-nine workers at a pesticide plant were examined clinically (for neurological impairment) and biochemically (lymphocytic NTE and plasma cholinesterase), and tested for tactile sensitivity and motor performance (Otto et aI., 1990). These workers were engaged in the production of a variety of OPs including diazinon, dimethoate, malathion, phentoate, EPN, leptophos, methamidophos, and trichorfon. Results were compared with those obtained from 180 workers from a fertilizer plant and 167 workers from a textile plant. Mean serum cholinesterases and lymphocytic NTE were lower in pesticide workers, although they were within normal ranges.
The proportion of workers with abnormal neurological findings (involuntary tremors and vibration sense) varied between plants. Tactile thresholds in the finger of the nondominant hand were higher in workers in the pesticide plant and the authors stated that this symptom was the most sensitive index of pesticide neurotoxicity. Toes were not tested. No changes were detected in the neurobehavioral tests. Assessment of exposure was missing (although some OPs that potentially cause OPIDP were manufactured), the incidence of various diseases, including neurological ones, was particularly high (in both cases and controls), and the fact that upper limb neuropathy is not expected in mild OPIDP creates problems in interpretating results. An epidemiological study on 90 pesticide applicators (Stokes et aI., 1995) led to the conclusion that prolonged OP exposure is associated with loss of peripheral nerve function. Exposure was assessed by means of urinary excretion of dimethylthiophosphate, one metabolite of azinphos-methyl. However, workers had been exposed to several other OPs and pesticides. The authors' conclusion was based on a significant increase in mean vibration threshold sensitivity for applicators' hands as compared to a matched control group. Feet were not affected. Long-term exposure was determined by questionnaire, but it is unclear whether poisonings had occurred in the past. Subjective symptoms were collected off and on season: headache, weight loss, and nightmares were reported more frequently among pesticide workers, but only headache was statistically increased during the season. Because toxic polyneuropathy does not exclusively affect upper limbs and because workers were exposed to many chemicals, there is no evidence that long-term lowlevel exposures to OPs cause loss of peripheral nerve functions. A cross-sectional study compared 168 spray operators with long-term exposures to OPs with 84 controls (London et aI., 1997). No evidence was found between exposure and loss of vibration sense. However, small differences were found on neurobehavioral test batteries based on information-processing parameters. The authors concluded that there was a small overall evidence of chronic effects of OP exposures, but indicated that exposure misclassification may have contributed to these findings. In another study, the same authors (London et aI., 1998) investigated neurological symptoms, vibration sense, and tremors in much the same population during the peak spraying season. Eighty-three nonspraying workers were used as the control group. Exposure, as in the previous study, was derived from a job-exposure matrix for pesticides in agriculture. Applicators significantly reported more dizziness, sleepiness, and headache, and had a higher overall neurological symptom score. Vibration sense and tremor outcome were not associated with past longterm OP exposure. A correlation was found between symptoms and either current exposure or episodes of past OP poisoning (see also Section 51.1.3.9). The effects of low-level exposure to foliar OP residues (primarily to azinphos methyl and possibly to phosmet and methylparathion) during one season were assessed in a cross-sectional study on 67 workers and 68 matched controls (Engel et at., 1998). Sensory and motor nerve conduction velocities, neuro-
51.4 Long-Term Exposures
muscular junction testing, and RBC AChE were measured. No differences were found between exposed and controls. 51.4.1.3 Psychiatric Effects Clinical Reports Schizophrenic and depressive reactions, with severe impairment of memory and difficulty in concentration were reported in 16 workers after variable exposures to OPs (Gershon and Shaw, 1961). This report was criticized because of serious flaws, including the lack of evidence for exposure, the detailed clinical description of only a few cases, and the inconsistency with larger studies (Barnes, 1961; Bidstrup, 1961). An anecdotal report suggested a causal relationship between psychiatric disturbances and exposure to a variety of pesticides including OPs in two pilots (Dille and Smith, 1964). Another anecdotal report linked the onset of psychosis in a farmer with previous spraying of demeton-S-methyl, but no casual relationship was established (Bradwell, 1994). Geographical Studies A geographical study was carried out to determine whether areas of high OP usage in Australia had a higher proportion of admissions for psychiatric disorders than low-usage areas (Stoller et aI., 1965). No evidence was found that schizophrenia, depressive states, psychoneuroses, or personality disorders were more common in high-usage areas than elsewhere. Increased risk of suicide was associated to pesticide exposure (mainly OPs) in an agricultural area (Parr6n et aI., 1996a). The rate of suicides was compared with rates where exposure to pesticides was low. Most suicide cases involved pesticides, but other factors that influenced suicide attitudes were not analyzed. Occupational Exposure Studies Workers who had unspecified exposures to OPs were compared with a control group on personality tests, structured interview, and cholinesterase levels (Levin et aI., 1976). Commercial sprayers, but not farmers, showed higher levels of anxiety and lower plasma cholinesterase than controls. The authors concluded that these findings were tentative until confirmed by additional studies. The effect of exposure stress in the absence of exposure was reported during a manufacturing accident with malathion (Markowitz et aI., 1986). The reactions of allegedly exposed workers were compared with a matched group. The exposed group showed high demoralization scores, particularly among those who admitted to little knowledge about toxic chemicals. Twenty-five greenhouse workers were compared to controls and showed a higher incidence of symptoms of depression and tremors (Parr6n et aI., 1996b). Exposure was not measured and blood cholinesterases were normal. Two groups, one of pesticide formulators (208 individuals) and another of applicators (172 individuals), were compared with matched controls (72 and 151 individuals, respectively; Amr et aI., 1997). Exposures to a variety of pesticides, including OPs, carbamates, organochlorine, and pyrethroids, were not
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quantified. Both exposed groups had a higher incidence of total psychiatric disorders, whereas formulators had a higher incidence of depressive neurosis that was related to the duration of employment. It is difficult to assess the role, if any, of the OPs, given the variety of pesticides to which these workers were exposed. A case control study investigated the link between exposure to pesticides and suicide in Canadian farmers (Pickett et al., 1998). Results excluded exposure to pesticides as an important risk factor for suicide among farmers. However, the chemicals used were not identified and were only divided between herbicides and insecticides. The latter certainly included OPs. Therefore, it cannot be ascertained from this study whether exposures to OPs were involved. 51.4.1.4 Neurohehavioral Effects The neurobehavioral effects of long-term exposures to low levels of OPs have been extensively reviewed over the last few years (D'Mello, 1993; ECETOC, 1998; Eyer, 1995; Jamal, 1995; Mearns et aI., 1994; Ray, 1998a, b; Steenland, 1996). Although much information has been published, results are contradictory and whether such exposures are linked with an increased risk of behavioral effects in humans is controversial. A study compared two groups of 53 and 68 asymptomatic workers with varying degrees of unquantified and unspecified exposure to OPs to controls (25 and 22 subjects, respectively) on a complex reaction time test. Results showed there was no indication that exposure at levels insufficient to produce clinical illness had any important effect on mental alertness (Durham et aI., 1965). Another study selected 23 workers who regularly used OPs and had used them within 2 weeks of the testing date (Rodnitzky et aI., 1975). Recent exposure was confirmed by lower plasma cholinesterase, but RBC AChE was normal. Types of pesticides were not reported. The results of tests for memory, signal processing, vigilance, language, and proprioceptive performance were no different from those of a matched control group. Neurobehavioral tests were performed before and after work shifts on 99 pest control workers with low-level, short-term exposure to diazinon (Maizlish et aI., 1987). Exposure was measured by means of the urinary metabolite diethylthiophosphate before and after the end of shifts. No changes in neurobehavioral functions were detected on a battery of seven tests. Similarly, no changes were seen when workers were subdivided according to the degree of exposure. Neuropsychological performance was assessed by test battery in a group of 49 pesticide applicators prior to and I month after the end of a 6-month pesticide spraying season. Results were compared with 40 controls (Daniell et a/., 1992). The nature and extent of pesticide exposure were assessed and reported in another paper (Karr et al., 1992). The comparison of seasonal RBC AChE changes according to exposure levels showed lower cholinesterase among higher exposure groups compared with lesser exposure group. No evidence of sig-
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Clinical Toxicology of Anticholinesterase Agents
nificant decrements in neuropsychological performance was reported. The neurobehavioral status was assessed in workers and kibbutz residents differently exposed to OPs and other pesticides (Richter et aI., 1992). Subjects were examined during the spraying season and afterward. Most neurobehavioral scores were poorer during the season. Exposure data were not reported and the authors drew attention to other risk factors such as work load and heat stress. In a cross-sectional study, neuropsychological performance in 146 sheep farmers was compared to 143 quarry workers (Stephens et aI., 1995a, b). The selection and the testing procedures for workers who belonged to the experimental group were different from those of controls. Long-term exposure data were assessed by means of a retrospective exposure questionnaire that used the number of sheep, dips, and years of employment as a surrogate. Farmers performed significantly worse than controls in tests to assess sustained attention and speed of information processing (simple reaction time, symbol digit substitution, and syntactic reasoning). A dose-response relationship was found only for one test (syntactic reasoning). Moreover, in another article, no association was found between the experience of acute symptoms and performance on neuropsychological tests (assessed on a subset of workers), and it was concluded that neuropsychological data reflect chronic effects that occur independently of acute effects (Stephens et al., 1996). Given the large differences between OP and quarry workers, it is doubtful that the small changes detected in the former should be attributed to low-level long-term exposures to OPs. Fifty-seven licensed applicators were compared on several neuropsychological tests to a control group of 34 farmers who had no exposure to pesticides (Fiedler et aI., 1997). Exposure to OPs was assessed with a questionnnaire on work history, but details were not given. None of the applicators had episodes of acute poisoning, and RBC AChE values were normal. Except for slower reaction time, no other difference in neuropsychological performances was detected between exposed and nonexposed subjects. Subclinical morbidity patterns, including symptoms, aiming at digit symbol tests, and measurement of RBC AChE, were investigated in 226 established farm workers and were compared with an equal number of controls and with 92 new farm workers (Gomes et al., 1998). Results indicated a higher incidence of symptoms (irritated conjunctiva, watery eyes, blurred vision, dizziness, headache, and muscular pain and weakness), reduced performance on the aiming and digit symbol tests, and reduced AChE activity in the group of established farmworkers. Although RBC AChE inhibition implies OP exposure, no actual exposure data were reported. Moreover, reduction of AChE was still within the coefficient of variation of the test. Nevertheless, because the above-reported symptoms are consistent with cholinergic overstimulation, it is likely that differences between groups reflect the effects of current exposures to OPs.
51.4.2 OTHER EFFECTS
Several toxic effects have been associated with long-term exposures to OPs. However, most of them are either isolated reports or are based on circumstantial evidence of exposure and probably are simply coincidental. A case report, for instance, suggested congestive cardiomyopathy caused by long-term exposure to OPs without signs of severe acute poisoning. Evidence of exposure was limited and the patient had a previous myocardial infarction, therefore, cardiomyopathy was likely a consequence of myocardial infarction rather than OP exposure (Fazekas and Kiss, 1980). Scveral cases of influenza-like symptoms associated with OP use in farmers during the sheep dipping season apparently were unrelated to RBC AChE inhibition (Murray et aI., 1992). Contact dermatitis and asthma have been linked to exposures to OP pesticides (Bryant, 1985; Deschamps et al., 1994; Xue, 1992). Whether the effects are due to sensitization or irritation, or if they are linked to other ingredients always present in commercial formulations of pesticides is unclear. Immunotoxicity of OPs has been suggested, although human data are mostly based on in vitro studies (Newcombe and Esa, 1992; Rodgers et aI., 1992; Sharma and Tomar, 1992). Hypotheses propose that exposures to OPs are linked to cancer development (Newcombe, 1992). One study was performed on workers engaged in the production of several OPs (trichlorfon, chlorfenvinphos, malathion, dichlorvos, fenitrothion, and formothion; Hermanowicz and Kossman, 1984). Exposure involved other chemicals and its assessment was rather approximate. RBC AChE and plasma ChE were reported on a group basis and correlated with the assessment of exposure. A marked impairment in neutrophil chemotaxis was found in workers who were likely to be exposed to OPs. The frequency of upper respiratory tract infections was higher in the exposed group compared with controls. No other types of infections showed increased frequency. The authors themselves concluded that a distinction cannot be made between OPs and other chemicals as possible factors for the described effects. Certainly more information is needed to ascertain whether immunotoxicity is an effect of OPs and what pathophysiological significance should be attributed to it. There is no evidence so far that any form of immunomediated clinical effect is linked to OP exposures. 51.4.3 BIOMONITORING OCCUPATIONAL EXPOSURES
Long-term occupational exposures to OPs are commonly monitored by measuring either urinary excretion of alkylphosphates or blood cholinesterases. The goal is to prevent adverse effects from OPs. 51.4.3.1 Assessment of Urinary Alkylpbospbates
Although OPs may be excreted unchanged, they are usually hydrolyzed, and the acidic and alcoholic moieties can be found in the urine of exposed subjects. Measurement of metabolites
51.4 Long-Term Exposures
is common practice in workers exposed to OPs, and several alkylphosphates have been identified, including dimethylphosphates, dimethy lthiophosphates, dimethy ldithiophosphates, and dimethylphosphorothioates derived from dimethylated OPs and the corresponding metabolites derived from diethylated OPs (Coye et aI., 1986). Several gas-chromatographic methods to measure urinary dialkylphosphates have been developed (Aprea et al., 1996; Nutley and Cocker, 1993). Measurement of excretion of the alcoholic moiety in exposed workers has been used less frequently. Examples include the measurement of 3,5,6-trichloropyridinol after exposure to chlorpyrifos and chlorpyrifos-methyl (Nolan et al., 1984), p-nitrophenol after exposure to parathion and parathion-methyl (Morgan et al., 1977), and mono- and dicarboxylic acid after exposure to malathion (Bradway and Shafik, 1977). Despite the numerous field studies where exposures have been assessed by means of urinary metabolites, not many data are suitable for a toxicological interpretation. The reasons are many. The first problem is related to usual agricultural practices, which lead to concurrent exposures to several OPs. As stated earlier, different OPs, each with its own toxicity, may be metabolized, yielding the same product. It is, therefore, difficult to assess the toxicological risk associated with such exposures. For instance, certain exposures to either parathionmethyl or chlorpyrifos-methyl caused comparable excretion of dimethylphosphates and dimethylthiophosphates. However, the risk derived from each OP is quite different, because they display a 3 orders of magnitude difference in acute toxicity (Moretto et aI., 1995). Depending on the OP, route of exposure, metabolism, and distribution, peak metabolite excretion might be reached at different times after the end of exposure. For instance, certain compounds such as chlorpyrifos show peak urinary excretion of ethylphosphates several hours after the end of exposure (Fenske and Elkner, 1990; Moretto et al., 1995). On the contrary, peak excretion of ethylphosphates derived from exposures to parathion occurs within a much shorter time (Morgan et aI., 1977). Moreover, alkylphosphates may have a different time course of excretion. For instance, diethylphosphate peaks earlier than diethylphosphorothioate after exposure to diazinon (Sewell et al., 1999). Therefore, the timing of urine sampling is crucial to assess the significance of a given concentration; for many OPs, the relevant timing information is missing. Finally, very little is known about the correlation between urinary metabolite excretion and the inhibition of AChE and/or plasma BuChE. In many cases, enzyme inhibition was not found (Griffin et aI., 1999; Kraus et aI., 1977, 1981; Krieger and Thongsinthusak, 1993; Popendorf et aI., 1979); in a few cases, minimal inhibition was found (Jauhiainen et aI., 1992; Spear et aI., 1977). Only at a time when occupational exposures were probably much more severe than nowadays, was a good correlation found between p-nitrophenol and RBC AChE inhibition in parathion exposed workers (Arterberry et aI., 1961). In conclusion, for the time being, it is difficult to give these data a toxicological significance beyond that of being a qualitative exposure index.
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51.4.3.2 Monitoring Blood Cholinesterase Blood cholinesterase activities have been used extensively to monitor the effects of occupational exposures to OPs. Guidelines have been developed on methods, interpretation of results, and actions to be taken (EPA, 1992; Plestina, 1984; WHO, 1986). However, these suggestion should be taken as general indications, particularly when interpreting single data, because the following issues must be considered (Lotti, 1995). Relationship between Inhibition of Blood Enzymes and Cholinergic Toxicity The postulate for using blood cholinesterases to biomonitor occupational exposures to OPs is that inhibition of these enzymes reflects either or both the degree of exposure or the corresponding enzyme inhibition in the nervous tissues. Because no physiological functions have been attributed to BuChE (confirmed by the fact that homozygote carriers of defective BuChE are healthy subjects; see succeeding text), the inhibition of this enzyme in any tissue most likely has no significance in terms of health. However, its inhibition in plasma means that exposure has occurred. This statement may not be true in the case of diseases (unlikely in occupational exposures) that cause depression of plasma BuChE, such as parenchimalliver diseases, acute infections, some anemias, and malnutrition. In this respect, an interesting observation is that patients with liver diseases not only have low plasma cholinesterases, but also may show a further reduction as a result of a level of exposure to OPs that causes no change of enzyme activity in normal persons or in persons affected by other diseases (Hayes, 1982). A possible explanation is the reduced ability of patients affected by liver diseases to detoxify certain OPs. Obviously, if BuChE inhibition is associated with inhibition of RBC AChE, then different conclusions should be drawn. How accurately inhibition of RBC AChE reflects that in the synapses is unknown and extrapolation is difficult, given the different access various OPs have to the blood and the nervous system. Animal data suggest that sometimes inhibition is similar, more often, inhibition of blood enzyme is higher (Su et aI., 1971) due to the particular protection of the nervous system that is offered by the blood-brain barrier. After recovery from exposure, extrapolation is even more difficult, given the different rates of recovery of AChE in the RBCs and nervous tissues, respectively. Nevertheless, based on clinical and toxicological data, a rough estimate of the levels of RBC AChE inhibition that require action are reported in the Table 51.10. It is also clear that because the access of xenobiotics to blood is always easier than to brain and because no evidence exists that OPs accumulate in the nervous system, the inhibition of RBC AChE usually overestimates the level in the brain. Spontaneous Reactivation and Reappearance of Blood Enzymes When measuring blood cholinesterase for biomonitoring purposes, the rates of reappearance of activities after inhibition should be taken into account. Such rates depend on spontaneous reactivation (in the case of carbamates and
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CHAPTER 51
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Table 51.10 Relationship between RBC AChE Inhibition and Preventive Actions When Monitoring Occupational Exposures to opsa
Table 51.11 Sensitivity of Plasma BuChE and RBC AChE in Humans to Various Insecticides a
RBCAChE
RBC AChE most inhibited
(% inhibition from
Dimefox (Edson, 1964)
Chlorfenvinphos (FAOIWHO, 1995)
Mevinphos (Rider et aI., 1972, 1975)
Chlopyrifos (Eliason et aI., 1969)
Improve hygenic conditions
Methyl-parathion (Rider et aI., 1970)
Demeton (Moeller and Rider, 1965)
As above plus removal of
Parathion (Hayes, 1982)
preexposure values)
Significance
Preventive action
20-29
Evidence of exposure
30-50
Hazard
subject from exposure >50
Poisoning
Plasma cholinesterase most inhibited
Diazinon (FAOIWHO, 1967) Dichlorvos (Rasmussen et aI., 1963)
Admit subject to the hospital
Fenitrothion (Vandekar, 1965) Malathion (Elliot and Barnes, 1963)
aData from WHO (1986) and Lotti (1995).
Monocrotophos (FAOIWHO, 1996) Trichlorfon (Abdel-Al et aI., 1970)
dimethylphosphates) and resynthesis of new enzymes. As previously discussed (Section 51.1.2), the rate constants of spontaneous reactivation and of aging vary according to the phosphorylating agent. Because rate constants also depend on the enzyme that is phosphorylated, they will be different when measured in RBC AChE or plasma BuChE (Aldridge and Reiner, 1972). However, the few studies in humans do not necessarily confirm these theoretical considerations. Data from school children treated orally with trichlorfon against schistosomiasis showed that plasma BuChE as well as RBC AChE recovered much slower than was predicted from in vitro spontanous reactivation studies (Reiner and Ple§tina, 1979). The synthesis of AChE occurs in the bone marrow, and its presence in the blood depends on the normal turnover of RBCs (i.e., 120 days). The synthesis of BuChE occurs in the liver and its turnover in plasma corresponds to about 20 days. Resynthesis of both enzymes after irreversible inhibition by OPs in the nervous systems of animals seems to occur at similar rates, corresponding to a half-life of 5-7 days. However, resynthesis is reflected in the blood quite differently because the localizations of AChE and BuChE differ. Thus, the reappearance of RBC AChE has been shown to occur at a rate of about 1% per day, whereas the rate of plasma BuChE is about 5% per day (Hayes, 1982). Sensitivity of Blood Enzymes to Inhibitors As stated previously, RBC AChE and plasma BuChE are different enzymes, and, therefore, they display different substrate specificity. Whereas interactions of OPs with esterases are analogous to interactions of substrates with esterases, it is clear that blood cholinesterases are differently inhibited by a given ~P. As shown in Table 51.11, plasma BuChE is generally more sensitive to inhibition than RBC AChE by most OPs used as insecticides. In cases of mild exposures, plasma BuChE may be the only inhibited enzyme. This observation should be interpreted as a sign of exposure, but not of poisoning. In cases of severe exposures, profound inhibition of plasma BuChE is always associated with similar inhibition of RBC AChE as it occurs in poisoned patients. The situation may be more complex when substantial repeated exposures are involved. In such a case, even if the plasma enzyme is more sensitive, a buildup of RBC
a Circumstances
of exposure vary.
AChE inhibition can occur, thus equalizing the inhibition of both enzymes at a certain time because of the different rates of reappearance of the two enzymes. Therefore, equal inhibition of the enzymes may represent the consequence of either a single substantial exposure or less severe but repeated exposures. Finally, RBC AChE may be more inhibited, even if the plasma enzyme is more sensitive when subjects are recovering from substantial exposures, given the prolonged life of the RBCs that carry inhibited AChE. Inter- and Intraindividual Variability of Blood Enzymes Intraindividual variability of both plasma BuChE and RBC AChE is high. Samples taken at intervals ranging from a few days to several years indicate that the coefficients of variation of both enzymes in unexposed subjects vary from 7 to 11 %. In some cases, an intraindividual variability of plasma enzyme up to 100% was detected in the course of several months (Hayes, 1982). Interindividual variability of these enzymes is even greater. The coefficients of variation of RBC AChE in unexposed subjects vary from 10 to 40%, whereas the corresponding values for plasma enzyme vary from 12 to 46%. Minor differences exist according to gender, age, and race (Hayes, 1982). A few people who have normal levels of RBC AChE are genetically deprived of plasma BuChE (0stergaard et al., 1992). It was observed that some of these people who were treated with succinylcholine during surgery exhibited an abnormal prolonged period of muscular paralysis after usual dosages of the drug. These patients were found to have a plasma BuChE much lower than normal (Bourne et al., 1952). Moreover, the enzyme is also qualitatively different from the norm, as for instance, in sensitivity to inhibitors. Because of this, a test was developed based upon lesser inhibition of the enzyme by dibucaine (Kalow and Staron, 1957). The dibucaine number (degree of plasma cholinesterase inhibition by dibucaine) discriminates three phenotypes: normal, intermediate, and atypical. The approximate frequency of these phenotypes has been estimated at 96, 3.9, and 0.05%, respectively (Harris and Whittaker, 1962). Other
References
tests are available to discriminate abnormal BuChE, based upon fluoride number (Harris and Whittaker, 1961), chloride number (Whittaker, 1968), scoline number (King and Griffin, 1973), and urea number (Hanel and Viby-Mogensen, 1977; see Section 1.3.6.4).
51.4.3.3 Detection of Hypersusceptible SUbjects Whereas OPs are inhibitors of plasma BuChE and are largely hydrolyzed by A-esterases, such as paraoxonase (PONI) and other carboxylesterases (aliesterases), inherited or acquired deficits of scavenger (plasma BuChE) or detoxifying (esterases) abilities have been suggested as potential factors for increased susceptibility to OPs (Loewenstein-Lichtenstein et aI., 1996; Saxena et aI., 1997). Although there is some evidence in experimental animals for hypersusceptibility based upon reduced ability to detoxify OPs, only one example is known in humans. An unexpected outbreak of malathion poisoning arose in workers occupationally exposed to commercial brands of malathion that contained high amounts of impurities. Among these, isomalathion was the most relevant because it inhibited the carboxylesterases that inactivativate malathion by hydrolyzing its carboxyl-ester linkages (Baker et aI., 1978). Recurrent suggestions have been made that genetically determined low levels of plasma BuChE increased the susceptibility to OP toxicity because of a reduced scavenger capability. However, a relationship between abnormal plasma BuChE and hypersusceptibility to OP poisoning has never been reported. Hypersusceptibility to succinylcholine would never have been discovered if some unusual people had not undergone succinylcholine treatment, an event that is probably no more common than heavy exposure to OPs. Moreover, when the biochemical characteristics of the normal and of the genetically determined defective enzymes were compared stoichiometrically with the plasma concentrations of inhibitors, no scavenger functions to plasma BuChE could be detected (Lotti and Moretto, 1995). Based on the polymorphism of PON 1 in human populations and the known role of this enzyme in the detoxification of some OPs (Davies et al., 1996; Mueller et aI., 1983), it has been inferred that the expression of this enzyme is involved in determining hypersusceptibility to OPs (Mackness et aI., 1998). However, so far, proof for this has been obtained only in animals (Costa et aI., 1999).
REFERENCES Abdel-AI, A. M. A., EI-Hawary, M. F. S., Kamel, H., Abdel-Khalek, M. K., and EI-Diwany, K. M. (1970). Blood cholinesterases, hepatic, renal and haemopoietic functions in children receiving repeated doses of "Dipterex." l. Egypt. Med. Assoe. 53,265-271. Abend, Y., Goland, S., Evron, E., Sthoeger, Z. M., and Geltner, D. (1994). Acute renal failure complicating organophosphate intoxication. Renal. Fail. 16, 415-417. Abou-Donia, M. B., and Lapadula, D. M. (1990). Mechanisms of organophosphorus ester-induced delayed neurotoxicity: Type I and type 11. Ann. Rev. Pharmaco!' Toxicol. 30, 405-440.
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Adams, R. G., Verma, P., lackson, A. l., and Miller, R. L. (1982). Plasma pharmacokinetics of intravenously administered atropine in normal human subjects. l. Clin. Pharmacol. 22,477-481. Ahlgren, l. D., Manz, H. l., and Harvey, 1. C. (1979). Myopathy of chronic organophosphate poisoning: A clinical entity? S. Afr. Med. l. 72,555-563. Aiuto, L. A., Pavlakis, S. G., and Boxer, R. A. (1993). Life-threatening organophosphate-induced delayed polyneuropathy in a child after accidental chlorpyrifos ingestion. l. Pediatr. 122, 658-660. Aldridge, W. N., and Reiner, E. (1972). "Enzyme Inhibitors as Substrates." North-Holland, Amsterdam. Alkondon, M., and Albuquerque, E. X. (1989). The nonoximc bispyridinium compound SAD- I 28 alters the kinetic properties of the nicotinic acetylcholine receptor ion channel: A possible mechanism for antidotal effects. l. Pharmacal. Exp. Ther. 250, 842-852. American EIectroencephalographic Society (1987). Statement on the clinical use of quantitative EEG. l. Clin. Neurophysiol. 4, 75. Ames, R. G., Steenland, K., Jenkins, B., ChrisIip, D., and Russo, J. (1995). Chronic neurologic sequelae to cholinesterase inhibition among agricultural pesticide applicators. Arch. Environ. Health 50, 440-443. Amr, M. M., Halim, Z. S., and Moussa, S. S. (1997). Psychiatric disorders among Egyptian pesticide applicators and formulators. Environ. Res. 73, 193-199. Aprea, c., Sciarra, G., and Lunghini, L. (1996). Analytical method for the determination of urinary alkylphosphates in subjects occupationally exposed to organophosphorus insecticides and in the general population. l. Anal. Taxica!. 20,559-563. Aring, C. D. (1942). The systemic nervous affinity of triorthocresyl phosphate (Jamaica Ginger Palsy). Brain 65, 34-47. Arterberry, J. D., Durham, W. F., Elliot, J. w., and Wolfe, H. R. (1961). Exposure to parathion: Measurement by blood cholinesterase level and urinary p-nitrophenol excretion. Arch. Environ. Health 3, 112-121. Arterberry, l. D., Bonifaci, R. w., Nash, E. w., and Quinby, G. E. (1962). Potentiation of phosphorus insecticides by phenothiazine derivatives. Possible hazard, with report of a fatal case. lAMA 182, 848-850. Baker, D. 1., and Sedgwick, E. M. (1996). Single fibre electromyographic changes in man after organophosphate exposure. Human Exp. Toxiea!. 15, 369-375. Baker, E. L., Zack, M., Miles, l. W., Alderman, L., Warren, M., Dobbin, R. D., Miller, S., and Teeters, W. R. (1978). Epidemic malathion poisoning in Pakistan malaria workers. Lancet i, 31-34. Balali-Mood, M., and Shariat, M. (1998). Treatment of organophosphate poisoning. Experience of nerve agents and acute pesticide poisoning on the effects of oximes. l. Physio!. Paris 92, 375-378. Bardin, P. G., and van Eeden, S. F. (1990). Organophosphate poisoning: Grading the severity and comparing treatment between atropine and glycopyrrolate. Crit. Care Med. 18, 956-960. Bardin, P. G., van Eeden, S. F., and Joubert, J. R. (1987). Intensive care management of acute organophosphate poisoning. A 7-year experience in the western Cape. S. Ajr. Med. J 72,593-597. Bardin, P. G., van Eden, S. F., Moolman, l. A., Foden, A. P., and Joubcrt, J. R. (1994). Organophosphate and carbamate poisoning. Arch. Intern. Med. 154, 1433-1441. Bames, J. M. (1961). Psychiatric sequelae of chronic exposure to organophosphorus insecticides. Lancet ii, 102-103 Barr, A. M. (1966). Further experience in the treatment of severe organic phosphate poisoning. Med. l. Aust. 1, 490-492. Benson, B. l., Tolo, D., and McIntire, M. (1992). Is the intermediate syndrome in organophosphate poisoning the result of insufficient oxime therapy? Clin. Toxieal. 30, 347-349. Bentur, Y., Nutenko, 1., Tsipiniuk A., Raikhlin-Eisenkraft, B., and Taitelman, U. (1993). Pharmacokinetics of obidoxime in organophosphate poisoning associated with renal failure. Clin. Toxicol. 31, 315-322. Bertolazzi, M., Caroldi, S., Moretto, A., and Lotti, M. (1991). Interaction of methamidophos with hen and human acetylcholinesterase and neuropathy target esterase. Arch. Toxico!. 65, 580-585. Bertoncin, D., Russolo, A., Caroldi, S., and Lotti, M. (1985). Neuropathy target esterase in human Iymphocytes. Arch. Environ. Health 40, \39-144.
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CHAPTER 51
Clinical Toxicology of Anticholinesterase Agents
Sundwall, A. (1961). Minimum concentrations of N-methylpyridinium-2aldoxime methane sulphonate (P2S) which reverse neuromuscular block. Biochem. Phannacal. 8,413-417. Susser, M., and Stein, Z. (1957). An outbreak of tri-ortho-cresyl phosphate (T. O. C. P.) poisoning in Durban. Br. J. Indust. Med. 14, 111-120. Sussman, J. L., Harel, M., Frolow, E, Oefner, c., Goldman, A., Toker, L., and Silman, I. (1991). Atomic structure of acetylcholinesterase from Torpedo californica: A prototypic acetylcholine-binding protein. Science 253, 872879. Suzuki, T., Morita, H., and Ono, K. (1995). Sarin poisoning in Tokyo subway. Lancet 345, 980. Tabershaw, I. R., and Cooper, W. C. (1966). Sequelae of acute organic phosphate poisoning. J. Occup. Med. 8, 5-20. Tafuri, J., and Roberts, .T. (1987). Organophosphate poisoning. Ann. Fmerg. Med. 16, 193-202. Tanaka, D., Jr., and Bursian, S. J. (1989). Degeneration patterns in the chicken central nervous system induced by ingestion of the organophosphorus delayed neurotoxin tri-ortho-tolyl phosphate. A silver impregnation study. Brain Res. 484, 240-256. Taylor, P. (1996a). Anticholinesterase agents. In "Goodman and Gilman's the Pharmacological Basis of Therapeutics" (J. G. Hardman and L. E. Limbird, eds.), 9th ed., pp. 161-176. McGraw-Hill, New York. Taylor, P. (1996b). Agents acting at the neuromuscular junction and autonomic ganglia. In "Goodman and Gilman's the Pharmacological Basis of Therapeutics" (1. G. Hardman and L. E. Limbird, eds.), 9th ed., pp. 177-197. McGraw-Hill, New York. Thompson, J. W, and Stocks, R. M. (1997). Brief bilateral vocal cord paralysis after insecticide poisoning. A new variant of toxcity syndrome. Arch. Otalaryngal. Head Neck Surg. 123, 93-96. Tomlin, C. D. S. (1997). "The Pesticide Manual," llth ed. British Crop Protection Council, Surrey, UK. Tosi, L., Righetti, C., Adami, L., and Zanette, G. (1994). October 1942: A strange epidemic paralysis in Saval, Verona, Italy. Revision and diagnosis 50 years later of tri-ortho-cresyl phosphate poisoning. J. Neural. Neurosurg. Psychiatry 57, 810-813. Tracey, J. A., and Gallagher, H. (1990). Use of glycopyrrolate and atropine in acute organophosphorus poisoning. Human Exp. Taxicol. 9,99-100. Tsao, T. C-Y., Juang, y-c., Lan, R.-S., Shieh, W-B., and Lee, C.-H. (1990). Respiratory failure of acute organophosphate carbamate poisoning. Chest 98,631-636. Tsatsakis, A. M., Aguridakis, P., Michalodimitrakis, M. N., Tsakalov, A. K., Alegakis, A K., Koumantakis, E., and Troulakis, G. (1996). Experiences with acute organophosphate poisonings in Crete. Vet. Human Taxcial. 38, 101-107. Tush, G. M., and Amstead, M.1. (1997). Pralidoxime continuous infusion in the treatment of organophosphate poisoning. Ann. Pharmacather. 31,441-444. Vale, J. A, and Scott, G. W. (1974). Organophosphorus poisoning. Guy's Hasp. Rep. 123, 13-25. Valero, A, and Golan, D. (1967). Accidental organic phosphorus poisoning: The use of propranolol to counteract vagolytic cardiac effects of atropine. Isr. J. Med. Sci. 3, 582-584. Vandekar, M. (1965). "Observations of the Toxicity of Two Organophosphorus and One Carbamate Insecticide in a Village Trial Performed by WHO Insecticide Testing Unit in Lagos During 1964", WHO Work. Doc. 65fToxl2.64, U.S. Govt. Printing Office, Washington, DC. Van Meter, W G., Karczmar, A. G., and Fiscus, R. R. (1978). CNS effects of anticholinesterases in the presence of inhibited cholinesterases. Arch. Int. Pharmacadyn.231,249-260. Vasilescu, C. (1982). Neuropathy after organophosphorus compounds poisoning. J. Neural. Neurosurg. Psychiatry 45, 942. Vasilescu, C., and Florescu, A (1980). Clinical and electrophysiological study of neuropathy after organophosphorus compounds poisoning. Arch. Taxicol. 43,305-315. Vasilescu, c., Alexianu, M., and Dan, A. (1984). Delayed neuropathy after organophosphorus insecticide (dipterex) poisoning: A clinical, electrophysiological and nerve biopsy study. J. Neural. Neurasurg. Psychiatry 47, 543-548.
Verberk, M. M., and Salle, H. J. A (1977). Effects of nervous function in volunteers ingesting mevinphos for one month. Taxicol. Appl. Pharmacal. 42, 351-358. Verpooten, G. A., and De Broe, M. E. (1984). Combined hemoperfusionhemodialysis in severe poisoning: Kinetics of drug extraction. Resuscitation 11,275-289. Wadia, R. S., Sadagopan, C., Amin, R. B., and Sardesai, H. V. (1974). Neurological manifestations of organophosphorus insecticide poisoning. J. Neural. Neurasurg. Psychiatry 37, 841-847. Wadia, R. S., Shinde, S. N., and Vaidya, S. (1985). Delayed neurotoxicity after an episode of poisoning with dichlorvos. Neural. India 33, 247-253. Wadia, R. S., Chitra, S., Amin, R. B., Kiwalkar, R. S., and Sardesai, H. v. (1987). Electrophysiological studies in acute organophosphate poisoning. J. Neural. Neurosurg. Psychiatry 50, 1442-1448. Wang, A-G., Liu, R.-S., Liu, J.-H., Teng, M. M.-H., and Yen, M. Y. (1999). Positron emission tomography scan in cortical visual loss in patients with organophosphate intoxication. Ophthalmology 106, 1287-1291. Wecker, L., Mrak, R. E., and Dettbam, W. D. (1985). Evidence of necrosis in human intercostal muscle following inhalation of an organophosphate insecticide. J. Enviran. Pathol. Taxicol. Oncal. 6, 171-175. Weeks, D. B., and Ford, D. (1989). Prolonged suxamethonium-induced neuromuscular block associated with organophophate poisoning. Br. J. Anaesth. 62,327. Weir, S., Minton, N., and Murray, V. (1992). Organophosphate poisoning in the U.K.: The National Poisons Information Service experience during 1984-1987. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B. Ballantyne and T. C. Marrs, eds.), pp. 463-470. Butterworth-Heinemann, Oxford. Weizman, Z., and Sofer, S. (1992). Acute pancreatitis in children with anticholinesterase insecticide intoxication. Pediatrics 90, 204-206. Whittaker, M. (1968). The pseudocholinesterase variants. Differentiation by means of sodium chloride. Acta Genet. 18, 566-562. WHO (1986). "Organophosphorus Insecticides: A General Introduction." Environmental Health Criteria 63, World Health Organization, Geneva. WHO (1998). "The WHO Recommended Classification of Pesticides by Hazard and Guidelines to Classification 1998-1999." WHO/PCS/98.21, World Health Organization, Geneva. Whorton, M. D., and Obrinsky, D. L. (1983). Persistence of symptoms after mild to moderate acute organophosphate poisoning among 19 farm field workers. J. Taxicol. Environ. Health 11,347-354. Willems, J. L. (1981). Poisoning by organophosphate insecticide: Analysis of 53 human cases with regard to management and drug treatment. Acta Med. Milit. (Belg) 134, 7-14. Willems, J. L., and Belpaire, EM. (1992). Anticholinesterase poisoning: An overview of pharmacotherapy. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B. Ballantyne and T. C. Marrs, eds.), pp. 536-544. Butterworth-Heinemann, Oxford. Willems, J. L., Langenberg, J. P., Verstraete, A. G., De Loose, M., Vanhaesebroeck, B., Goethals, G., Belpaire, EM., Buylaert, W. A., Vogelaers, D., and Colardyn, E (1992). Plasma concentrations of pralidoxime methylsulphate in organophosphorus poisoned patients. Arch. Taxical. 66, 260-266. Willems, J. L., De Bisschop, H. c., Verstraete, A. G., Declerck, c., Christiaens, Y., Vanscheeuwyck, P., Buylaert, W. A., Vogelaers, D., and Colardyn, E (1993). Cholinesterase reactivation in organophosphorus poisoned patients depends on the plasma concentrations of the oxime pralidoxime methylsulphate and of the organophosphate. Arch. Toxical. 67, 79-84. Wilson, B. W, Padilla, S., Henderson, J. D., Brimijoin, S., Dass, P. D., Elliot, G., Jaeger, B., Lanz, D., Pearson, R., and Spies, R. (1996). Factors in standardizing automated cholinesterase assays. J. Taxicol. Environ. Health 48, 187-195. Worek, E, Kirchner, T., Backer, M., and Szinicz, L. (1996). Reactivation by various oximes of human erythrocyte acetylcholinesterase inhibited by different organophosphorus compounds. Arch. Taxicol. 70,497-503. Worek, E, Eyer, P., and Szinicz, L. (1998a). Inhibition, reactivation and aging kinetics of cyclohexylmethylphosphonofluoridate-inhibited human cholinesterases. Arch. Taxicol. 72,580-587.
References
Worek, E, Widmann, R, Knopff, 0., and Szinicz, L. (1998b). Reactivating potency of obidoxime, pralidoxime, HI 6 and HL5 7 in human erythrocyte acetylcholinesterase inhibited by highly toxic organophosphorus compounds. Arch. Toxicol. 72,237-243. Worek, E, Diepold, c., and Eyer, P. (1999a). Dimethylphosphoryl-inhibited cholinesterases: Inhibition, reactivation, and aging kinetics. Arch. Toxicol. 73,7-14. Worek, E, Mast, U., Kiderlen, D., Diepold, c., and Eyer, P. (l999b). Improved determination of acetylcholinesterase activity in human whole blood. Clin. Chim. Acta 288, 73-90. Xintaras, C., and Burg, J. R (1980). Screening and prevention of human neurotoxic outbreaks: Issues and problems. In "Experimental and Clinical Neurotoxicology" (P. S. Spencer and H. H. Schaumburg, eds.), pp. 663674. Williams & Wilkins, Baltimore. Xintaras, c., Burg, J. R, Tanaka, S., Lee, S. T., Johnson, B. L., Cottrill, C. A., and Bender, J. (1978). "NIOSH Health Survey of Velsicol Pesticide Workers, Occupational Exposure to Leptophos and Other Chemicals. " DHEW (NIOSH) Publication 78-136, U.S. Govt. Printing Office, Washington, DC. Xue, S. Z. (1992). Acute anticholinesterase poisoning in China. In "Clinical & Experimental Toxicology of Organophosphates and Carbamates" (B.
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Ballantyne and T. C. Marrs, eds.), pp. 502-510. Butterworth-Heinemann, Oxford. Yilmazalar, A., and Ozyurt, G. (1997). Brain involvement in organophosphate poisoning. Environ. Res. 74, 104-109. Yokoyama, K., Ogura, Y., Kishimoto, M., Hinoshita, F., Hara, S., Yamada, A., Mimura, N., Seki, A., and Sakai, O. (1995). Blood purification for severe sarin poisoning after the Tokyo subway attack. lAMA 274, 379. Yoshida, M., Shimada, E., Yamanaka, S., Aoyama, H., Yamamura, Y., and Owada, S. (1987). A case of acute poisoning with fenitrothion (sumithion). Human Toxicol. 6,403-406. Zadik, Z., Blachar, Y., Barak, Y., and Levin, S. (1983). Organophosphate poisoning presenting as diabetic ketoacidosis. l. Toxicol.-Clin. Toxicol. 20, 381-385. Zoppellari, R, Borron, S. W., Chieregato, A., Targa, L., Scaroni, I., and Zatelli, R (1997). Isofenphos poisoning: Prolonged intoxication after intramuscular injection. Clin. Toxicol. 35,401-404. Zwiener, R. J., and Ginsburg, C. M. (1988). Organophosphate and carbamate poisoning in infants and children. Pediatrics 81,121-126.
CHAPTER
52 Carbamate Insecticides Donald J. Ecobichon Queen's University
52.1 INTRODUCTION Early testing of the natural carbamate, physostigmine (eserine) from the calabar bean (Physostigma venenosum) and the synthetic derivative, neostigmine, revealed that these highly polar compounds possessed no insecticidal activity. Aliphatic esters of carbamic acid were synthesized in the early 1930s and, while showing herbicidal and fungicidal activities, were not insecticidal. These agents will be discussed in other chapters (Chapter 66 for herbicides and Chapter 77 for fungicides). Interest in the carbamates was not renewed until the mid-1950s when there was a search for insecticides having anticholinesterase activity, more selectivity, and less mammalian toxicity than some of the organophosphorus esters then in use. This led to the synthesis of several potent aryl esters of methyl carbamic acid, these agents becoming the insecticides of the 1960s and 1970s. While a large number of carbamates have been synthesized, relatively few were developed further, the pesticide market being limited to less than 20 agents. The early history of carbamate insecticide development has been discussed by Kuhr and Dorough (1976) and Cremlyn (1978).
52.2 NOMENCLATURE As is shown in Fig. 52.1, the structure of all carbamate insecticides is based on carbamic acid (the monoamide of carbon dioxide), a highly unstable compound decomposing into carbon dioxide and ammonia. Carbamic acid may be stabilized by forming salts such as ammonium carbamate or by synthesizing alkyl or aryl esters. Replacement of one hydrogen associated with the nitrogen by a methyl group results in the formation of N-monomethy1carbamic acid which, when combined with an aryl ester substituent, results in significant alterations in various physicochemical properties and introduces insecticidal activity (e.g., bendiocarb, carbaryl, propoxur). An additional group of carbamate insecticides are derivatives of aliphatic oximes rather than esters, resembling aldehydes or ketones, known collectively as methy1carbamoyloximes, and possessing a high degree of toxicity (e.g., aldicarb and methomyl). Handbook of Pesticide Toxicology Volume 2. Agents
The majorIty of the carbamate insecticides in use are N-monomethyl carbamates, frequently referred to as N-methylcarbamates or just methy1carbamates. In this chapter, the insecticides will be referred to by the name commonly used in the literature, thereby simplifying discussion. Table 52.1 lists the methy1carbamates currently used in pest control with their chemical names and chemical structures (Baron, 1991).
52.3 CHEMISTRY The nature of the substituent groups alters both the physicochemical properties of the insecticide and the biological activity. Most of these insecticides dissolve readily in organic solvents but are only slightly soluble in water, thereby conferring varying degrees of lipid solubility. The exceptions are the methy1carbamoyloximes, the "oxime" carbamates aldicarb and methomyl, which are highly water soluble. A wide range of melting points (50 to 150°C) is found for these agents, determined largely by the size of the substituent group. Vapor pressures range from less than 5 x 10- 6 to 5 X 10-2 mmHg (Melnikov, 1971). While high melting points and low vapor pressures enhance the environmental stability of the compound, decomposition can be markedly enhanced by increased temperatures, a 10°C increase raising the hydrolysis rate two- to three-fold (Aly and EI-Dib, 1971; Fukuto et al., 1967). The environmental stability of carbamates is severely affected by photodegradation at short ultraviolet wavelengths (254 nm) and by oxidation upon exposure to air. These aspects of decomposition are discussed succinctly by Kuhr and Dorough (1976). Alkyl esters tend to be relatively unstable in the environment, in contrast to aryl esters. Stability can be enhanced by attaching additional substituents either to the aryl structure or to the carbamoylated nitrogen. While carbamates decompose slowly in water at an acidic pH, alkalinity enhances degradation since the substituent groups tend to draw electrons from around the ester linkage, thereby weakening it and accelerating the hydrolysis by hydroxide ions. Considering some of the structurally different agents shown in Table 52.1, incubation in an alkaline solution (0.01 M sodium barbital buffer, pH 9.3)
1087
Copyright © 200 1 by Academic Press. All rights of reproduction in any form reserved
1088
CHAPTER 52
Carbamate Insecticides I
I
!
0
I
I
I
I
I
,
H-;O-C-N-;-Hl I
Aryl alcohol Aliphatic oxime (R-N-O-) alcohol (R-C-O-)
H2 \ Methyl Dimethyl
Figure 52.1 The basic structure of carbamic acid, the monoamide of carbon dioxide, shoving the positions of substituant methyl, aryl, or aliphatic groups to produce methy1carbamate insecticides.
resulted in the determination of biological half-lives for methiocarb, carbaryl, mexacarbate, and propoxur of 0.4,0.5,2.3, and 3.1 hours, respectively (Abdel-Wahab et aI., 1966). Mono- or dimethylation of the carbamoyl nitrogen results in stabilization of the ester bond. N-monomethylcarbamates degrade slowly in the environment; for example, carbaryl at pH 7.0 has a halflife of 10 days. Dimethylcarbamates are exceedingly stable, the half-life of dimetilan (l-dimethyl-carbamoyl-5-methylpyrazol3-yl dimethylcarbamate) being approximately 100 days at pH ranging from 6 to 10 (Kuhr and Dorough, 1976). In addition to the direct-acting, anticholinesterase methylcarbamates, certain derivative agents such as benfuracarb, carbosulfan, mecarbam, and thiodicarb, known as procarbamates, have insecticidal activity but low mammalian toxicity until they are biotransformed to release biologically active agents or to yield nontoxic, readily excreted products. Fukuto (1983) showed that substitution of the remaining hydrogen on the carbamyl nitrogen of methylcarbamates reduced the mammalian toxicity due to slower conversion of the derivative to the original toxic insecticide. For example, carbosulfan and thiocarb are sulfide derivatives of carbofuran and methomyl, respectively. Most of the carbamate ester insecticides have low vapor pressures, which results in poor volatility at usual temperatures (Fig. 52.2). However, as is shown for aldicarb, oxamyl, and pirimicarb, increasing the temperature can markedly alter the vapor pressure, a factor that must be considered when using these agents in tropical countries. If other carbamate esters behave in the same manner, many would become highly volatile in climates having high temperatures. The propoxur toxicity incident in southern Nigeria, discussed by Vandekar (1965), is an example in point. Propoxur has a vapor pressure of 6.5 x 10-6 mmHg at 20°C; spraying huts and roofs at ambient temperatures of 70°C (140°F) caused acute toxicity among spraymen who wore some protective equipment. The effects were considered to be due to revolatilized propoxur from surfaces rather than from the suspended spray aerosol. Most of the carbamates in commercial use have relatively low water solubility (Table 52.2), a high level of solubility in polar solvents (ethanol, isopropanol, methanol, acetone), and limited-to-moderate solubility in nonpolar solvents (benzene, toluene, xylenes). This lipophilicity enhances the insec-
ticidal potency, the agents readily penetrating insect cuticles and tissues, but it also presents problems of oral and/or dermal absorption in other animal species, and enhanced storage in tissues. There are, however, exceptions, the high water solubility of some carbamates playing important roles in absorption, distribution (both in vivo and environmental), storage, and elimination, as well as governing the regulation of use. Example: note the degree of water solubility of the "oxime" carbamates, aldicarb, methomyl, oxamyl (Table 52.2). These agents are restricted for use on crops with a low water content. There has been illegal use of aldicarb resulting in consumer poisonings from melons and hydroponically grown cucumbers and widespread contamination of groundwater and community drinking water (Fiore et aI., 1986; Goes et aI., 1980; Goldman et aI., 1990a, 1990b; Zaki et aI., 1982).
52.4 TOXICOKINETICS 52.4.1 ABSORPTION
The most likely route of exposure to carbamates is via the skin in an occupational setting. The lipophilicity of this class of agents and the fact that most formulations contain organic solvents and emulsifiers insure a rapid dermal penetration and absorption into the systemic circulation. Temperature and humidity play important roles; high temperature and relative humidity enhance absorption, environmental conditions being reflected in less clothing being worn, greater areas of skin being exposed, and greater subdermal vasodilatation and perspiration, all resulting in a more complete absorption. Carbamates are readily absorbed in the gastrointestinal tract, the efficiency of absorption being somewhat guided by the vehicle(s) in which they are administered or are formulated. Exposure to low levels of carbamate residues in fresh fruits and vegetables may occur where regulatory tolerances have been established for food crop use. Residues in edible foods may be less efficiently absorbed, being trapped or bound in the food bolus. Under certain circumstances, inhalation may be an important route of exposure. The vapor pressures of some carbamates (Fig. 52.2) make them vulnerable to rapid revolatilization when applied under climatic conditions of high temperature in excess of 60-70°C. The previously mentioned accidental poisonings by propoxur in southern Nigeria are a case in point (Vandekar, 1965). The spraymen, applying a 5.0% suspension of propoxur on hut walls and roofs, had to terminate the operation within 2 to 3 hours when severe symptoms of toxicity were observed due to revolatilized agent rather than to the spray aerosol. Similar conditions would be encountered in greenhouses and mushroom barns, areas of high temperature and high humidity. These conditions alter the behavior of aerosols, the change being primarily to keep them suspended in air for long periods of time.
52.4 Toxicokinetics
1089
Table 52.1 Structure, Common Names, and Chemical Names of Carbamate Insecticides Agent
CAS number
Aldicarb
CAS 116-06-3
IUPAC chemical name
Structure Me
I
2-methyl-2-(methylthio)pro-
I
pionaldehyde O-methyl-
MeS.C.CH: N.O.CO.NHMe
TEMIKTM Bendiocarb
Me
CAS 22781-23-3
OC ~
FICAM™ ROTATE™
carbamoyloxime
I
°\;Me
/\e
2,2-dimethyl-l,3-benzodioxol4-yl methylcarbamate
°
O.C.NHMe
8
Carbaryl
CAS 63-25-2
Carbofuran
CAS 1563-66-2
I-naphthyl methylcarbamate
w,~
2,3-dihydro-2,2-dimethyl-7 benzofuranyl methylcarbamate
MeNHCO.O
Carbosulfan
CAS 55285-14-8
2,3-dihydro-2,3-dimethyl7 -benzofuranyl( (dibutyl-
ADVANTAGE™
amino )thio) methylcarbamate
MARSHAL™
Formetanate
CAS 23422-53-9 HCl
HCl CARZOL™
aminophenyl methylcarbamate hydrochloride
I
O.CO.NHMe
DICARZOL™ Methiocarb
0-'' '0',
3-dimethylaminomethylene-
CAS 2032-65-7
4-methylthio-3,5-xylyl methylcarbamate
Methomyl
CAS 16752-77-5
~
MeS"
C-N-OCNHMe
M./
LANNATETM Mexacarbate
CAS 315-18-4
ZECTRAN
Me
o.~~ }'I"~
S-methyl N-(methylcarbamoyloxy) thioacetimidate 3,5-dimethyl-4-(dimethylamino )phenyl methylcarbamate
Me
Oxamyl
CAS 23135-22-0
o
0
11 11 Me,N.C.?=N.O.C.NHMe
SM.
VYDATE™ Pirimicarb
CAS 23103-98-2
N,N-dimethyl-2-methylcarbamoyloxyimino-2-(methylthio )acetamide 2-dimethylamino-5,6-dimethy1pyrimidin-4-yl dimethylcarbamate
(continues)
1090
CHAPTER 52
Carbamate Insecticides
Table 52.1 (continued) Agent
CAS number
Propoxur
CAS 114-26-1
Structure
IUPAC chemical name 2-isopropoxyphenyl methylcarbamate
Thiodicarb
CAS 59669-26-0
Dimethyl N,N-(thiobis(methyl-
rH, ~ CH,-S-C=N-O-C-7-CH,
imino)carbony loxy)-bis-
S
(ethanimidothioate)
I
CH,-S-r=N-O-ji-N-CH, CH,
52.4.2 BIOTRANSFORMATION A number of excellent reviews consider the biotransformation of carbamate insecticides, induding those of Knaak (1971), Ryan (1971), Fukuto (1972), Kuhr and Dorough (1976), Wilkinson (1976), Kulkarni and Hodgson (1980), and the IPCS
W21
0
(1986). The initial response of any species exposed to a carbamate ester is to convert the chemical into more polar forms for ready excretion via the urine. To achieve this, the organism calls upon Phase I and Phase 11 detoxification mechanisms in tissues to create water-soluble, easily excreted, and less toxic by-products. While these insecticidal esters are suscep-
CARBAMATE ESTERS VAPOR PRESSURE (mmHg)
1/100
•
VP (mmHg)
w3
1/1,000
1/10,000
DlllETILAN
·~A e
0
ALllOXYCARB
PIRIMICARB
/
/0~OMETHOMYV oo )y 1
CARBARYL
THIODICARB
METHACARB
1/100,000
BUFENCARB
0/ ... 0
/ PROPOXUR
o o
w6
CARBOFllRAN
BENDIOCARB
1/1.000,000 o
10
20
i
CARBOSULF AN
30
40
50
60
70
TEMPERATURE (CC) Figure 52.2 The vapor pressures of carbamate insecticides determined at 20°C, 25°C, or 30°, with examples of altered vapor pressures at elevated temperatures as might be encountered in tropical countries.
52.5 Mechanism Table 52.2 Relative Water Solubility of Carbamate Ester Insecticidesa,b Agent
Solubility (g/L)
Aldicarb
6.0
Bendiocarb
0.04
Carbaryl
0.7
Carbofuran
0.7
Carbosulfan Fonnetanate HCl Methiocarb Methomyl Mexacarbate Oxamyl Pirimicarb
0.0003 >500 0.01 58 0.1 280 2.75
Propoxur
2.0
Thiodicarb
0.035
a Data from Baron (1991) and the Merck Index, 12th edition (1996).
bMeasured at 20--25°C.
tible to a variety of enzyme-catalyzed detoxification reactions, the principal biotransformation pathways involve oxidation and hydrolysis, with conjugation of some of the cleaved products (Ecobichon, 1994a). The nature and position of the substituent groups on the ether oxygen or the nitrogen exert an important role over the rate and pathway of biotransformation. Being esters, carbamate insecticides are susceptible to hydrolysis by nonspecific carboxylesterases ubiquitously distributed throughout the tissues of species from insects to humans. The products formed are identical to many of those produced by chemical (alkali, water) hydrolysis in that an aryl alcohol plus methyl- or dimethyl-carbamic acid will be formed. The unstable methylated carbamic acids will rapidly decompose into carbon dioxide and mono- or dimethylamine. Rates of hydrolysis in vivo are governed by the molecular structure of the agent, the specificity or selectivity of the carboxylesterases for particular agents, and interspecies differences. Carbamate esters are actually poor substrates for many tissue esterases. The hydrolysis of the various carbamate esters is highly individualistic, only a certain percent hydrolysis occurring with different agents (Schlagbauer and Schlagbauer, 1972). A generalization that carbamates can be hydrolyzed by tissue enzymes requires rigorous testing with several carbamates as substrates. The ubiquitous distribution of the reactive hemoprotein, cytochrome P-450, and the various isoenzymatic forms in tissues of all life forms, point to a commitment to the oxidative detoxification of a broad spectrum of both endogenous and exogenous chemicals as a protective measure. These hemoprotein isoenzymes, in conjunction with molecular oxygen, flavoproteins, cytochrome-b5 and reduced nicotinamide adenine dinucleotide phosphate (NADPH), can initiate a variety of enzymatic oxidative/reductive reactions depending upon the nature of the substituent groups on the carbamate ester. Oxidative reactions can be simplified into two groups: (1) oxidation of appropriate side chains, for example, hydroxylation of N-methyl groups
1091
and/or hydroxylation of methyl substituents on aryl moieties to form hydroxymethyl groups, N-demethylation of secondary amines attached to the aryl moiety; and (2) ring hydroxylation through the formation of an epoxide intermediate. In addition, thiocarbamates may undergo S-oxidation by these same oxidative mechanisms; for example, aldicarb can be converted into a sulfoxide and/or a sulfone, depending upon the species being studied. In conjugative or Phase 11 detoxification reactions, a functional group on the molecule, introduced as a consequence of hydrolytic or oxidative biotransformation, is enzymatically reacted with an endogenous substance in the tissues of the life form to produce water-soluble, biologically inactive, and readily excreted products. Depending upon the species of plant or animal being studied, a variety of products may be formed but, in general, the products may be classified as sulfates, glucuronides, glucosides, amino acid conjugates, acetylated amines, or glutathione conjugates, the last being excreted as mercapturic acid derivatives. In mammalian species, the cleaved aryl substituent(s) are conjugated to produce sulfates, glucuronides, and mercapturates. Biotransformation/degradation in aquatic systems, plants, and by microorganisms has been reviewed (IPCS, 1986). Hydrolysis of the carbamate ester bond is the major degradation pathway in soils. In plants, oxidative processes result in ring hydroxylation followed by conjugation with either amino acids (cysteine), phosphates, or sugars to form glycosides. Hydrolysis can occur in some plant species. 52.4.3 ELIMINATION
There is little evidence of extensive carbamate bioaccumulation since biotransformation is relatively rapid. There is at least one report of persistent toxicity in a human intoxication, the signs and symptoms disappearing slowly when the afflicted individual was removed from the source (Branch and Jacqz, 1986a). However, this effect might have been related to slow recovery from agent-induced neuropathy, or altered metabolism, rather than the clearance of any body burden. Excretion of the water-soluble by-products of detoxification occurs relatively rapidly via the urine and/or feces in most vertebrate species. Glucuronide and sulfate derivatives of the aryl substituents are the major products found in the urine. Small amounts of the parent carbamate may be excreted in the urine. Mercapturates are usually found in mammalian feces if they are not broken down in the intestinal tract, reabsorbed systemically, and recycled to form other products to be excreted in the urine.
52.5 MECHANISM Like the organophosphorus ester insecticides, the carbamates elicit toxicity by inhibiting nervous tissue acetylcholinesterase (AChE). However, it is a transient, reversible inhibition, since there is a relatively rapid reactivation of the enzyme in the
1092
CHAPTER 52
Carbamate Insecticides
~Ka k
k
k
EH ~EHAB ~EA ~EH
+
AB
-1
+
+
BH
AOH
Figure 52.3 A schematic diagram shoving the mechanism of interaction between a methylcarbamate insecticide (AB) and acetylcholinesterase (EH), depicting the unstable intermediate complex (EHAB), the carbambylated enzyme (EA), the leaving group (BH), and the spontaneously decarbamoylated enzyme (EH) and the released methyicarbamic acid (AOH).
presence of "tissue" water. The biological effects of the accumulating acetylcholine (ACh) tend to be of short duration, in terms of hours rather than in days to weeks as is seen with organophosphorus esters. As is shown in Fig. 52.3, a reversible carbamate-AChE complex (EHAB) is formed, followed by the hydrolysis of the ester bond and the loss of the aryl or alkyl substituent (BH), the result being a carbamylated enzyme (EA) which is unstable and hydrolyzes in the presence of water to release free and active enzyme (EH) (Ecobichon, 1996). The differences between organophosphorus and carbamate ester inhibitors lie in the rate constants for the various steps in the reaction(s) (Table 52.3). Both classes of insecticides have high affinity constants (Ka = k-ll kl) for the active center of the enzyme, the interaction with the enzyme (EHAB) being almost instantaneous (Hastings et aI., 1970; Reiner, 1971). The rate of carbamy1ation of the enzyme depends largely on molecular complementarity and reactivity, the latter depending on the nature of the leaving group, for example, phenolic and oxime substituents being somewhat better than benzyl alcohols. While carbamylation appears to be reversible from the point of view of the enzyme, it is not reversible from the point of view of the carbamate which is cleaved and loses anticholinesterase potency in the process (Baron, 1991). Thus, the carbamylation constant, K2, will vary considerably between carbamate esters. Acetylcholinesterase inhibition varies in degree with the rate of the EHAB-to-EA complex formation and the relative Ka of each compound. The decarbamylation constant, K3, would be the same for all N-methylcarbamates, the moiety (A) adhering to the enzyme being identical in all cases, with aqueous hydrolysis at the same rate resulting in the formation of free, uninhibited enzyme (EH). By contrast, the phosphorylation of AChE is regulated by (1) the electron-withdrawing power of the "leaving" substituent, which is highly variable between chemicals; and (2) the nature of the alkyl (methyl, ethyl, isopropyl, methylarnido, ethylamido, etc.) substituents on the ester. The rate of reactivation of AChE is governed by the rate constants, K2 and K3, frequently quite different from those for carbamate esters (Table 52.2). The phosphorylated enzyme can be quite stable, aqueous hydrolysis being very slow in many cases. The degree of inhibition of nervous tissue AChE and/or plasma pseudocholinesterase (PChE) by carbamates is variable, being dependent upon the specificity of the agent for the active site of the enzyme, the rate constants for complex formation,
Table 52.3 Kinetic Rates of Inhibition of Cholinesteerases By Carbamate and Organophosphorus Esters Kinetic
Reaction rates
Parameter
constantsa
Organophosphorus
Carbamate
Complex Formation
LJ/kJ
Rapid (high affinity)
Rapid (high affinity)
Inhibition Rate
k2
Rapid to moderately rapid
Variable
Reactivation Rate
k3
Slow to extremely slow
Relatively rapid
aSee Fig. 52.3.
spontaneous reversal of the complex, the carbamylation of the enzymes, and the decarbamylation stage. Carbamate variability is reflected in the relative rate(s) of recovery of AChE and PChE and the level of exposure. In mild-to-moderate cases of intoxication, carbamates may have little effect on PChE while severely inhibiting the AChE (both erythrocytic and nervous tissue). In severe intoxications, both PChE and AChE will be markedly inhibited. As an example, in a case of a suicidal attempt with a propoxur formulation, the blood sample taken within an hour of visiting the emergency room revealed no activity of either erythrocytic AChE or PChE, but the sample taken 6 hr later showed 60% inhibition of erythrocytic AChE and no residual inhibition of PChE (Ecobichon, unpublished). The transient nature of carbamate-induced inhibition of AChE poses several problems in the attempt to measure the level of inhibition. Care must be taken to keep blood and tissue samples cold or frozen during transportation to the laboratory prior to analysis. For example, blood samples should be kept on ice, centrifuged under refrigerated conditions to recover both the plasma and the erythrocyte fractions, and frozen at -20°C immediately until assayed. Spontaneous reversal of the inhibition is rapid and can be accelerated by (1) the time interval between sampling and analysis; (2) the dilution of the sample; (3) the addition of substrate, usually acetylcholine at high concentration, which competes successfully for the enzymatic active site in either of the EHAB and EA complexes; and (4) the duration of the assay time. Laboratory assays of cholinesterase inhibition must be very rapid (less than 3 minutes), and must employ minimal dilution and minimal amounts of substrate. Modifications can be made to the colorimetric assay of Ellman et al. (1961) to meet the restrictive criteria mentioned above. It is fallacious to measure cholinesterase activities in biological fluids and tissues collected in subchronic and chronic exposure studies 24 hours after the last exposure. Such assays should be done immediately following the last treatment since, as was seen in the case of acute aminocarb (4-dimethylaminom-tolyl methylcarbamate) toxicity in rats, recovery of the vital cholinesterases was complete by 6 hours post-treatment (Vassilieff and Ecobichon, 1983). Little or no inhibition would be observed if the activities were measured 24 hours following the last exposure.
52.6 Toxicology
1093
Table 52.4 Signs and Symptoms of Anticholinesterase Insecticide Poisoning Nervous tissue and receptors affected
Site affected
Manifestations
Parasympathetic autonomic
Exocrine glands
Increased salivation, lacrimation, perspiration
Eyes
Miosis (pinpoint and nonreactive), ptosis, blurring of vision, conjunctival
Gastrointestinal tract
Nausea, vomiting, abdominal tightness, swelling and cramps, diarrhea,
Respiratory tract
Excessive bronchial secretions, rhinorrhea, wheezing, edema, tightness in chest,
Cardiovascular system
Bradycardia, decrease in blood pressure
(muscarinic receptors)
injection, "bloody tears"
postganglionic nerve fibers
tenesmus, fecal incontinence bronchospasms, bronchoconstriction, cough, bradypnea, dyspnea
Parasympathetic and sympathetic
Bladder
Urinary frequency and incontinence
Cardiovascular system
Tachycardia, pallor, increase in blood pressure
Skeletal muscles
Muscle fasciculalions (eyelids, fine facial muscles), cramps, diminished
autonomic fibers (nicotinic receptors) Somatic motor nerve fibers
tendon reflexes, generalized muscle weakness in peripheral and respiratory
(nicotinic receptors)
muscles, paralysis, flaccid or rigid tone Restlessness, generalized motor activity, reaction to acoustic stimuli, tremulousness, emotionallability, ataxia Brain (acetylcholine receptors)
Central nervous system
Drowsiness, lethargy, fatigue, mental confusion, inability to concentrate, headache, pressure in head, generalized weakness Coma with absence of reflexes, tremors, Cheyne-Stokes respiration, dyspnea, convulsions, depression of respiratory centers, cyanosis
Source: From Ecobichon and Joy (1982).
The toxicity of carbamates in mammals can be predicted in vitro by the degree to which they inhibit AChE activity, and in vivo by the severity of the clinical manifestations (Feldman, 1999).
52.6 TOXICOLOGY 52.6.1 MODE OF ACTION The insecticidal carbamates, like organophosphorus esters, exert their effects by inhibiting nervous tissue AChE found in the synaptic spaces and on the postsynaptic membranes of all neurons, using acetylcholine as a chemical neurotransmitter. The role of this enzyme is to terminate, by hydrolysis, the biological actions of the neurotransmitter, thereby restoring the acetylcholine receptors to a state where they can receive the next chemical stimulus. With the loss of this regulating mechanism, the accumulating, nondetoxified acetylcholine (ACh) continues to stimulate specific receptor types, eliciting a spectrum of characteristic clinical signs and symptoms of intoxication (Cranmer, 1986; Ecobichon, 1994b, 1996). Due to the transient nature of carbamate-inhibited nervous tissue AChE, acute intoxication by carbamates is generally resolved within a few hours. Depending upon the level of exposure, the clinical signs and symptoms may appear quite rapidly, be of mild-to-severe intensity, but last for a relatively short duration, disappearing within six hours.
Acetylcholine is an important neurotransmitter at parasympathomimetic, postganglionic nerve endings that are not under voluntary control (autonomic pathways) and which include the exocrine glands, the eyes, the gastrointestinal tract, the respiratory tract secretions, the cardiovascular system, and the bladder (Ecobichon, 1994b). Such neuronal junctions are stimulated specifically by the chemical muscarine and are blocked by atropine, an agent used in treating intoxications to alleviate what are called muscarinic effects, which frequently appear early in any carbamate intoxication. Acetylcholine is also a neurotransmitter at the interneuron ganglia of both the parasympathomimetic and the sympathomimetic divisions of the autonomic nervous system, the major effects seen being a stimulation of the ganglia of sympathetic, adrenergic neurons and the adrenal medulla (releasing epinephrine), with observed clinical signs in the cardiovascular system (tachycardia, vasoconstriction) resulting in increased heart rate and blood pressure and pallor. These neuronal junctions are also stimulated by nicotine, giving rise to the term nicotinic receptors. Acetylcholine stimulates skeletal neuromuscular junctions under voluntary control (the somatic nervous system), these neuromuscular receptors characteristically being stimulated by nicotine and blocked by the agents d-tubocurarine and succinyldicholine. These receptors are known as nicotinic receptors. Overstimulation of such receptors by acetylcholine causes generalized increased motor activity with muscle fascicula-
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tions. An excess of neurotransmitter may lead to receptor blockade' the evident clinical signs being skeletal muscle paralysis and/or generalized muscle weakness, as well as respiratory distress due to paralysis of the diaphragmatic and intercostal muscles (Ecobichon, 1996). Acetylcholine has important roles in the central nervous system, cholinergic brain receptors being both muscarinic and nicotinic in nature. The respiratory center is cholinergic in nature, controls the respiratory rate (overstimulation causes blockade' respiration is impaired or stops), and responds to atropine treatment. Convulsions are elicited through centrally located neurons, and a host of other effects (disorientation, anxiety, memory loss, drowsiness, lethargy, fatigue, general malaise) appear to have central origins. The acronym "MUDDLES" (i.e., miosis, urination, diarrhea, diaphoresis, lacrimation, excitation of the CNS, salivation) is an accurate description of the principal effects of AChE inhibition (O'Malley, 1997). A detailed listing of clinical signs and symptoms observed in animals and humans is presented in Table 52.4 (Ecobichon, 1996). The appearance of none, some, or all of the symptomatology is largely dependent upon the compound and the level of exposure (Cranmer, 1986; Vandekar, 1965; Vassilieff and Ecobichon, 1983). The rate(s) of recovery will be dependent upon the rate(s) of biotransformation and excretion of the particular chemical, most intoxications being brief but with some signs, particularly in humans, persisting for weeks after exposure. The persistent peripheral- and central-mediated symptoms will be addressed in a later section. The acute toxicity of different carbamate insecticides correlates well with their anticholinesterase activity, particularly with the inhibition of erythrocytic AChE (Vandekar et aI., 1971; Vassilieff and Ecobichon, 1983). Intoxications showing obvious cholinergic signs of toxicity may be accompanied by little or no inhibition of cholinesterase activity, this phenomenon being due to a number of assay problems: (1) the selection of the proper enzyme for assay, the plasma PChE being less sensitive to carbamates than the erythrocytic AChE; (2) the selection of an inappropriate substrate for the enzyme being assayed; (3) the ease with which the carbamoylated cholinesterase spontaneously reactivates following dilution, lysis in the case of erythrocytes, or addition of substrate, all factors related to the assay method being used; and (4) the interval between exposure and blood sampling, during which time the carbamate may be degraded or the inhibition may be reversed in vivo (Berry, 1971; Ecobichon and Comeau, 1973; Iverson, 1975; Reiner, 1971; Wilhelm and Reiner, 1973). Particular attention should be paid to the analytical method, which should incorporate minimum exposure-to-collection intervals, minimum dilution of sample, minimum assay time, minimum substrate concentration, and the appropriate pH.
52.6.2 ACUTE TOXICITY-ANIMAL One index of acute toxicity is reflected by an LD50 value determined in suitable animal species, the agent being administered
Table 52.5 Acute Oral Toxicity of Carbamate Insecticides (Technical)Q Chemical Aldicarb
Species
Sex
rat
both
0.46-1.23
mouse
both
0.38-1.50
rabbit
1.3
guinea pig Bendiocarb
rat
mouse
Carbaryl
1.0 both
34--156
M
138
both
350--657
both
28-45
M
175
both
173-380
rabbit
both
35-40
guinea pig
F
35
dog
both
rat
both
mouse
both
108-650
rabbit
?
710
guinea pig cat
233-850
250--795
?
swine
125-250 1500--2000 >1000
monkey rat
ca. 300
280
dog
Carbofuran
LDso (mg/kd
M
5.3-13.2 2.0
mouse
19
dog rat
both
90--250
mouse
both
33-124
rabbit
both
37-53
rat
both
15-26
mouse
both
13-25
dog
both
19
rat
both
13-135
guinea pig
both
14--100
dog
both
10--25
Methomyl
rat
both
12-48
Mexacarbate
rat
both
8.5-12.0
Oxamyl
rat
both
2.5-16.0
mouse
both
2.3-3.3
guinea pig
M
7.1
rat
F
mouse
F
107
dog
both
100--200
rat
both
80--191
mouse
both
Carbosulfan
Formetanate HCl
Methiocarb
Pirimicarb
Propoxur
37-109 40
guinea pig Thiodicarb
68-221
rat
both
mouse
both
39-136 226
guinea pig
M
160
rabbit
both
556
monkey
both
467.2
aData modified from Baron (1991). bValues determined using different vehicles.
52.6 Toxicology Table 52.6 Acute Dennal Toxicity of Carbamate Insecticides (Technical)a Sex
LDso (mg/kg)b
rat
both
rabbit
M
3.2->10 5.0--20 566 >5000 >1000 >2000 >2000 >10200 >300-->5000 >2000 >1000-->2400 556->1500 >2000 >1200 740 >500 >500 1000-->2400 >500 2540 >6310
Chemical
Species
Aldicarb Bendiocarb
rat
both
Carbaryl
rat
both
Carbofuran
rat
both
rabbit
both
Carbosulfan
rabbit
both
Fonnetanate HCl
rabbit
both
Methiocarb
rat
both
rabbit
both
Methomyl
rat
M
rabbit
both
Mexacarbate
rabbit
both
Oxamyl
rat
M
rabbit
M
rat
F
Pirimicarb
rabbit Propoxur Thiodicarb
rat
both
rabbit
M
rat
M
rabbit
both
QData modified from Baron (1991). bYalues detennined using different vehicles.
via the route(s) by which humans are most likely to acquire the chemical (Ecobichon, 1996). To this end, for comparative purposes, Tables 52.5, 52.6, and 52.7 list the oral, dermal, and inhalation LD50s of the carbamate ester insecticides of commercial interest, these tables being reproduced from the 1991 edition of Hayes' and Laws' Handbook of Pesticide Toxicology (Baron, 1991). The specific references for any particular LD50 may be found in that text. Table 52.7 Acute Inhalation Toxicity of Carbamate Insecticides (Technical)a Chemical
Species
Carbaryl
rat
Carbosulfan
rat
Sex
both
Fonnetanate HC]
rat
both
Methiocarb
rat
both
Methomyl
rat
M
Oxamyl
rat
both M
Pirimicarb
rat
?
Propoxur
rat
M
Thiodicarb
rat
both
QData modified from Baron (1991). bYalues detennined over different time intervals (1--6 hr).
LDso (mg/L)b
0.005-0.023 0.61-1.53 0.29-2.8 >0.322 0.45 0.12-0.17 0.064 ca. 0.3 >1.44 0.116-0.22 >0.20
1095
The signs and symptoms of carbamate-induced, acute toxicity observed in various animal species should be comparable for the different insecticides, given that adequate, toxic doses have been administered. Considering carbaryl as a prototype carbamate ester of moderate toxicity, the following signs will be seen in mammals in approximate order of appearance, beginning some 15 to 30 minutes after oral administration: salivation, lacrimation, increased respiration with rales due to bronchial secretions, urination, defecation, and muscle fasciculations and tremors progressing to mild-to-moderate convulsions within 90 minutes of treatment. More severe intoxications may be characterized by pupillary constriction, profuse salivation, chromodacryorrhea, respiratory difficulty, loss of bladder and bowel control, muscular spasms and weakness, prostration, and incoordination. While most of the symptoms will disappear within 6 hours of exposure, a few, such as diarrhea, chromodacryorrhea, and muscle weakness, may persist beyond 24 hours posttreatment. Death is due to respiratory collapse if intoxication is severe. A number of studies have examined the behavioral effects of anticholinesterase-type insecticides immediately following treatment. Carbaryl produces CNS depressant effects, making it obvious that ACh plays a significant role in memory, cognitive, and motor functions; many of the adverse effects are ameliorated by such cholinolytic agents as atropine or scopolamine (Kurtz, 1977; Takahashi et aI., 1991). The acute administration of carbaryl (1.0, 3.0, 5.0, and 10 mg/kg) to rats resulted in a dose-related decrease in variable interval response rates in a learned procedure of pushing a lever to receive a food pellet (Anger and Wilson, 1980). The rate decreases were 55 to 77,81 to 94, and 88 to 100 percent at 3.0, 5.0, and 10 mg/kg, depending upon the route (ip or im) of administration. In other acute experiments, both propoxur and carbaryl caused post-treatment reductions in motor activity (open field and figure eight mazes) in a dose-dependent manner (Ruppert et aI., 1983). However, maze activity recovered within 30 and 60 minutes, while the brain AChE activities remained depressed for 120 to 240 minutes for propoxur and carbaryl, respectively. These results suggest several possibilities, including no association between behavior and AChE or some threshold effect of ACh counteracted by the spontaneous recovery of sufficient AChE activity. A more recent intoxication in both sheep and humans involved aldicarb contamination of a buckwheat field into which the sheep had been moved (Grendon et aI., 1994). Of the 318 sheep, 288 died rapidly from acute poisoning, exhibiting respiratory distress, hypersalivation, miosis, diarrhea, and seizures. Reduced erythrocytic AChE activity was measured in five animals tested, and levels of aldicarb ranging from 0.19 to 344 ppm were detected in the rumen contents of 13 of the exposed animals. The remaining live sheep, given atropine, showed some clinical improvement but continued to have poor appetites, showed body weight loss, and, within 3 weeks, either had died or were euthanized. The shepherd was affected with difficulty in breathing and a burning sensation in his throat. Those arriving to assist the owner experienced classical acute signs and
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symptoms. Chronic symptoms, evident in some of the humans, will be considered in a later section. From acute animal studies, reports in the literature suggest that carbamates possess another mechanism of action in addition to that of inhibition of nervous tissue AChE. In some experiments, animals died within a few minutes of receiving the agent, seemingly from a marked anesthetic-like or "narcotic" effect accompanied by severe respiratory difficulty (dyspnea) and eventual respiratory failure (Vandekar et al., 1971). These effects have been observed with intravenous and intraperitoneal administration but not with oral administration. The "narcotic" effect was produced only by carbamates of low toxicity (high LDso values) (Ecobichon, 1994a). Hypotheses have suggested that such agents cause a complete blockade of nerve conduction by direct action at the level of sodium ion transport across axonal membranes and/or at motor end plate, postsynaptic, ACh receptors, both effects indicating a possible interaction of the agent with membranes to cause perturbation. A similar effect was noted following the intravenous injection of some organophosphorus ester insecticides of low toxicity (Heath, 1961). 52.6.3 ACUTE TOXICITY-HUMAN
Despite statements to the effect that "most" carbamate ester insecticides are relatively safe and produce only transient, short-term toxicity in animals following acute administration, carbamate toxicity does occur in humans, particularly in cases of ingestion by accident or with suicidal intent (Ecobichon, 1994a; Hayes, 1982). Invariably, acute toxicity in humans is associated with one of the more acutely toxic carbamates such as aldicarb, methomyl, or propoxur. Because of long and extensive use, several reported carbaryl intoxications have been summarized in the literature (Cranmer, 1986; Dickoff et aI., 1987; Farago, 1969; Hayes, 1982). Fatalities have occurred, one particular case being well documented (Farago, 1969). In this situation, a 39-year-old male purposefully drank approximately 500 mL of Sevin-80™ (80% concentration of carbaryl), with death occurring some 6 hours after ingestion, even with prompt hospitalization, gastric lavage, and antidote administration. Quantitative analysis of tissues and fluids revealed carbaryl concentrations (ppm): stomach lavage fluid (2,446), stomach contents (148), intestinal contents (176), blood (14), liver (29), kidney (25), and urine (31). No measurements of cholinesterases were reported (Farago, 1969). Overexposure to mexacarbate as a consequence of leakage from a high-pressure pump line in a cockpit resulted in acute intoxication of a copilot (Richardson and Batteese, 1973). Approximately 1lO minutes post-exposure, the copilot experienced the characteristic cholinergic symptoms. On landing, the affected individual was unable to stand, shook uncontrollably, and developed paresthesia and paralysis of the hands and arms and slurred speech while in transit to the hospital. After atropine treatment was initiated at the hospital, the symptoms disappeared rapidly and the patient was discharged three hours
after admission, the only residual effects being headache and weakness for the remainder of the day. Carbofuran-induced occupational intoxication has occurred, two plant employees being affected while preparing a lO% granular formulation (Tobin, 1970). Taken to their respective physicians within 3 hours after the onset of symptoms, one patient received atropine, while the second was not treated since the symptoms appeared to be regressing. The atropine-treated patient recovered fully within 30 minutes, while the untreated patient recovered over the course of 2 to 3 hours. Hayes (1982) reported several interesting poisoning and voluntary consumption cases never published in the literature, giving insight to the amount(s) required to elicit toxicity. In one case, a physician, testing the efficacy of carbaryl as an anthelmintic, ingested 250 mg of carbaryl (2.8 mg/kg) and experienced sudden, violent epigastric pain and profuse sweating within 20 to 30 minutes, followed by a gradually developing lassitude and vomiting (twice). By 3 hours after self-administration, and after having taken 3.0 mg of atropine, improvement of the symptoms occurred and, by 4 hours, the individual had recovered completely. In a second case described by Hayes (1982), arising from a personal communication from a professional colleague who was also testing the anthelmintic efficacy of carbaryl, an oral dose of 420 mg of carbaryl (5.45 mg/kg) was ingested on an empty stomach. The signs and symptoms appeared in the order shown in Table 52.8. The severity of the symptoms reached a maximum by 120 minutes after ingestion of the carbaryl. By the third hour after ingestion, the symptoms were dissipating, and the patient felt "practically normal" by the fourth hour. In a limited volunteer study reported by Wills et al. (1968), volunteers receiving acute oral doses of carbaryl (0.5, 1.0, and 2.0 mg/kg) showed Table 52.8 Time Sequence of Appearance of Symptoms in a Carbaryl Intoxicationa Time interval Symptoms observed
(min)
Blurred vision persisting for 10 to IS min Nausea Lightheaded Nausea, lightheaded, continuing sweating Hyperperistalsis Persistent nausea without vomiting or diarrhea
120
Weakness Pulse rate-normal 64/min Respiratory rate-I8/min No pinpoint pupil No lacrimation, salivation or rales Answered questions readily and correctly Improvement of symptoms
180
Some increase in strength Practically normal, walking about
240 a Data from Hayes, Jr. (1982). b Atropine
administered, 20 mg at 10 min 2.8 mg at 17 min.
52.6 Toxicology no subjective or objective effects. In a further study in which carbaryl doses of 0.06,0.12, and 0.13 mg/kg/day were administered for six consecutive days to volunteers, few conclusive abnormalities attributable to carbaryl were observed. The suicidal poisoning described by Dickoff et al. (1987) details the case of a 23-year-old male who swallowed 100 mL of Ortho-Liquid Sevin™ containing 27% carbaryl, showed the classical signs and symptoms reported earlier but, following recovery from the acute cholinergic toxicity, developed an acute weakness in the arms and legs associated with a peripheral, axonal neuropathy. The apparent delayed neurotoxicity will be addressed in a later section. One unpublished incident reported by Hayes (1982) involved the aerial application of carbofuran in place of carbaryl on corn, with the rapid onset of mild-to-moderate symptoms and the rapid recovery from the intoxication. Within 12 hours of application, 142 teenaged boys and girls entered the sprayed field to remove tassels from the plants. Within 6 hours, 74 complained of dizziness, nausea and/or blurred vision; some 45 received medical aid, with 29 being hospitalized and all but one individual being released within a few hours. Propoxur has frequently been associated with occupational intoxications, volunteer trials, accidents and suicidal attempts (Hayes, 1982; Vandekar et aI., 1968, 1971). Voluntary ingestion of propoxur at concentrations of 1.5 mg/kg resulted in a depression of erythrocytic AChE to 27% of normal at 15 minutes, with a return to normal activity by 120 minutes postingestion. Symptoms such as discomfort, head pressure, blurred vision, pallor, nausea, sweating, increased pulse rate (from 76 to 140/minute), increased blood pressure (from 130/90 to 175/95 mmHg) were observed by 60 minutes after ingestion and, by 30 minutes, pronounced nausea, repeated vomiting, and profuse sweating were observed, these symptoms persisting through 45 minutes after ingestion. By 60 minutes, the individual was feeling better, the signs and symptoms disappearing and, by 120 minutes, was feeling well enough to eat (Vandekar et aI., 1971). Ingestion ofpropoxur at a level of 0.36 mg/kg caused a rapid decrease in erythrocytic AChE to 57% of normal activity within 10 minutes and recovery within 180 minutes, and produced initial abdominal discomfort, blurred vision, moderate facial redness, and sweating lasting about 5 minutes, with recovery by 3 hours after ingestion (Vandekar et aI., 1971). A more recent, suicidal attempt, using a tickand-flea preparation of propoxur, has been described, giving a detailed list of classic signs and symptoms of severe toxicity over an 8-hour period, including: unconsciousness, labored breathing, bilateral pinpoint pupils, salivation, reduced respiratory movements, regular heart rhythm but with frequent premature ventricular contractions, incontinence with watery stool, cyanosis in the extremities, sweating, no response to painful stimuli, no gag or deep-tendon reflexes, downward deflection of toes during the plantar reflex, myotonic jerks of all extremities, and grand-mal seizure (Remaley et aI., 1988). The patient spontaneously awakened approximately 8 hours after hospital admission and was discharged 4 days after the episode.
1097
The most toxic of the carbamate esters is the systemic insecticide aldicarb, registered for use on citrus fruits, cotton, potatoes, peanuts and soybeans. It is not registered for use on any fruits or vegetables having a high water content. Surprisingly, this highly water-soluble chemical has been the source of periodic outbreaks of poisoning, usually associated with the inappropriate, even illegal, use in hydroponically grown cucumbers (CDe, 1979; Goes et aI., 1980), various melons (CDC, 1986; Goldman et al., 1990a, b), or the contamination of drinking water in New York and Wisconsin (Fiore et aI., 1986; Sterman and Varma, 1983; Zaki et aI., 1982). Levels of aldicarb in cucumbers ranged from 6.6 to 10.7 ppm (Goes et aI., 1980). In the melons, the active ingredient was not the parent insecticide but the equally water-soluble, biologically active metabolite, aldicarb sulfoxide (Goldman et al., 1990a). Estimates of the amounts of aldicarb sulfoxide ingested and responsible for intoxications ranged from 0.0011 to 0.06 mg/kg body weight (Goldman et aI., 1990a). In drinking water derived from groundwater sources, levels of aldicarb ranged between 8 and 75 ).Lg/L in Suffolk County, NY, while in Wisconsin, the levels ranged between 1.0 and 61 ).Lg/L (Fiore et aI., 1986; Zaki et aI., 1982). In all of the cucumber and melon intoxications, classic cholinergic symptoms (diarrhea, nausea, vomiting, sweating, blurred vision, abdominal pain, dyspnea, muscle fasciculations, headache, and, in some cases, loss of function of arms and legs) were observed, persisting some 4 to 12 hours followed by complete recovery (Risher et al., 1987). The exposures to aldicarb in drinking water were less conclusively related to specific signs and symptoms of intoxication (Sterman and Varma, 1983; Zaki et aI., 1982). In an experimental study with human subjects, aldicarb was administered in single oral doses of 0.025,0.05, 0.1 mg/kg body weight, with consequent manifestations of a variety of cholinergic symptoms at the highest level, all of which disappeared by 6 hours after administration (Union Carbide report quoted by Risher et aI., 1987). While abnormal reductions in erythrocytic AChE activity (25% of pre-exposure activity) were measured at the highest dose, the inhibition was rapidly reversible and preceded the disappearance of the symptoms. Methomyl appears periodically as an insecticidal toxic ant in humans, either from occupational exposure or through accidental or suicidal ingestion (Simpson and Bermingham, 1977). One recorded acute poisoning in Jamaica involved unleavened bread prepared with methomyl mistakenly used in place of common table salt, a level of some 1000 ppm being detected (Liddle et aI., 1979). Consumption of the bread was rapidly fatal to three individuals; another was asymptomatic, while the fifth showed generalized muscle fasciculations and respiratory distress. It was estimated that, in those who died, the amounts ingested were equivalent to 12 to 15 mg/kg body weight. In Japan, a 31-year-old woman committed suicide by incorporating methomyl in food, this being eaten by her three children as well (Araki et aI., 1982). A 9-year-old child survived. Autopsies revealed congestion of the stomach lining and lungs, edema, and hemorrhaging due to acute circulatory failure. The amounts ingested were estimated at 55 mg/kg for the mother and 13 mg/kg for a 6-year-old child.
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Carbamate Insecticides
Concerning the more toxic carbamate ester insecticides, aldicarb and methomyl, it has been demonstrated that central effects of these toxicants are more severe in children than in adults; symptoms in adults were miosis, muscle fasciculations, slowing of the heart, and broncorrhea, whereas in children, stupor/coma, hypotonia, and diarrhea were significant effects. Feldman (1999) suggested that the observed differences might be reflected in differences in the permeability of the bloodbrain barrier of children and adults. 52.6.4 CHRONIC TOXICITY-ANIMAL
Subchronic and chronic studies have been conducted in various animal species, including mice, rats, dogs, swine, and monkeys, although many have never been published since they were confidential documents submitted to regulatory agencies in support of product registrations. Summaries of the general chronic effects of carbamate insecticides on various target sites have been published (Cranmer, 1986; Ecobichon, 1994a; IPCS, 1986). Much of the toxicity is associated with the nervous systems, neuromuscular dysfunction, and neurobehavioral changes. The chronic toxicity of carbaryl has been the most extensively reviewed (Cranmer, 1986). Baron (1991) has published a relatively complete inventory of the effects resulting from the subchronic and chronic treatment of mice, rats, and dogs with several carbamate insecticides; much of this data was gleaned from unpublished reports from chemical manufacturers submitted at the time of product registration. Consistently, depression of erythrocytic and plasma cholinesterases were seen associated with cholinergic effects, particularly at higher dosages. Only with some agents were changes seen in food consumption, growth and development, organ weights, hematological and clinical chemistry measurements, urinalysis, and gross and histopathological parameters. The data from chronic dietary exposure studies are summarized in Table 52.9. 52.6.4.1 Neurophysiological Effects
There is evidence from animal studies that long-term exposure to carbamates such as sodium diethylthiocarbamate (a copperchelating agent) and tetraethylthiuram (a rubber vulcanizer and a therapeutic agent in chronic alcoholism) can elicit neurological effects, possibly due to metabolism to carbon disulfide, a neuropathic agent (Barry, 1953; Gardner-Thorpe and Benjamin, 1971; Moddel et aI., 1978; Waibel et aI., 1957; Watson et al., 1980). Detailed descriptions of anatomic lesions including degeneration and vacuolization in the peripheral and central nervous systems of rabbits and chickens have been published, suggesting that the lesion pattern was similar to the "dying back" process described for some organophosphorus esters (Cavanagh, 1969, 1973; Edington and Howell, 1966, 1969; Howell and Edington, 1968). The elicited neurotoxicity has been attributed to the biotransformation of the dithiocarbamates to yield carbon disulfide, a known neuropathic agent (Brugnone et aI., 1993; Johnson et aI., 1998).
Miller et al. (1969) demonstrated that a single dose of carbaryl (20 mg/kg) given to miniature swine caused a 44% and 75% inhibition of cerebral cortex and brain stem AChE, respectively, and caused a hindlimb paralysis even though no obvious effects were seen upon histopathological examination. Severe carbaryl-induced neuromuscular effects were observed in another study in swine (Smalley et aI., 1969). High dosages (150 mg/kg/day over 73 to 83 days or 150 mg/kg daily for 28 days followed by 300 mg/kg/day over the next 18 or 57 days) caused severe neuromuscular effects. Reluctance to stand was observed first, followed by a peculiar stance, the rear legs being carried well forward under the body, the animals appearing to be "walking" on their dew claws. There was greatly exaggerated flexion of the rear legs, the animals having difficulty in backing up or sitting down. Forcing the animals to move caused marked incoordination, ataxia, muscle tremors, and clonic contractions. Muscular lesions consisted of a myodegeneration of traumatic or ischemic nature, an acute hyaline and vacuolar degeneration, and an acute degenerative process associated with dystrophic calcification (Smalley et aI., 1969). Carbamate esters have caused severe neuropathy in adult chickens following repeated oral administration, although this neuropathy is different from organophosphorus-induced delayed neuropathy (OPIDN) (Fisher and Metcalf, 1983; Hollingshaus and Fukuto, 1982). In young (3-week-old) chicks, both carbaryl and aldicarb affected locomotor activity for 6 weeks after cessation of the subacute (7 days) exposure (FarageElawar, 1989a, 1989b). The treated chicks walked with an abnormal gait, taking shorter steps but with a wider stance suggestive of ataxia. Some paralysis was seen for up to 40 days post-treatment. Once again, this neuropathy was unlike that seen in OPIDN. While the short-term exposure of rats and dogs to repeated oral doses of carbaryl, carbofuran, and propoxur resulted in the inhibition of plasma, erythrocytic, and brain cholinesterases accompanied by typical acute signs of toxicity, there was little evidence of persistent effects on the central and/or peripheral nervous systems (Cranmer, 1986; IPCS, 1986; Krechniak and Foss, 1982). While rats dying in an acute carbaryl study showed congestion of the brain and meninges, such morphological changes have not been reported in animals treated with less than near-lethal concentrations (Boyd and Boulanger, 1968). Cranmer (1986) cited studies in which morphological changes in the brain ganglia in rabbits and increased brain protein in rats were reported, but suggested that such effects occurred only at doses that reduced cholinesterase activities. Rats and dogs receiving oral aldicarb for up to two years showed no adverse effects (Risher et aI., 1987). Two studies have suggested that long-term exposure to carbamate esters can cause neurotoxicity. In a two-year rat study of carbaryl at levels causing no inhibition of blood cholinesterases or clinical signs, the animals showed electroencephalographic (EEG) changes and had a decreased maze performance (Desi et aI., 1974). In monkeys, the EEG patterns were not adversely affected at carbaryl levels of 1.0 mg/kg/day (Santolucito and Morrison, 1971). Tolerance to carbamates has been reported
1099
52.6 Toxicology
Table 52.9 Chronic Toxicity of Carbamate Insecticidesa Max. dosage
Food consump.
+
Chemical
Species
(mglkg)
Aldicarb
Rat
0.5
Rat
0.3
Dog
0.25
Bendiocarb Carbaryl Carbofuran
Carbosulfan
Rat
200ppm
Dog
500ppm
Rat
400ppm
Dog
1250 ppm
Mouse
500ppm
Rat
100ppm
Dog
500ppm
Rat
500 ppm 2500ppm
Methiocarb
Dog
IOOOppm
Mouse
2500ppm
Rat Dog
Methomyl
Mexacarbate
Oxamyl
Mouse
Wt.
Death
Clin.
tology
Urina-
chem.
Pathology
ysis
+ +
+
+ + + + + + + + + +
+
+
+ +
+ +
+ +
+
+ +
+ + + +
+
+
+
+
+ +
+
+
60ppm
+
800ppm 20--26
Dog
1000ppm 300ppm
Rat
250ppm
Dog
325 ppm
Rat
150ppm
Dog Pirimicarb
+ + + + +
600 ppm
Rat Mouse
+
Hema-
Organ Growth
+ + + + + +
+ + + +
+ +
250--750
Dog
4.0ppm
+
+ + +
150ppm
Rat
+
+
+
ppm Propoxur
Rat
2000 ppm
Rat
750 ppm
Dog
2000 ppm
+ +
+ +
+
"+," while no effects are represented by "-." Maximum dosage is in milligram per kilogram of body weight (mglkg) unless presented as parts per million (ppm) in food.
a Data derived from Baron (1991). Positive effects on a parameter are indicated by
(Costa et aI., 1982). For example, male rats receiving carbaryl (200 mg/kg/day) orally for 3 days/week for 90 days showed no overt toxicity (Dikshith et aI., 1976). In the above-mentioned carbaryl swine study, Smalley et al. (1969) described a clinical syndrome of chronic intoxication characterized by progressive myasthenia, incoordination, ataxia, intention tremor, and clonic muscular contractions terminating in paraplegia and prostration. Moderate-to-severe edema was found in myelinated tracts of the cerebellum, brain stem, and upper spinal cord, as well as fragmentation of myelin sheaths, moderate swelling and rupture ofaxons, necrosis of cellular components in spinocerebellar tracts, and vascular congestion, hypertrophic endothelium, and vascular degeneration and hemorrhage in the gray columns. The authors attributed the pathological effects to vascular changes induced by carbaryl (Smalley et aI., 1969).
52.6.4.2 Neurobehavioral Effects Behavioral changes have been associated with carbamate exposure in many animal studies, although such adverse effects were usually detected immediately following acute administration of sufficient chemical to inhibit cholinesterases, suggesting that the effects were the result of cholinergic-mediated stress. Most studies have involved carbaryl (Cranmer, 1986). Single intraperitoneal doses of carbaryl (8.0 mg/kg) reduced shock avoidance by 50% in treated rats for 30 minutes, while a dose of7.3 mg/kg caused a 50% reduction in a positive reward response test (food presentation) (Goldberg et aI., 1965). In a food reward test in cats, inhaled carbaryl (40 mg/m 3 ) caused a deficit only immediately after exposure (Yakim, 1967). Spontaneous locomotor activity was reduced in rats in a 60-minute period after acute oral carbaryl (0.56 and 2.24 mg/kg), whereas the
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Carbamate Insecticides
daily administration of a higher dose for 14 consecutive days had no effect on "wheel-turning" (Singh, 1973). In contrast, acute intoxication of rats with carbaryl (10 mg/kg) resulted in increased locomotor activity in a familiar environment but reduction in exploratory activity in a new environment (Singh, 1973). A decrease in rats' working memory was carbaryl dosedependent shortly after treatment (Heise and Hudson, 1985). In a feeding study of carbaryl and arprocarb (2-isopropoxyphenylN-methylcarbamate) in rats, the amounts given in the diet to achieve body dosages of 10 and 20 mg/kg over 50 days resulted in the animals showing increased difficulty in performing tasks, forgetting already learned skills (De si et aI., 1974). Over 4-month periods, rats inhaling carbaryl (12 to 23 mg/m 3 for 4 hours daily) showed decreased performance in a maze task for food reward, but if the treatment was extended every two weeks for four months, performance was normal (Viter, 1978). A possible association could be suggested between behavioral alterations and cholinesterase inhibition/acute toxicity, the adverse effects being ACh-related. Learning in monkeys appeared to be affected by small, chronic doses of carbaryl. This led to the use of the technique called "chain acquisition" task in which monkeys learned one set of equivalent response sequences each day; they were trained to make four out of the 12 possible response sequences in a certain order, with the correct responses being changed
every day (Anger and Setzer, 1979). Carbaryl was administered orally (up to 50 mg/kg) or intramuscularly (1.0,3.0,5.0, and 10 mg/kg) to trained macaque monkeys. Oral carbaryl caused no consistent effects on performance, whereas the injected carbaryl elicited significant decreases in total session time and increased errors in performance at and above dosages of 3.0 mg/kg.
52.6.4.3 Mutagenicity
As a class, methylcarbamates are not mutagenic, negative results being obtained in an overwhelming majority of in vitro gene mutation assays using microbial systems, cultured mammalian cells, and such in vivo systems as Drosophila and dominant lethal mutation tests (Baron, 1991; IPCS, 1986). Table 52.10 lists the methylcarbamate insecticides that showed mutagenicity in in vitro and in vivo test systems. While weak mutagenicity has been identified as a property of several carbamates, high, almost toxic dosages were used and, frequently, the results either could not be replicated or were derived from nonstandardized protocols that could not be compared with results from acceptable techniques.
Table 52.10 Mutagenic Potential of Carbamate Insecticides Chemical
Test systems
Aldicarb
Carbaryl
Effectsa
References Rashid and Mumma (1986)
S. typhimurium
DNA damage
Mouse bone marrow cells
CA
Debuyst and Van Larebeke (1983)
Cultured human lymphocytes
SCE
Gonzales and Matos (1984)
Saccharomyces cerevisiae
M
Guerzoni et al. (1976)
S. typhimurium
M
Egert and Greim (1976)
Cultured rodent cells
M
Ahmed et al. (1977)
Cultured rodent cells
CA
Ishidate and Odashima (1977)
Cultured rodent cells
spindle poison
Onfelt (1983)
Human fibroblasts
DNA damage
Ahmed et al. (1977)
D. melanogaster
CA
Hoque (1972)
Rats
mitotic abnormalities
Baron (1991)
S. typhimurium
M
Moriya et al. (1983)
Cultured rodent cells
M
Wojciechowski et al. (1982)
Cultured rodent cells
CA
Debuyst and Van Larebeke (1983)
Formetanate HCl
Cultured human lympho-
M
Baron (1991)
Methomyl
Saccharomyces cerevisiae
M
Guerzoni et al. (1976)
Human lymphocytes
SCE
Debuyst and Van Larebeke (1983)
Cultured mammalian cells
CA
Baron (1991)
Saccharomyces cerevisiae
M
Guerzoni et al. (1976)
Mouse dominant lethal
CA
Baron (1991)
Carbofuran
cytes
Mexacarbate Propoxur Pirimicarb Thiodicarb a Abbreviations:
Cultured mammalian cells
CA
Pilinskaya (1981, 1982)
Human lymphocytes
CA
Pilinskaya (1981)
Saccharomyces cerevisiae
DNA damage
Baron (1991)
M, mutagenic; CA, chromosomal aberrations; SCE, sister-chromatid exchanges.
52.6 Toxicology 52.6.4.4 Reproductive Effects
There is little evidence that methylcarbamate insecticides cause reproductive anomalies in mammals. The common, positive findings have been embryotoxicity and/or fetotoxicity associated with the administration of high dosages and concomitant maternal and possibly nutritional toxicity (IPCS, 1986). There is little evidence of these carbamates causing teratogenicity other than through nutrition-related problems (IPCS, 1986). 52.6.4.5 Carcinogenicity
There has been no evidence of the potential of methylcarbamate insecticides to cause carcinogenicity (IPCS, 1986). However, in studies where positive effects have been found, caution must be observed in interpreting the data because of design inadequacies (limited dosage range, duration of study, numbers of animals, etc.) and the extremely high doses administered. Most of these studies are summarized in the IPCS document (IPCS, 1986); more recent reports have not been seen. 52.6.5 CHRONIC TOXICITY-HUMAN
As has been observed with organophosphorus ester intoxications, persistent effects may result following either acute, single, high-level or repeated, even long-duration, low-level exposure to carbamate esters (Ecobichon, 1994a, b). Involvement of both the peripheral and central nervous systems has given rise to distinctive neurophysiological and/or behavioral anomalies. In 1982, a bizarre case of a 55-year-old male was described following a thorough soaking of skin and clothing by a waterwettable preparation of carbaryl (Ecobichon, 1982). Within 3 to 4 weeks of initial antibiotic treatment for "bacterial meningitis," the patient was reporting headaches, blurred vision, photophobia, peripheral numbness, tingling sensations in the hands and legs, muscle weakness, vertigo, incoordination, lethargy, tiredness, forgetfulness, and loss of recent memory. More alarmingly, behavioral changes persisting for several months manifested themselves as frustrated rage, inability to control temper, severe headaches, and short periods of blackouts. Even at 18 months post-exposure, the patient was unable to drive a car, still experienced photophobia, had persistent short-term memory loss and mild peripheral paresthesia, and suffered from lassitude, lethargy, and muscle weakness. The behavioral aspects persisted, being partially controlled by anticonvulsant and antipsychotic drugs. At the time of this case, no other reports had been published concerning slow developing and/or persistent symptoms arising from exposure to carbamate ester insecticides. Subsequently, a number of such cases have appeared in the literature, all reporting persistent adverse effects (Ecobichon, 1994a; Feldman, 1999). 52.6.5.1 Neurophysiological Effects
Indicative of delayed neurotoxicity following acute exposure to high levels of carbamate insecticides, Dickoff et al. (1987) reported a case involving the ingestion of a liquid preparation of
1101
carbaryl (500 mg/kg), showing a weakness in the arms and legs as well as sensory loss following the acute cholinergic crisis, with electrophysiological alterations consistent with a peripheral axonal neuropathy. Recovery began one week after exposure and progressed for 9 months. However, while recovery appeared to be complete, bilateral severe ankle and toe weakness persisted, accompanied by reduced propioception in the toes and tactile preception below mid-calf. In a second, suicidal case, the ingestion of a metolcarb (m-tolyl-methylcarbamate) formulation (estimated dosage of 1.0-1.2 mg/kg) resulted in neurological damage (Umehara et aI., 1991). Fibrillation potentials and sharp positive waves were observed at rest, with reduced recruitment patterns during muscle contractions. A sural nerve biopsy, performed 38 days after the poisoning, revealed reduced densities oflarge and small myelinated fibers, degenerated axons, and denervated Schwann cell clusters. At 3 months post-exposure, motor symptoms had improved, with a reduction in numbness, although deep tendon reflexes in the extremities were absent. By 6 months, upper motor neuronal signs were no longer observed. In a more recent case of acute exposure to a1dicarb, men handling poisoned sheep showed persistent symptoms similar to those mentioned above, five of the six individuals still experiencing persistent neurotoxicity some three years after exposure (Grendon et aI., 1994). Feldman (1999) described two cases of chronic symptoms following an aerial overspray of carbaryl in a forest spraying program. Both individuals had symptoms persisting for more than six years after the exposure, one (a 35-year-old male) showing hoarseness, bronchorrhea, dizziness, and a peripheral neuropathy. The second individual (a 53-year-old male) suffered recurrent bouts of abdominal cramps and diarrhea, anxiety attacks with associated flashback memories, fatigue, numbness in the feet which gradually affected the hands, a clumsy gait, reduced sensation to mid-calf developing within three months of exposure, and some effects (slowed nerve conduction, reduced muscle action potential amplitude and motor activity in the lower extremities) for some five years postexposure. Electromyographical and nerve conduction studies indicated a peripheral neuropathy with chronic denervation and reinnervation in distal muscles. In contrast to the acute-exposure situation, reports linking incidents of-low-to-moderate, subchronic or chronic exposure to chronic toxicity are almost nonexistent. In one study in which volunteers took daily oral doses of carbaryl (0.06 or 0.13 mg/kg) for six weeks, with concomitant monitoring of plasma and whole blood cholinesterase activities, other blood and urinary biochemical parameters, and electroencephalograms, no deleterious changes were attributable to the agent (Wills et aI., 1968). However, the level of exposure could be considered to be low. One case report of a low-level exposure to carbaryl stands out as being contrary to everything known and expected of carbamate-related toxicity. The propositus patient, a 75-yearold male, and his family (wife and son) were exposed over a period of 8 to 10 months to carbaryl (a 10% dust formulation) applied some six times to the basement area of a home to control fleas (Branch and Jacqz, 1986a). The insecticide
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was dispersed throughout the home by the central air conditioning located in the basement. While the entire family experienced a range of acute influenza-like symptoms (headache, malaise, epigastric discomfort, abdominal colic, diarrhea, nausea, muscle spasms, cough) within 3 days of the initial application, the wife and son appeared to recover. However, the father's signs and symptoms became progressively more severe (headache with intense pressure, tinnitus, vertigo, mild ataxia, rhinorrhea, excessive lacrimation, weakness in major skeletal muscle groups, fasciculations, somnolence, agitation, mental confusion). Both the plasma and erythrocytic cholinesterase activities were below normal values and consistent with an anticholinesterase intoxication. During intervals of living away from the home, the patient's symptoms dissipated, but they returned when he reentered the house for any extended period. Leaving the house permanently, the patient's well-being improved, but it was two months before the plasma PChE was within normal range. Symptoms that failed to abate included alterations in sleep patterns (episodic awakening with headache and tinnitus) and mental confusion that persisted for a further two years. A computerized tomographic scan revealed cerebral atrophy which had not been observed on a pre-exposure scan. A persistent, neurological deficit became more severe, being defined as a stocking-and-glove peripheral neuropathy. There has been considerable criticism of this last case report, including that of drawing a valid inference of carbaryl intoxication based on one case in the absence of any occupational exposure literature of similar agent-related toxicity. No indoor measurements of aerial or surface concentrations of carbaryl were ever made, thereby preventing an estimation of exposure. The advanced age of the patient could have played a role in the rate of carbaryl detoxification. The concomitant treatment of the patient with cimetidine, an H2-receptor antagonist used in treating gastric acidity and peptic ulcers, might have had an effect. In animal studies, cimetidine has been shown to inhibit carbaryl biotransformation, thereby increasing the systemic bioavailability of the agent (Branch and Jacqz, 1986b). However, taken in context with other case reports, the claim of a carbaryl-induced intoxication appears valid. Regular occupational exposure to methomyl, particularly in the packaging area of the manufacturing plant, resulted in a high incidence of anticholinesterase symptoms (constricted pupils, nausea, vomiting, blurred vision, increased salivation, muscle weakness, fatigue) and a high number of hospitalizations (Morse et aI., 1979). A decreased vibratory sensation was noted in 19.8% of the workforce, with hospitalized workers having significantly more vibratory sensory loss than nonhospitalized workers. However, no depression of either plasma or erythrocytic cholinesterase activities was noted. It is of interest to note that some acute carbamate intoxications have resulted in the later development of persistent respiratory problems that were exacerbated by exposure to other chemicals such as solvents, household pesticides, hairsprays, and perfumes (Feldman, 1999; Grendon et aI., 1994). The descriptions are reminiscent of multiple chemical sensitivity, an olfactory sensitizing and triggering phenomenon described by
Ashford and Miller (1998). One of the earliest reports of odor aversion as a consequence of pesticide exposure was that of Tabershaw and Cooper (1966), who reported that several of their cases could no longer tolerate smelling or contact with pesticides, reacting with nausea and vomiting after even a whiff of the agents and being forced to give up work involving contact with agrochemicals. The genesis of this idiosyncratic reaction has not yet been ascertained but appears to have both neurological and psychogenic components. 52.6.5.2 Neurobehavioral Effects Sufficient numbers of cases of both acute and chronic intoxications by organophosphorus ester, anticholinesterase insecticides have revealed a recognizable pattern of delayed and/or persistent neurobehavioral anomalies that can be detected and assessed by appropriate neuropsychological evaluation (Ecobichon, 1994b, 1999; Feldman, 1999). Not surprisingly, close examination of acute and chronic carbamate insecticide intoxications reveals a similar pattern of persistent behavioral effects even though neuropsychological testing of affected individuals has not been reported. The long-term follow-up of acutely or chronically carbamate-exposed individuals has been poor but, in several of the incidents discussed in the previous section, observed symptoms have included vertigo, incoordination, disturbed sleep patterns, anxiety attacks, mood changes, chronic lethargy and fatigue, agitation, mental confusion, and difficulty in performing simple tasks or making decisions (Branch and Jacqz, 1986a; Ecobichon, 1982; Grendon et aI., 1994). Feldman (1999) describes the detailed neuropsychological assessment of two individuals some five or six years following an overspraying by carbaryl applied aerially, using the Wechsler Adult Intelligence Scale (WAIS) test battery. As has been observed with organophosphorus ester insecticide intoxications, many of the component tests were within normal values, but impairment, deficiency, or slowing of performance were detected in: a confrontational naming task (Boston Naming Test), memory tests requiring a delayed recall of information; psychomotor (Digit Symbol Test) evaluation, and tasks of visual spatial organization, cognitive tracking, and reasoning. The Profile of Mood States (POMS) and the Minnesota Multiphasic Personality Inventory (MMPI) revealed persistent fatigue, depression and/or anxiety states. Discrepancies were revealed between verbal and performance intelligence quotients. It is obvious that greater use could be made of refined neurological and behavioral test batteries to evaluate short- and long-term neurological effects of carbamate exposure. 52.6.5.3 Reproductive Effects Few specifics are known about the potential of short- and longterm influence of carbamate insecticide exposure on any aspects of reproduction in either male or female agricultural workers. One paper has reported an increase in sperm abnormalities (abnormally shaped heads) in production workers who had been exposed to carbaryl at the time of sampling (Wyrobek et aI., 1981). Neither the sperm count nor the presence of double
1103
52.8 Exposure Limits
fluorescent bodies was changed. Formerly exposed workers, removed from carbaryl-related occupational activities for an average of 6.3 years, showed only a marginally significant elevation in sperm abnormalities, results suggesting that the carbarylinduced morphological effects may not be reversible or may be only slowly so. A dose-dependent change in sperm morphology could not be established. More research needs to be conducted in this area of toxicology.
52.7 TREATMENT The symptoms (Table 52.8) associated with carbamate insecticide intoxication are associated with the accumulating, unmetabolized neurotransmitter ACh at the nerve endings of the parasympathetic and sympathetic autonomic ganglia, the postganglionic parasympathetic nerve endings, and the neuromuscular junctions of the somatic, motor neurons, as a consequence of the inhibition of the nervous tissue AChE. The reversible nature of the AChE inhibition would suggest that the symptoms would be transient although, depending upon the level of exposure, possibly moderate to severe in nature. Atropine is the antidote of choice, antagonizing the action of ACh by blocking the receptor site for the transmitter at parasympathetic nerve fibers innervating exocrine glands, gastrointestinal tract, respiratory tract, eyes, heart, and bladder, as well as exerting a direct, central effect on the respiratory center (Ecobichon et aI., 1977; Feldman, 1999; Namba et al., 1971). Alleviation of these muscarinic signs will be best achieved by administering frequent small doses (0.5 to 1.0 mg) subcutaneously until there is dilatation (mydriasis) of the pupils and the face becomes flushed and/or sweating disappears. The patient should be carefully titered using these signs as physiological endpoints since excess atropine can cause severe toxicity. This is of particular importance in carbamate ester poisoning where the enzyme-insecticide complex is unstable; the enzyme may be decarbamoylated by the excess ACh and the carbamate ester may be metabolized in a short period of time. Atropine is ineffective in counteracting the nicotinic, neuromuscular effects of the accumulated ACh. Acetylcholinesterase reactivators such as the pyridinium oximes, 2-PAM, P2S, and toxogonin, have been used in carbamate-induced intoxications but with mixed results, their use remaining controversial. Signal animal (rats, dogs) experiments involving carbaryl intoxications revealed that the protective effect of atropine was markedly reduced by the concomitant administration of 2-PAM (Carpenter et aI., 1961; Natoff and Reiff, 1973; Sanderson, 1961). This observation was confirmed in one human carbaryl-related poisoning where it was noted that the patient's condition deteriorated rapidly following the administration of2-PAM (Farago,1969). This led to a generalized condemnation of oxime use as an antidote in carbamate insecticide intoxication (Harris et aI., 1989; Lifshitz et aI., 1994; Natoff and Reiff, 1973; Sterri et al., 1979). However, beneficial effects of oxime treatment were seen in treating aldicarb,
Table 52.11 Exposure Limits For Carbamate Insecticidesa Chemical
OS HA (PEL)
ACGIH
NIOSH REL
IDLH
(TLY)
Carbaryl
5.0
5.0
100
5.0
Carbofuran
0.1
ND
0.1
Methomyl
O.lb 2.5 b
2.5
ND
2.5
Propoxur
0.5 b
0.5
ND
0.5
aYalues represent concentrations (mg/m3) in air. REL values represent a timeweighted average for a lO-hr exposure, while TLY values represent a timeweighted average for an 8-hr exposure. bYa1ues "vacated" by OS HA in 1993.
mecarbam (S-(N-ethoxycarbonyl-N-methyl-carbamoylmethyl) O,O-diethyl phosphorodithioate) and methomyl intoxications (Natoff and Reiff, 1973; Sterri et aI., 1979). In cases involving carbofuran, methiocarb, mexacarbate, thiodicarb, and trimethacarb, 2-PAM was ineffective but did not exacerbate the clinical symptoms or interfere with the antidotal effectiveness of atropine (Baron, 1991; FAOIWHO, 1982). Similar results were seen for P2S with pirimicarb and for toxogonin with methomyl (FAOIWHO, 1982; Sanderson, 1961). Overall, the studies to date suggest the efficacious use of oximes with aliphatic oxime carbamate (aldicarb, methomyl, and possibly mecarbam) intoxications but not their use in carbaryl- or other carbamate-related intoxications (Feldman, 1999). The myorelaxant agent, diazepam, should be considered in treatment regimens of all but the mildest cases of carbamate intoxications to relieve anxiety, to counteract some central nervous system-related symptoms not affected by atropine.
52.8 EXPOSURE LIMITS Given the importance of methyl carbamate insecticides in the agricultural industry, it is surprising that so few exposure limit values have been established for these agents by the Occupational Safety and Health Administration (OSHA), the National Institute for Occupational Safety and Health (NIOSH), and the American Conference of Governmental Industrial Hygienists (ACGIH). As is shown in Table 52.11, the promulgated values for a few carbamates include permissible exposure levels (PELs), recommended exposure limits (RELs), immediately dangerous to life and health (IDLH) levels, and threshold limit values (TLVs) or time-weighted average (TWA) exposure levels for 8 to 10 hours. In addition, the U.S. Environmental Protection Agency (EPA) has introduced maximum drinking water contamination levels (MCLs) only for aldicarb (0.007 mg/L) and carbofuran (0.04 mg/L), chemicals that have been associated with problems of surface and groundwater contamination.
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CHAPTER 52
Carbamate Insecticides
REFERENCES Abdel-Wahab, A. M., Kuhr, R. J., and Casida, J. E. (1966). Fate of 14C_ carbonyl-Iabeled aryl methylcarbamate insecticide chemicals in and on bean plants. 1. Agric. Food Chem. 14, 290-298. Ahmed, E E., Lewis, N. J., and Hart, R W (1977). Pesticide-induced ouabain resistant mutants in Chinese hamster V79 cells. Chem. Biol. Interact. 19, 369-374. Aly, O. M., and E1-Dib, M. A. (1971). Studies on the persistence of some carbamate insecticides in the aquatic environment. 1. Hydrolysis of SevinTM, Baygon™, PyroIan™ and DimetiIan™ in waters. Water Res. 5,1191-1200. Anger, W. K., and Setzer, J. V. (1979). Effect of oral and intramuscular carbaryl administration on repeated acquisition in monkeys. 1. Toxicol. Environ. Health 5, 793-808. Anger, W. K, and Wilson, S. M. (1980). Effect of carbaryl on variable interval response rates in rats. Neurobehav. Toxicol. 2, 21-24. Araki, M., Yonemitsu, K, Kambe, T., Idaka, D., Tsunenari, S., Kanda, M., and Kambara, T. (1982). Forensic toxicological investigations on fatal cases of carbamate pesticide methomyl (Lannate®) poisoning. Ipn. 1. Legal Med. 36,584-588. Ashford, N., and Miller, C. (1998). "Chemical Exposures. Low Levels and High Stakes," 2nd ed. Van Nostrand Reinhold, New York. Baron, R. L. (1991). Carbamate insecticides. In "Handbook of Pesticide Toxicology" (W J. Hayes, Jr. and E. R Laws, Jr., eds.), Vol. 3, Ch. 17, pp. 11251189. Academic Press, New York. Barry, W K (1953). Peripheral neuritis following tetraethylthiuram-disulfide treatment. Brit. Med. 1. 2, 937. Berry, W. K (1971). Acceleration by free carbamate of the spontaneous reactivation of carbamylated acetylcholinesterase. Biochem. Pharmacol. 20, 3236-3238. Boyd, E. M., and Boulanger, M. A. (1968). Insecticide toxicology. Augmented susceptibility to carbaryl toxicity in albino rats fed purified casein diets. 1. Agric. Food Chem. 16, 834-838. Branch, R. A., and Jacqz, E. (l986a). Subacute neurotoxicity following longterm exposure to carbaryl. Am. J. Med. 80, 741-745. Branch, R A., and Jacqz, E. (1986b). Is carbaryl as safe as its reputation? Does it have a potential for causing chronic neurotoxicity in humans? Am. J. Med. 80, 659-664. Brugnone, E, Maranelli, G., Guglielmi, G., Ayyad, K., Soleo, L., and Elia, G. (1993). Blood concentrations of carbon disulfide in dithiocarbamate exposure and in the general population. Int. Arch. Occup. Environ. Health 64, 503-507. Carpenter, C. P., Weil, C. S., Palm, P. E., Woodside, M. W., Nair, rn, J. H., and Smyth, Jr., H. E (1961). Mammalian toxicity of I-naphthyl-Nmethylcarbamate (Sevin insecticide). J. Agric. Food Chem. 9, 30-38. Cavanagh, J. B. (1969). Toxic substances and the nervous system. Brit. Med. Bull. 25, 268-273. Cavanagh, J. B. (1973). Peripheral neuropathy caused by chemical agents. CRC Crit. Rev. Toxicol. 2, 365-417. Centers for Disease Control (CDC) (1979). Suspected carbamate intoxications-Nebraska. Morbidity and Mortality Weekly Report 28, 133-134. Centers for Disease Control (CDC) (1986). Aldicarb food poisoning from contaminated melons-California. Morbidity and Mortality Weekly Report 35, 254-255. Costa, L. G., Schwab, B. W, and Murphy, S. D. (1982). Tolerance to anticholinesterase compounds in mammals. Toxicology 25, 79-97. Cranmer, M. E (1986). Carbaryl: A toxicological review and risk analysis. Neurotoxicology 7, 247-328. Cremlyn, R (1978). "Pesticides. Preparation and Mode of Action." John WiIey and Sons, New York. Debuyst, B., and Van Larebeke, N. (1983). Induction of sister-chromatid exchanges in human Iymphocytes by aldicarb, thiofonax and methomyl. Mutat. Res. 113, 242-243. Desi, 1., Gonczi, L., Simon, G., Farkas, 1., and Kneffel, Z. (1974). Neurotoxicologic studies of two carbamate pesticides in subacute animal experiments. Toxicol. Appl. Pharmacol. 27, 465-476.
Dickoff, D. J., Gerber, 0., and Turovsky, Z. (1987). Delayed neurotoxicity after ingestion of carbamate pesticide. Neurology 37, 1229-1231. Dikshith, T. S. S., Gupta, P. K, Gaur, J. S., Datta, K K, and Mathur, A. K (1976). Ninety day toxicity of carbaryl in male rats. Environ. Res. 12, 161170. Ecobichon, D. J. (1982). Carbamic Acid Ester Pesticides. In "Pesticides and Neurological Diseases" (D. J. Ecobichon and R M. Joy, eds.), Ch. 6, pp. 220-221. CRC Press, Boca Raton, FL. Ecobichon, D. J. (l994a). Carbamic acid ester insecticides. In "Pesticides and Neurological Diseases" (D. J. Ecobichon and R. M. Joy, eds.), 2nd ed., Ch. 5, pp. 258-262. CRC Press, Boca Raton, FL. Ecobichon, D. J. (l994b). Organophosphorus ester insecticides. In "Pesticides and Neurological Diseases" (D. J. Ecobichon and R M. Joy, eds.), 2nd ed., Ch. 4, pp. 211-220. CRC Press, Boca Raton, FL. Ecobichon, D. J. (1996). Toxic effects of pesticides. In "Casarett and Doull's Toxicology. The Basic Science of Poisons" (c. D. Klaassen, ed.), 5th ed., Ch. 22, pp. 655-662. McGraw-HiII, New York. Ecobichon, D. J. (1999). Biological monitoring: Neurophysiological and behavavioral assessments. In "Occupational Hazards of Pesticide Exposures: Sampling, Monitoring, Measuring" (D. J. Ecobichon, ed.), Ch. 8, pp. 209230. Taylor and Francis, Philadelphia. Ecobichon, D. J., and Comeau, A. M. (1973). Pseudocholinesterases of mammalian plasma: Physicochemical properties and organophosphate inhibition in eleven species. Toxicol. Appl. Pharmacol. 24, 92-100. Ecobichon, D. J., and Joy, R. M. (1982). Carbamic acid ester pesticides. In "Pesticides and Neurological Diseases." CRC Press, Boca Raton, FL. Ecobichon, D. J., Ozere, R L., Reid, E., and Crocker, J. E S. (1977). Acute fenitrothion poisoning. Can. Med. Assoc. J. 116, 377-379. Edington, N., and Howell, J. M. (1966). Changes in the nervous system of rabbits following the administration of sodium-diethylthiocarbamate. Nature 210, 1060-1062. Edington, N., and Howell, J. M. (1969). The neurotoxicity of sodium diethylthiocarbamate in the hen. Acta Neuropathol. 12,339-347. Egert, G., and Greim, H. (1976). Formation of mutagenic nitroso-compounds from ephedrine and the pesticides carbaryl, dodin and prometryn in the presence of nitrite at pH 1. Naunyn-Schmiedberg 's Arch. Pharmacol. 293, Supp. R66. Ellman, G. L., Courtney, K. D., Andres, Jr., v., and Featherstone, R. M. (1961). A new and rapid colorimetric determination of acetylcholinesterase activity. Biochem. Pharmacol. 7, 88-95. Farage-Elawar, M. (1989a). Enzyme and behavioral changes in young chicks as a result of carbaryl treatment. J Toxicol. Environ. Health 26, 119-131. Farage-Elawar, M. (I 989b). Toxicity of aldicarb in young chicks. Neurotox. Teratol. 10, 549-554. Farago, A. (1969). Suicidal, fatal Sevin® (1-naphthyl-N-methyl carbamate) poisoning. Arch. Toxicol. 24, 309-315. Feldman, R G. (1999). Carbamates. In "Occupational and Environmental Neurotoxicology" (R. G. Feldman, ed.), Ch. 23, pp. 442-465. Lippincott-Raven Publishers, Philadelphia. Fiore, M. c., Anderson, H. A., Hong, Z., Golubjatnikov, R, Seiser, J. E., Nordstrom, D., Hanrahan, L., and Belluck, D. (1986). Chronic exposure to aldicarb-contaminated groundwater and human immune function. Environ. Res. 41, 633-645. Fisher, S. W, and Metcalf, R L. (1983). Production of delayed ataxia by carbamic acid esters. Pestic. Biochem. Physiol. 19, 243-253. Food and Agricultural Organization/World Health Organization (FAOIWHO) (1982). "Pesticide Residues in Food: 1981 Evaluations: The Monographs." FAO Plant Product Protection Paper No. 42. Food Agric. Organ. V.N., Rome. Fukuto, T. R (1972). Metabolism of carbamate insecticides. Drug Metab. Rev. 1, 117-150. Fukuto, T. R (1983). Structure-activity relationships in derivative of anticholinesterase insecticides. In "Pesticide Chemistry: Human Welfare and the Environment" (J. Miyamoto and P. C. Keamey, eds.), Vol. I, pp. 203213. Pergamon Press, New York.
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Fukuto, T. R., Fahmy, M. A. H., and Metcalf, R. L. (1967). Alkaline hydrolysis, anticholinesterase and insecticidal properties of some nitro-substituted phenyl carbamates. 1. Agric. Faad Chem. 15,273-277. Gardner-Thorpe, and Benjamin, S. (1971). Peripheral neuropathy after disulfiram adminstration. 1. Neural. Neurosurg. Psychiat. 34, 253-259. Goes, A. E., Savage, E. P., Gibbons, G., Arronson, M., Ford, S. A., and Wheeler, H. W. (1980). Suspected foodborne carbamate pesticide intoxications with the ingestion of hydroponic cucumbers. Am. 1. Epidem. 111, 254-259. Goldberg, M. E., Johnson, H. E., and Knaak, J. B. (1965). Inhibition of discrete avoidance behavior by three anticholinesterase agents. Psychapharmacalagia 7, 72-76. Goldman, L. R., Beller, M., and Jackson, R. J. (1990a). Aldicarb food poisonings in California. 1985-1988: Toxicity estimates for humans. Arch. Enviran. Health 45,141-147. Goldman, L. R., Smith, D. E, Neutra, R. R., Saunders, L. D., Pond, E. M., Stratton, J., Walker, K, Jackson, R. J., and Kizer, K. W. (l990b). Pesticide food poisoning from contaminated watermelons in California. Arch. Enviran. Health 45, 229-236. Gonzales, C. M., and Matos, E. (1984). Induction of sister-chromatid exchanges in cultured human lymphocytes by aldicarb, a carbamate pesticide. Mutat. Res. 138, 175-179. Grendon, J., Frost, E, and Baum, L. (1994). Chronic health effects among sheep and humans surviving an aldicarb poisoning incident. Vet. Human Taxical. 36,218-223. Guerzoni, M. E., DelCupolo, L., and Ponti, 1. (1976). Mutagenic activity of pesticides. Riv. Sci. Tecnal. Alimenti. Nutr. 6, 161-165. Harris, L. w., Talbot, B. G., Lennox, W. J., and Anderson, D. R. (1989). The relationship between oxime-induced reactivation of carbamylated acetylcholinesterase and antidotal efficacy against carbamate intoxication. Taxicol. Appl. Pharmacal. 98, 128-133. Hastings, E L., Main, A. R., and Iverson, E (1970). Carbamylation and affinity constants of some carbamate inhibitors of acetylcholinesterase and their relation to analogous substrate constants. 1. Agric. Faad Chem. 18, 497-502. Hayes, Jr., W. J. (1982). Carbamate pesticides. In "Pesticides Studied in Man" (w. J. Hayes, Jr., ed.), Ch. 8, pp. 436-462. Williams and Wilkins, Baltimore. Heath, D. E (1961). Abnormal effects. In "Organophosphorus Poisons. Anticholinesterases and Related Compounds," Ch. XVI, pp. 338-339. Pergamon Press, New York. Heise, G. A., and Hudson, J. D. (1985). Effects of pesticides and drugs on working memory in rats: Continuous delayed response. Pharmacal. Biachem.
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Behav.23,591-598. HoIIingshaus, J. G., and Fukuto, T. R. (1982). The effect of exposure to pesticides on delayed neurotoxicity. In "Effects of Chronic Exposures to Pesticides on Animal Systems" (J. Chambers and J. D. Yarborough, eds.), pp. 85-120. Raven Press, New York. Hoque, M. Z. (1972). Carbaryl, a new chemical mutagen. Curr. Sci. 41, 855856. Howell, J. M., and Edington, N. (1968). The neurotoxicity of sodium diethyldithiocarbamate in the hen. 1. Neuropathal. Exp. Neural. 27,464-472. International Program on Chemical Safety (IPCS) (1986). "Carbamate Pesticides: A General Introduction." Environmental Health Criteria 64. World Health Organization, Geneva. Ishidate, Jr., M., and Odashima, S. (1977). Chromosome tests with 134 compounds on Chinese hamster cells in vitro. A screening for chemical carcinogens. Mutat. Res. 48, 337-354. Iverson, E (1975). Affinity and carbamylation rate constants of propoxur in reaction with erythrocyte and serum cholinesterase. Biachem. Pharmacal. 24, 1537-1538. Johnson, D. J., Graham, D. G., Amamath, v., Amamath, K, and Valentine, W. M. (1998). Release of carbon disulfide is a contributing mechanism in the axonopathy produced by N,N-diethyldithiocarbamate. Taxical. Appl. Pharmacal. 148, 288-296. Knaak, J. B. (1971). Biological and nonbiological modifications of carbamates. Bull. W.H.O. 44,121-131. Krechniak, J., and Foss, W. (1982). Cholinesterase activity in rats treated with propoxur. Bull. Enviran. Cantam. Taxical. 29, 599-604.
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Kuhr, R. J., and Dorough, H. W. (1976). "Carbamate Insecticides: Chemistry, Biochemistry and Toxicology." CRC Press, Boca Raton, PL. Kulkami, A. P., and Hodgson, E. (1980). Metabolism of insecticides by mixed function oxidase systems. Pharmacal. Ther. 8, 379-475. Kurtz, P. J. (1977). Behavioral and biochemical effects of the carbamate insecticide, Mobam. Pharmacal. Biachem. Behav. 6, 303-310. Liddle, J. A., Kimbrough, R. D., Needham, L. L., Cline, R. E., Smrek, A. L., Yert, L. w., and Bayse, D. D. (1979). A fatal episode of accidental methomyl poisoning. Clin. Taxical. 15, 159-167. Lifshitz, M., Rotenberg, M., Sofer, S., Tamiri, T., Shahak, E., and Almog, S. (1994). Carbamate poisoning and oxime treatment in early children: A clinical and laboratory study. Pediatrics 93, 652-655. Lifshitz, M., Shahak, E., Bolotin, A., and Sofer, S. (1997). Carbamate poisoning in early childhood and in adults. Clin. Taxicol. 35, 25-27. Lima, J. S., Alberto, C., and Reis, G. (1995). Poisoning due to illegal use of carbamates as a rodenticide in Rio de Janeiro. Clin. Taxicol. 33, 687-690. Melnikov, N. N. (1971). Chemistry of pesticides. Residue Rev. 36,1-480. Merck Index, 12th ed. (1996). Merck Research Laboratories, Division of Merck and Co., Inc., Whitehouse Station, NJ. Miller, E., Reinwall, J., Brouwer, J., Ear, E L., and Loon, E. J. (1969). Effects of acute administration of carbaryl on cholinesterase levels in the CNS of swine. Taxical. Appl. Pharmacal. 14, 622-623. Moddel, G., Bilbao, J. M., Payne, D., and Ashby, P. (1978). Disulfiram neuropathy. Arch. Neurol. 35, 658-660. Moriya, M., Ohta, T., Watanabe, K., Miyazawa, T., Kato, K., and Shirasu, Y. (1983). Further mutagenicity studies on pesticides in bacterial reversion assay systems. Mutat. Res. 116, 185-216. Morse, D. L., Baker, E. L., Kimbrough, R. D., and Wisseman, C. L. (1979). Propanil-chloracne and methomyl toxicity in workers of a pesticide manufacturing plant. Clin. Taxical. 15, 13-21. Namba, T., Nolte, C. T., Jackrel, J., and Grob, D. (1971). Poisoning due to organophosphate insecticides. Am. 1. Med. 50, 475-492. Natoff, I. L., and Reiff, B. (1973). Effect of oximes on the acute toxicity of anticholinesterase carbamates. Taxical. Appl. Pharmacal. 25, 569-573. O'Malley, M. (1997). Clinical evaluation of pesticide exposure and poisonings. Lancet 349, 1161-1166. Onfelt, A. (1983). Spindle disturbances in mammalian cells. I. Changes in the quantity of free sulfhydryl groups in relation to survival and C-mitosis in V79 Chinese hamster cells after treatment with coIcemid, diamide, carbaryl and methyl mercury. Chem. Bial. Interact. 46, 201-217. Pilinskaya, M. A. (1981). Study of the cytogenetic effect of a number of pesticides in human peripheral blood Iymphocyte culture at various initial levels of chromosomal aberrations. Cytal. Genet. 15,74-76. Pilinskaya, M. A. (1982). The cytogenetic effect of pesticide pirimor in a human peripheral blood lymphocyte culture in viva and in vitro. Cytal. Genet. 16, 45-49. Rashid, K A., and Mumma, R. O. (1986). Screening pesticides for their ability to damage bacterial DNA. 1. Enviran. Sci. Health Part B 21, 319-334. Reiner, E. (1971). Spontaneous reactivation of the phosphorylated and carbamylated cholinesterases. Bull. W.H.O. 44, 109-112. Remaley, A. T., Hicks, D. G., Kane, M. D., and Shaw, L. M. (1988). Laboratory assessment of poisoning with a carbamate insecticide. Clin. Chem. 34, 1933-1936. Richardson, E. M., and Batteese, R. 1. (1973). An incident ofZectran poisoning. 1. Maine Med. Assac. 64, 158-159. Risher, J. E, Mink, E L., and Stara, J. E (1987). The toxicologic effects of the carbamate insecticide aldicarb in mammals: A review. Environ. Health Res. 72,267-281. Ruppert, P. H., Cook, L. L., Dean, K. E, and Reiter, L. W. (1983). Acute behavioral toxicity of carbaryl and propoxur in adult rats. Pharmacal. Biachem. Behav. 18, 569-584. Ryan, A. J. (1971). The metabolism of pesticidal carbamates. CRC Crit. Rev. Taxicol. 1,33-54. Sanderson, D. M. (1961). Treatment of poisoning by anticholinesterase insecticides in the rat. 1. Pharm. Pharmacal. 13, 435-442.
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Santolucito, J. A., and Morrison, G. (1971). EEG of Rhesus monkeys following prolonged low-level feeding of pesticides. Toxicol. Appl. Pharmacol. 19, 147-154. Schlagbauer, B. G. L., and Schlagbauer, A. W. J. (1972). The metabolism of carbamate pesticides-A literature analysis. Part I and Part n. Residue Rev. 42,1-84;42,85-90. Simpson, G. R., and Bermingham, S. (1977). Poisoning by carbamate pesticides. Med. 1. Austr. 2, 148-149. Singh, J. M. (1973). Decreased performance behavior with carbaryl-An indication of clinical toxicity. Clin. Toxicol. 6, 97-108. Smalley, H. E., O'Hara, P. J., Bridges, C. H., and Radeleff, RD. (1969). The effect of chronic carbaryl administration on the neuromuscular system of swine. Toxicol. Appl. Pharmacol. 14,409-419. Sterman, A. B., and Varma, A. (1983). Evaluating human neurotoxicity of the pesticide aldicarb: When man becomes the experimental animal. Neurobehav. Toxicol. Terato!' 5,493-495. Sterri, S. H., Rognerud, B., Fiskum, S. E., and Lyngaas, S. (1979). Effect of toxogonin and P2S in the toxicity of carbamates and organophosphorus compounds. Acta Pharmacol. Toxicol. 45, 9-15. Tabershaw, I. R., and Cooper, W. C. (1966). Sequelae of acute organic phosphate poisoning. 1. Occup. Med. 8, 5-20. Takahashi, RN., Poli, A., Morato, G. S., Lima, T. C. M., and Zanin, M. (1991). Effect of age on behavioral and physiological responses to carbaryl in rats. Neurotox. Terato!' 13,21-26. Tobin, J. S. (1970). Carbofuran a new carbamate insecticide. 1. Occup. Med. 12, 16-19. Umehara, E, Izumo, S., Arimura, K, and Osame, M. (1991). Polyneuropathy induced by m-tolyl methyl carbamate intoxication. 1. Neurol. 238, 47-48. Vandekar, M. (1965). Observations on the toxicity of carbaryl, folithion and 3-isopropoxyphenyl-N-methyIcarbamate in a village-scale trial in Southern Nigeria. Bull. WH.O. 33, 107-115. Vandekar, M., Heyadat, S., Plestina, R, and Ahmady, G. (1968). A study of the safety of O-isopropoxyphenyl-methyIcarbamate in an operational field trial in Iran. Bull. WH.O. 38, 609-623.
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CHAPTER
53 Aldicarb: Current Science-Based Approaches in Risk Assessment Abraham J. Tobia, Pierre-Gerard Pontal, Peter McCahon and Neil G. Carmichael Aventis CropScience
J oseph P. Rieth lSC, Inc.
Rick Williams RTI, Inc.
53.1 INTRODUCTION Toxicology-based human health risk assessment is evolving continuously. As new data are developed allowing greater understanding of chemical effects, dose relationships, and modes of action, improvements in the reliability of the subsequent risk assessments follow. This is particularly true in the case of pesticides, because this class of chemicals has undergone a virtually continuous process of registration and reregistration since the advent of the Environmental Protection Agency in the United States (U.S. EPA) as well as increased regulatory activity in Europe and Japan since the early 1970s. As a result, each currently registered pesticide has a very robust toxicological database from which one can assess potential health risks, and the wealth of information continues to grow as new study types are developed and conducted. With the advent of the Food Quality Protection Act of 1996 (FQPA) in the United States further emphasis has been placed on the quality of the risk assessment. Whereas previous risk assessment practices did indeed ensure the protection of infants and children, with FQPA emphasis is now placed on increasing the certainty of the assessment. A variety of new science-based techniques have been and are being developed to increase the certainty of the risk assessment process. This chapter describes new approaches, which reduce the uncertainty in the risk assessment of aldicarb. It begins with a description of aldicarb chemistry, uses, and biological mode of action. This is followed by a brief overview of risk assessment practices in the crop protection industry. The next Handbook of Pesticide Toxicology Volume 2. Agents
section describes the available toxicology and exposure studies used to evaluate the potential risk of the product. Finally, two elements of aldicarb risk assessment are presented. The first demonstrates the use of the pharmacokinetics of reversibility of cholinesterase inhibition following aldicarb exposure to adjust the exposure component of the risk assessment. The second describes special dermal toxicity studies used to evaluate the potential toxicity to occupational users of the aldicarb-formulated product Temik
53.2 ALDICARB: DESCRIPTION, USE AND BIOLOGICAL MODE OF ACTION Technical aldicarb belongs to the N -methyl carbamate chemical family. The pure (technical) material is a white crystalline solid with a water solubility of approximately 6000 ppm at 25°C and is stable at room temperature. The empirical formula of aldicarb is C7H14N202S, with a molecular weight of 190.3. The structural formula is shown in Fig. 53.1, and the (IUPAC) and Chemical Abstract Service (CAS) names are as follows: IUPAC-2-methyl-2-(methylthio)propionaldehyde O-(methylcarbamoyl) oxime CAS-2-methyl-2-(methylthio)propanalO-[(methylamino) carbonyl] CAS number-116-06-30 Aldicarb is a carbamate insecticide used in agriculture for the control of insects, mites, and nematodes. The product is mar-
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Copyright © 200 1 by Academic Press. All rights of reproduction in any fonn reserved.
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CHAPTER 53
Figure 53.1
Aldicarb Risk Assessment
Structural formula of aldicarb.
keted under the trade name Temik as a 15% (active ingredient) granule in the United States, and as a 5, 10, or 15% granule worldwide. Aldicarb has a number of unique characteristics which make it an invaluable tool for crop protection. First, aldicarb controls pests from three divergent animal groups: insects, mites, and nematodes. This range of activity for a single product is rare for this class of crop protection chemicals, and because of it, a single application of aldicarb can replace two or more applications of alternative pesticides. Second, aldicarb has systemic activity whereby the product can be applied beneath the soil, and uptake through the roots allows distribution throughout the plant, with subsequent control of chewing and sucking pests. Aldicarb is formulated solely as a dust-free granule and is not produced as a liquid formulation. This type of formulation significantly reduces potential dermal and inhalation exposures, which makes the product much safer from an occupational perspective. Granular Temik is also much safer for the environment than liquid-formulated insecticides. It affords longer control, reducing the number of applications. Also, it is applied beneath the surface of the soil to a depth of up to several inches. Both of these factors significantly reduce the negative impact on beneficial insects, fish, birds, and other wildlife because the product is not available for exposure. Aldicarb was discovered by Union Carbide Corporation in 1962. The first U.S. registration of Temik was received in 1970 for use on cotton. The four major crops on which Temik is currently used are citrus, cotton, peanuts, and potatoes, and it is also registered for use on nine other crops in the United States. Temik is typically applied via tractor-mounted equipment used to place the granules at a depth of 2-6 inches beneath the surface of the soil. Developments in positive displacement metering devices allow the application of precise amounts of the material into the application area. It is often applied in conjunction with other cultural practices such as planting, cultivating, or fertilizing. Most often, application is once per use season. Technical aldicarb is produced as an integrated 35% solution and then formulated into a granular product. Two types of granules are produced, one made of corncob grit and the other with gypsum clay. Both granules have binding agents and the production method produces a virtually dust-free product. Carbamate insecticides are reversible cholinesterase inhibitors for which recovery is primarily a function of the rate at which the active chemical is hydrolytic ally decarbamylated by the cholinesterase enzyme (Alvarez, 1992; O'Brien, 1967; Rotenberg and Almog, 1995). This process, commonly called spontaneous reactivation, is often measured in minutes. This
is in contrast to the organophosphates, which are generally irreversible inhibitors of cholinesterase; the inhibition is due to significantly stronger binding between the chemical and enzyme by the process of phosphorylation. Recovery following exposure to organophoshates is primarily through prolonged reactivation of the inactivated enzyme and synthesis of new enzyme, and is typically measured in days or weeks. It is primarily because of the difference in recovery times that carbamates are considered to pose significantly less risk to exposed humans, relative to organophosphate repeated exposure.
53.3 CURRENT PRACTICES IN PESTICIDE RISK ASSESSMENT The general process of human health risk assessment based on toxicological data is similar for all chemicals, that is, for pharmaceuticals, crop protection chemicals and industrial and naturally occurring chemicals. At a minimum, two types of information are required for a reliable risk assessment, knowledge of the "hazard" the chemical may cause, and data on possible levels of exposure to humans (Cohrssen and Covello, 1989). Hazard refers to the potential toxic effect or effects the material may cause; this information is typically gleaned either from descriptive toxicity tests in animals or from controlled studies in humans such as clinical trials or epidemiological studies (or a combination of both). Exposure data may be known exactly, as in the case of the prescribed dose of a pharmaceutical, or can be measured on those individuals coming into contact with the chemical, using a variety of techniques (U.S. EPA, 1987). In essence, human health risk assessment involves the determination of a level of exposure to a chemical which is expected to be safe for humans. This process is based on the well-established principles that for every toxic effect there must be a dose sufficiently large to cause that effect. Comparison of hazard and exposure data will not provide a safe exposure level, but can indicate if an exposure level is safe. Hazard information for registered pesticides is based on a very extensive toxicological database. Table 53.1 provides a listing of the types of toxicological studies conducted on aldicarb, which have been cited, reviewed and accepted by multiple international regulatory authorities [Baron, 1994; California EPA, 1998; Food and Agriculture Organization-World Health Organization (FAOIWHO), 1992; International Agency for Research on Cancer (IARC), 1991; International Programme on Chemical Safety (IPCS), 1991]. Studies are performed to examine the full spectrum of possible toxicological effects. These include acute, short-term, and long-term studies; tests for carcinogenicity and mutagenicity; reproductive and developmental effects; and neurotoxicity tests. The studies are conducted by various routes of exposure, for example, oral, dermal, and inhalation, and the studies are typically designed to approximate the route and duration of a potential human exposure.
53.3 Current Practices in Risk Assessment Table 53.1 Toxicological Studies Conducted on Aldicarb Acute Oral LDSO Dermal LDsO Inhalation LCSO Potential for eye irritation
1109
Table 53.2 Human Health Risk Assessments for Aldicarb Acute dietary exposure Chronic dietary exposure Short-term dermal occupational exposure Intermediate-term dermal occupational exposure Long-term dermal occupational exposure
Potential for skin irritation Potential for sensitization Subchronic 7-day dietary study in rats
Short-term inhalation occupational exposure Intermediate-term inhalation occupational exposure Long-term inhalation occupational exposure
90-day dietary study in rats 7-day dietary study in mice 14-day dietary study in dogs 90-day dietary study in dogs Chronic 2-year dietary study in rats (two studies) 2-year dietary study in rats with a mixture of aldicarb, aldicarb sulfoxide, and aldicarb sulfone Neurotoxicity Acute neurotoxicity study 90-day neurotoxicity study Developmental neurotoxicity study Human volunteer studies Controlled clinical studies (two studies) Worker exposure monitoring Studies on product formulations 21-day dermal study Oral LDso Dermal LDso Inhalation LCSO Potential for eye irritation Potential for skin irritation
The risk assessment is performed by first determining the human exposure scenario of interest and then selecting the toxicological study of most relevance to the human situation as well as the relevant end point in this study (usually the most sensitive). Thus, human health risk assessment involves a number of different risk assessments, which reflect each particular human exposure scenario. A listing of the types of risk assessments required for aldicarb can be found in Table 53.2. The potential routes of exposure for aldicarb include oral exposure, from low-level residues in food, and dermal and inhalation exposure of workers handling the product. Once the end point and study of interest for the appropriate risk assessment have been determined, the next step in the process is the determination of the no-observed-adverse-effect level (NOAEL). The NOAEL for the study is the dose level at which no hazard has been detected. Because these studies are generally conducted in animals, the NOAEL for a given study represents a safe dose for the species tested. The prediction of a safe dose for humans is derived by dividing the NOAEL from the animal study by an appropriate safety factor (also called
uncertainty factor); thus the safe human dose is lower than the NOAEL from the animal study. The safe dose for human intake is described in a number of ways, for example, a dose "without appreciable health risk" (WHO, 1987) and reference dose (RID; V.S. EPA, 1993), which was previously known as allowable or acceptable daily intake (ADI). The safety factor historically considered appropriate to provide a safe dose for human exposure for pesticides is 100 when it is derived from an appropriate animal study; typically the RID is set as the NOAEL from an appropriate animal study divided by 100. For excellent reviews of the origin and justification for the use of a lOO-fold safety factor in human health risk assessment see Swartout et al. (1998), Dourson et al. (1992), and Renwick and Lazarus (1998). The lOO-fold safety factor takes into consideration two sources of uncertainty in the risk assessment: (i) the toxicology study was conducted in animals but the objective is the protection of humans (interspecies extrapolation); (ii) there is variability in the human population and sensitive individuals need to be protected (intraspecies extrapolation). A lO-fold margin of safety is generally considered to provide adequate protection for each of these sources of uncertainty, hence the margin of safety of 100, which accounts for both sources simultaneously in the risk assessment. For the protection of infants and children, an additional safety factor with a default value of lOx has been mandated by FQPA (1996). When implemented, this will create a safety margin of 1000. Once the RID has been established, it can be compared to the expected exposure. If anticipated human exposures are less than the RID, then such exposures are considered safe. A related approach is to divide the NOAEL by the exposure value and calculate the margin of safety value directly. With this method, also called the margins of exposure (MOE) approach, values greater than 100 (or other numbers if considered appropriate) are considered adequate to protect human health. In cases where the initial risk assessment indicates that the exposure is higher than the RID, additional measures must be taken to ensure human safety. The risk assessment can be improved by reducing the uncertainty contained within it. This can be accomplished by developing new information, which allows a better understanding of the biological processes resulting in the hazard, and also by improving the knowledge of the exposures actually taking place. Direct measurement of the true
1110
CHAPTER 53
Aldicarb Risk Assessment
exposure levels through the conduct of a worker exposure study may demonstrate that the real exposure values are less than the estimated values. In addition exposure levels can be closely estimated from knowledge of related products for which exposure studies have been conducted. There are also ways in which improved toxicology information can be used to increase the accuracy of risk assessments by reducing the uncertainty. For example, if the mechanism of toxicity in the animal model can be demonstrated with certainty, and if it can be shown that this mechanism does not exist in humans, then a higher NOAEL based on another toxicity end point may be appropriate. This would have the effect of increasing the safety relative to the exposure value. Another way to reduce the uncertainty is to study the effects of the chemical in humans if human exposure will occur as a result of the product use. If the information used in the risk assessment has been determined in a well-conducted study in humans and demonstrates that there is concordance between the animal and human databases, then the safety factor used for the interspecies extrapolation is not necessary, and a safety factor of 10 (for the intraspecies extrapolation only) is considered adequate. Therefore, it is normal and appropriate practice to utilize the human study to augment the derivation a NOAEL and the RID. In the past, risk assessments were conducted primarily for long-term or lifetime exposure to pesticides. However, more recently, regulatory agencies (such as V.S. EPA) have also begun to conduct risk assessments for acute (1 day), short-term (1-7 days), and intermediate (7 days to 3 months) time periods for oral (dietary), dermal, and inhalation routes of exposure. The process is generally the same as described previously, except that the RID approach is used to assess chronic risk whereas the MOE approach is used for the short-term and intermediate assessments. There are other methodologies used to calculate the short-term assessments, but this chapter will focus mainly on the methodology utilized by the V.S. EPA. It should be noted that there are other risk assessment methods which do not rely on the NOAEL, such as the benchmark dose (BMD) approach, which describe the dose-response data mathematically; then a point-of-departure (POD) approach is used in which a predefined effect level, such as dose predicted to give a 10% effect (EDlO), is used rather than the NOAEL. The safety factor is then applied to the dose-response curve with the POD as the starting point for the hazard component of the risk assessment. These types of models have advantages in certain situations, for example, where no NOAEL has been established in a study. These models have not been widely used in a regulatory context because there is currently no scientific consensus to drive policy decisions.
as a developmental neurotoxIcIty study. Aldicarb has high acute toxicity. Toxicity is that commonly associated with acetylcholinesterase inhibition (ChEI) caused by a carbamate pesticide, that is, cholinergic symptoms. These symptoms are dose-dependent, are rapidly reversible, and do not occur at expected human exposure levels. Aldicarb is neither genotoxic nor carcinogenic. It does not cause developmental or reproductive effects in the absence of maternal toxicity. The degradation pathway for aldicarb involves a combination of oxidation to aldicarb sulfoxide and then aldicarb sulfone and hydrolysis of parent, sulfoxide, and sulfone, to low toxicity compounds.
53.4.2 ALDICARB ACUTE TOXICITY Tables 53.3 and 53.4 provide acute toxicity data for aldicarb technical and for the formulated product Temik 15G. Aldicarb technical is highly toxic by the oral, dermal, and inhalation routes. Aldicarb is not a sensitizer. Aldicarb sulfoxide has similar potency with regard to acetylcholinesterase inhibition as aldicarb itself. Aldicarb sulfone is approximately 25 times less toxic than either aldicarb or aldicarb sulfoxide.
53.4.3 ALDICARB SUB CHRONIC TOXICITY In assessing the subchronic toxicity of aldicarb, the most sensitive indicator of exposure is cholinesterase inhibition. A number of subchronic and subacute (e.g., 14-day) oral studies have been conducted on aldicarb, aldicarb sulfoxide, and aldicarb sulfone. Results from study to study are consistent; for the sake of simplicity, only the longer term oral studies and a 21-day dermal study are discussed. Table 53.5 provides a summary of these studies. In an oral study, rats were fed aldicarb in their diet for 93 days at dose levels of 0, 0.02, 0.1, and 0.5 mg/kg/day. The noTable 53.3 Acute Toxicity Data for Aldicarb Technicala Toxicity Species
Findings
Acute oral
Rat
LDSO = 1.2 mg/kg
Rabbit
LDSO = 544 mg/kg
Rat
LCSO = 0.0039 mg/l
Rabbit
Moderately irritating
III
Rabbit
Slightly irritating
IV
Guinea
Not a sensitizer
Not
toxicity Acute dermal
53.4.1 SUMMARY OF ALDICARB TOXICITY Aldicarb has a very robust toxicity database including developmental, reproductive, and neurotoxicity studies, as well
II
toxicity Acute inhalation toxicity Primary eye
53.4 ALDICARB TOXICOLOGY PROFILE
category
Study
irritation Primary dermal irritation Dermal sensitization
pig
applicable
a Aldicarb technical is approximately 35% aldicarb in dichloromethane.
53.4 Toxicology Profile
1111
Table 53.6 Acute Neurotoxicity Study in Rats: Cholinesterase Inhibition in FemalesPercentage Relative to Control, 0.75 h
Table 53.4 Acute Toxicity Data For Temik IS G Toxicity Study
Species
Findingsa
Acute oral
Rat
LDSO = 2.14 (m)/2.46 (t)
Rabbit
LDso > 2000 mglkg
Rat
Not applicable-
Rabbit
Moderately irritating
III
Rabbit
Not irritating
IV
Guinea
No data-technical
toxicity Acute dermal
mg/kg III
toxicity Acute inhalation
sensitization
0.05
46.6
8.6
5.1
0.1
73.3
30.6
15.6
0.5
94.1
54.2
50.4
(21-day dermal study)
irritation Dermal
Brain
Plasma
53.4.4 ALDICARB NEUROTOXICITY STUDIES
irritation Primary dermal
RBC
Dose (mg/kg)
granular product
toxicity Primary eye
% Inhibition
category
pig
is not a sensitizer
Not applicable
am = male; f = female.
observed-adverse-effect level was 0.5 mg/kg/day based on the lack of effects on red blood cell ChE. In addition, body weight and food consumption were decreased at the highest dose level. An oral dog study was conducted to investigate the ChEI dose-response curve of aldicarb. During the 5-week study, the dogs were fed diets mixed with aldicarb technical at levels of 0, 0.35, 0.7, and 2 ppm (0.013, 0.023, and 0.069 mg/kg/day in males, and 0.012, 0.025, and 0.067 mg/kg/day in females). There was neither mortality nor any changes in body weight, food consumption, or clinical observation data indicative of a compound effect. A NOAEL based on erythrocyte cholinesterase was established at 0.07 mg/kg/day. There have been a number of subchronic and subacute studies using aldicarb sulfoxide and aldicarb sulfone. As stated in the acute toxicity section, the sulfoxide metabolite is comparable in toxicity to aldicarb; the sulfone metabolite is less toxic. In each case, ChEI is the most sensitive indicator of exposure.
There is a complete neurotoxicity database on aldicarb consisting of acute, subchronic, and developmental neurotoxicity studies. In addition, there is a "time to peak behavioral effects" study of a single oral administration of aldicarb technical. Also, there are acute neurotoxicity studies on both aldicarb sulfoxide and aldicarb sulfone. As discussed earlier, effects on ChEI are always the most sensitive indicators of both exposure and toxicity in the case of aldicarb. The aldicarb dose-effect relationship for ChEI is consistent across studies. A dose of 0.05 mg/kg/day gives the first indications of erythrocyte cholinesterase inhibition with no concomitant brain cholinesterase inhibition or behavioral changes. At 0.2 mg/kg/day, marked erythrocyte ChEI is observed accompanied by measurable inhibition in the brain and moderate clinical signs. Higher dose levels result in marked erythrocyte and brain ChEI and clinical signs, the magnitude of which increases with dose. A summary of the ChEI effects from three aldicarb neurotoxicity studies is provided in Tables 53.6-53.8. In these tables, data on ChEI are shown for females only, to simplify the presentation, but the effects for males show the same magnitude of effects.
Table 53.5 Subchronic Toxicity Data for Aldicarb Dose Level
NOAEU
LOEU
Study
Species
(mg/kg/day)
(mg/kg/day)
(mg/kg/day)
Toxicological endpoints
Oral toxicity,
Rat
0,0.02,0.1,0.5
0.1
0.5
Plasma ChEI at the highest dose tested. In addition, food consumption and body weight were decreased at this
93-day
dose level. RBC ChEI was not affected at all dose levels. Oral toxicity,
Dog
5-week
0,0.35,0.7, and 2 ppm; 0, 0.013, 0.023,
0.023 (m), 0.025 (t)
0.069 (m), 0.067 (t)
and 0.069 mg/kglday (m);
Plasma ChE! was over 20% compared to controls at the highest dose tested. RBC ChE! was not affected at all dose levels.
0,0.012,0.025, and 0.067 mg/kg/day (t) Dermal toxicity,
Rat
0, 100, 250, 500
100
250
Plasma ChE! at 250 mg/kglday dose level. RBC NOAEL
21-day,
was 250 mg/kglday; brain ChE! NOAEL was at least
Temik 15G
500 mg/kglday.
am
= male; f = female.
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CHAPTER 53
Aldicarb Risk Assessment
Table 53.7 Time to Peak Behavioral Effects Study in Rats, Females-Percentage Relative to Predose Value, I h DOSE (mg/kg)
% Inhibition plasma
% Inhibition RBC
0.1
79.7
39.5
0.4
93.1
62.2
0.6
93.9
78.4
In an oral feeding developmental neurotoxicity study in rats, the dose levels were 0, 0.05, 0.1, and 0.3 mg/kg/day. This study provides strong evidence that aldicarb does not cause permanent effects on the nervous system, and that the young are not more sensitive to the effects of aldicarb than mature animals. The maternal NOAEL was 0.05 mg/kg/day based on miosis at 0.1 mg/kg/day. The developmental NOAEL was 0.05 mg/kg/day based on postweaning body weight decrement, reduced hindlimb grip strength, and foot splay in F] females on postpartum day 35. Cholinesterase (ChE) activity was measured in the maternal animals on gestation day 7, and lactation days 7 and 11. Cholinesterase inhibition was not detected at 0.05 mg/kg/day, probably because the blood was collected two hours postdose and the enzyme had spontaneously reactivated by this time. In a subchronic toxicity study, this same dose level resulted in 24% erythrocyte ChE!. Thus, the systemic toxicity NOAEL was 0.05 mg/kg/day and the maternal NOAEL for red blood cell (RBC) ChEI was similar. These results demonstrate the lack of increased sensitivity to developing animals relative to adults because there were no developmental effects even in the presence of slight maternal ChEL In an acute gavage study, rats were treated with aldicarb sulfoxide at doses of 0, 0.05, 0.1, and 0.5 mg/kg/day. Cholinesterase activity was not measured in this study. There were no deaths in the study and no significant effects on body weights or body weight gain. Food intake values were reduced for males in the 0.5 mg/kg/day dose. Significant functional observation battery (FOB) effects were seen at the 0.5-mg/kg/day dose level at the time of peak effect. Significant decreases in motor activity were also seen at this same dose. There were no neuropathology effects. The NOAEL for FOB and motor activity was 0.1 mg/kg/day; the NOAEL for histopathology was 0.5 mg/kg/day. Table 53.8 13-Week Neurotoxicity Study in Rats: Cholinesterase Inhibition Females at 4 Weeks Relative to Control, 0.75 h
In an acute neurotoxicity gavage study in rats, the dose levels for aldicarb sulfone were 0, 1, 10, and 20 mg/kg/day. Cholinesterase activity was not measured in this study. No deaths occurred in the study. At the 20-mg/kg/day dose level, males and females showed significant decreases in food consumption, and males exhibited a significant reduction in body weight gain. At the time of peak effect, significant FOB effects were seen at the 10- and 20-mg/kg/day dose levels. Neuropathological evaluations revealed no effects at any dose. The NOAEL for FOB and motor activity was 1 mg/kg/day; the NOAEL for histopathology was 20 mg/kg/day.
53.4.5 ALDICARB DEVELOPMENTAL AND REPRODUCTIVE TOXICITY STUDIES
There is a complete developmental and reproductive toxicity database including a developmental neurotoxicity study (discussed in the previous section). Aldicarb does not cause developmental or reproductive effects in studies in the absence of maternal (or parental) toxicity. The following section discusses the study results, and these are summarized in Table 53.9. In an oral gavage developmental study, rats were given doses of 0, 0.125, 0.25, and 0.5 mg/kg/day. Maternal toxicity was indicated by maternal death and clinical signs were observed at the upper dose levels. Gestational parameters were not affected. No increased incidence of malformations was observed in the absence of clear maternal toxicity. The NOAEL for fetal toxicity was 0.25 mg/kg/day; fetal effects at the highest dose included dilated ventricles. In an oral gavage rabbit developmental study with doses of 0, 0.1,0.25, and 0.5 mg/kg/day, there were no fetal effects. Maternal toxicity was clearly established in the upper two dose levels with an increase in severity being observed at the highest dose level tested. The maternal NOAEL was 0.1 mg/kg/day based on decreased body weight and clinical signs. In a two-generation reproductive toxicity study, rats were fed a diet with 0,2,5, 10, or 20 ppm (ca. 0, 0.1, 0.25, 0.5, or 1.0 mg/kg/day). Cholinesterase inhibition and body weight changes in parents were observed at the upper dose levels. The maternal NOAEL was 0.25 mg/kg/day. The reproductive NOAEL was 0.5 mg/kg/day based on decreased pup weight and reduced viability. There were no reproductive effects in the absence of parental toxicity.
53.4.6 ALDICARB MUTAGENICITY STUDIES % Inhibition Dose (mg/kg/day)
Plasma
0.05
64.9
24.0
2.1
0.2
92.7
71.1
33.1
0.3
94.9
70.3
56.7
RBC
Brain
Studies covering gene mutations, chromosomal aberrations, unscheduled DNA synthesis, and dominant lethal effects were all negative for aldicarb. There is no concern for mutagenicity for aldicarb. A limited battery of genotoxicity studies on aldicarb sulfoxide and sulfone are also negative.
53.4 Toxicology Profile
1113
Table 53.9 Developmental and Reproductive Toxicity Data for Aldicarb NOAEL
LOEL
Study
Species
Dose level
(mg/kg/day)
(mg/kg/day)
Toxicological endpoints
Developmental
Rat
0,0.125,0.25,
Maternal toxicity, 0.125;
Maternal,0.25;
Maternal toxicity was indicated by death, reduced
toxicity
fetal toxicity, 0.25
0.5 mg/kg/day
body weight gain and food consumption, and
fetal,0.5
clinical signs of cholinesterase inhibition at 0.5 mg/kg/day, and reduced food consumption and body weights at 0.25 mg/kg/day. Fetal toxicity was indicated by increased dilated ventricles and reduced ossification of the 6th sternebrae. Developmental
Rabbit
toxicity
0,0.1,0.25,0.5
Maternal toxicity, 0.1;
Maternal, 0.25;
mg/kg/day
fetal toxicity: 0.5
fetal, >0.5
Signs of maternal toxicity included pale kidneys, hydroceles on oviducts, and decreased body weight. There was no fetal toxicity.
Two-generation
Rat
0,2,5,10,20
Parental,0.25; reproductive, 0.5
Parental: 0.5;
Maternal toxicity included plasma and RBC Cheri
reproductive
ppm; ca.
toxicity
0,0.1,0.25,0.5,
The reproductive LOEL is based on decreased pup
1.0 mg/kg/day
weights and reduced viability.
53.4.7 ALDICARB CHRONIC TOXICITY AND ONCOGENICITY STUDIES
Aldicarb has been shown to have no oncogenic potential when administered to rats and mice in lifetime experiments. Cholinesterase inhibition is the most sensitive indicator of ex-
reproductive, 1.0
and body weight changes at the upper dose levels.
po sure in chronic studies in rats and dogs. A discussion of chronic toxicity and oncogenicity data follows and Table 53.10 summarizes the study results. In a 2-year study, rats were fed aldicarb at levels of 0, I, 10, and 30 ppm in the diet (equivalent to ca. 0,0.05,0.5, and 1.5 mg/kg/day). There were no compound-related effects on sur-
Table 53.10 Chronic Toxicity and Oncogenicity Data for Aldicarb NOAEU
LOELa
Study
Species
Dose levels
(mg/kg/day)
(mg/kg/day)
Toxicological endpoints
Chronic toxicity or
Rat
0, I, 10,30 ppm;
0.05 (m), 0.59 (f)
0.5 (m), 1.5 (f)
Highest dose tested equivalent to greater than an LDSO
oncogenicity
ca. 0, 0.05, 0.5, 1.5
dosage when administered by gavage. Only clinical effect
mg/kg/day
was limited use of the tail at the highest dose tested. Body weights and body weight gains reduced at this dose level. Atrophy of the iris at the high dose.
Oncogenicity
Rat
0,2,6ppm;
Not evaluated
Not evaluated
None.
Not evaluated
Not evaluated
There were slight increases in mortality and slight
ca. 0,0.1,0.3 Chronic toxicity or
Rat
oncogenicity
0, 0.3 mg/kg/day for aldicarb; other
depressions in growth at certain stages for some of
doses for other
the test materials.
materials (e.g., a1dicarb sulfoxide) Oncogenicity
Mouse
0,0.1,0.3,0.7
Not evaluated
Not evaluated
Mortality and an increase in hematomas and lymphoid
Not evaluated
Not evaluated
None.
0.027
0.055
Plasma ChEI occurred at 0.055 mg/kg/day. Brain ChEI
mg/kg/day Oncogenicity
Mouse
0, 2, 6 ppm; ca. 0,
neoplasia were observed at the highest dose tested.
0.29, and 0.86 mg/kg/day Chronic toxicity
Dog
0, 1,2,5, 10 ppm; ca. 0, 0.027, 0.055, 0.13 mg/kg/day
am = male; f = female.
occurred at 0.13 mg/kg/day.
1114
CHAPTER 53
Aldicarb Risk Assessment
vival. It should be noted that the high dose of 1.5 mg/kg/day was greater than the LDso and was tolerated every day over the course of the study. This was possible because the aldicarb was administered via the diet, and the total dose was ingested in fractionated amounts throughout the day, allowing for ChEI reversibility between consumption periods. The principal clinical effect observed was limited use of the tail in high-dose males and females. Body weights and body weight gains were reduced in high-dose males and females. Also, atrophy of the iris occurred in this dose group. There was no evidence of direct organ toxicity, and no evidence of oncogenic effects. The NOAELs were 0.05 mg/kg/day in males and 0.59 mg/kg/day in females based on erythrocyte ChE!. It is noteworthy that the rats could tolerate such a dosing regimen over their entire lifespan, demonstrating that recovery is complete, accumulation of aldicarb does not occur, and there are no persistent effects following such exposure. In a National Cancer Institute (NCI) study, rats were fed aldicarb in the diet at concentrations of 0, 2, and 6 ppm (equivalent to ca. 0, 0.1, and 0.3 mg/kg/day). There was no mortality attributed to aldicarb and no effect on body weight was noted. It was concluded that aldicarb was not oncogenic; the NOAEL was the highest dose tested. In a third rat study, rats were fed aldicarb at dose levels of 0 and 0.3 mg/kg/day. In addition, other groups were fed aldicarb sulfoxide at dose levels of 0, 0.3, and 0.6 mg/kg/day, aldicarb sulfone at dose levels of 0,0.6, and 0.24 mg/kg/day, or a mixture of aldicarb sulfoxide and aldicarb sulfone at doses of 0, 0.5, and 1.2 mg/kg/day. Neither aldicarb nor either of its major metabolites was found to be oncogenic. There were slight increases in mortality and slight depressions in growth at certain stages for some of the test materials. Cholinesterase activity was measured at 6, 12, and 24 months during the study. Plasma, erythrocyte, and brain ChE activity were examined 24 hours after animals were removed from test diets. No ChEI was noted other than a possible slight inhibition with respect to plasma ChE. There have been three mouse oncogenicity studies conducted on aldicarb. The first is a National Cancer Institute study in which mice were fed 0, 2, or 6 ppm of aldicarb in the diet (equivalent to ca. 0, 0.29, and 0.86 mg/kg/day). It was concluded that aldicarb was not oncogenic. No effects on mortality or body weights were noted. In a second study, mice were fed aldicarb at doses of 0, 0.1, 0.2,0.4, and 0.7 mg/kg/day. Mortality was evident in males at the two highest dose levels, and in females at the three highest dose levels during the first few months of the study. Following this period, aldicarb was mixed in the diet in a different manner which eliminated the acute toxicity. Based on the mortality observed in the study, these data are not considered appropriate for the evaluation of an oncogenic response. In a third study, mice were fed aldicarb at dose levels of 0, 0.1,0.3, and 0.7 mg/kg/day. There was no effect on mortality or growth. Inclusion of aldicarb in the diet did not result in an increased incidence of oncogenic response. In a one-year study in dogs, groups of beagles were fed dietary concentrations of 0, 1,2, 5, and 10 ppm daily for 52 weeks
(equivalent to ca. 0, 0.027, 0.055, 0.13, and 0.24 mg/kg/day). The study was designed to produce maximum ChE! by limiting feeding time to 2 hours per day to mimic a bolus administration of aldicarb. Cholinesterase activity was measured from blood samples approximately 2 hours after the feeding period. There were no observable effects other than ChE!. The NOAEL for erythrocyte ChEI was 0.027 mg/kg/day. The lowest observed effect levels (LOELs) for erythrocyte and brain ChEI were 0.055 and 0.24 mg/kg/day, respectively. In another one-year dog feeding study, aldicarb sulfone was administered at dietary concentrations of 0, 5, 25, and 100 ppm (ca. 0, 0.125, 0.625, and 2.5 mg/kg/day). Cholinesterase determinations were taken approximately 2 hours after feeding to measure maximum ChE!. No mortality or treatment related clinical signs were seen. At the high dose, slight changes in spleen and thyroid-parathyroid weights and slight effects in the mandibular lymph nodes and adrenal cortex were observed. The NOAEL based on erythrocyte ChEI was 0.625 mg/kg/day. 53.4.8 HUMAN VOLUNTEER STUDIES
In a series of studies reported in 1973, groups of four adult male volunteers were administered aldicarb orally in aqueous solution at dose levels of 0.025,0.05, and 0.1 milligram per kilogram of body weight (mg/kg bw). Clinical signs were recorded and whole blood cholinesterase activity was measured up to 6 hours after administration of the sample. Total urine voided was collected and aldicarb-excretion patterns for the initial 8 hours after dosing were evaluated. In addition, spot samples were taken at 12 and 24 hours. In other studies, two additional subjects were administered aldicarb in water solution at dose levels of 0.05 and 0.26 mg/kg bw. Dose levels of 0.1 and 0.26 mg/kg bw are considered to be high doses. Acute signs, typical of anticholinesterase agents, were observed at the high-dose levels (0.1 and 0.26 mg/kg bw) within 1 hour of aldicarb administration. Cholinesterase depression at these very high dose levels was observed in all volunteers within 1-2 hours after treatment. Within the first 6 hours of treatment, cholinesterase depression and clinical signs of poisoning were within normal levels. There were no signs of treatment observed at the 0.05-mg/kg bw dose level. Urine analysis showed that approximately 10% of the administered dose was excreted as carbamates within the first 8hour interval. Cholinesterase analyses confirmed the same rapid inhibition and recovery pattern with man as had been observed in experimental animals. In 1992 Aventis Crop Science conducted a human volunteer study, which was conducted and performed under globally accepted ethical guidelines established for such work. This was a double blind, placebo controlled study, in which aldicarb was administered as a single oral dose to healthy male and female subjects. The doses administered were: placebo (22 subjects16 males and 6 females); 0.01 mg/kg bw (8 males); 0.025 mg/kg bw (8 males and 4 females); 0.05 mg/kg bw (8 males and 4 females); and 0.075 mg/kg bw (4 males). Volunteers were screened before entry for general medical history by examination and laboratory tests including hematology, clinical
53.5 Use of Pharmacokinetics in Aldicarb Risk Assessments chemistry, and urinalysis. Clinical measurements were made at intervals before and after dosing. These included vital signs (systolic and diastolic blood pressure, pulse rate), pulmonary function tests, pupil size, electrocardiographs (ECGs), salivation, and clinical signs of nausea, vomiting, diarrhea, sweating, abdominal cramps, involuntary movement, or slurred speech. Samples were taken for urinalysis, clinical chemistry (including red blood cell and plasma cholinesterase activity), and hematology before and after dosing. There were no clinically significant changes in vital signs, pupil size, pulmonary function, ECGs, salivation, clinical signs, clinical chemistry (apart from cholinesterase), hematology, or urinalysis in the study. Cholinesterase activity was the only parameter affected during the study. Red blood cell and plasma cholinesterase was maximally depressed at 1 hour after dosing and had recovered by 8 hours in all subjects. The fall in cholinesterase activity and recovery was dose-related. No biologically significant depression of erythrocyte cholinesterase activity (> 20%) was seen in subjects treated with 0.01 or 0.025 mg/kg bw or in plasma cholinesterase at 0.01 mg/kg bw. Depression in cholinesterase activity > 20% was seen in erythrocytes at 0.05, and 0.075 mg/kg bw and in plasma at 0.025, 0.05, and 0.075 mg/kg bw. A single volunteer (0.075-mg/kg bw group, actual dose 0.06 mg/kg bw) reported some sweating, which is not considered to be related to aldicarb. Nevertheless, the NOAEL for clinical signs was reported as 0.05 mg/kg bw and the NOAEL based on erythrocyte cholinesterase inhibition was 0.025 mg/kg bw.
53.5 USE OF PHARMACOKINETICS IN ALDICARB RISK ASSESSMENTS Although it has been known since the initial discovery of the carbamates that this class of compounds present reduced risk compared to organophosphates due to the rapid reversibility of cholinesterase inhibition, a quantitative analysis of ChE reversibility is desirable because it allows the conduct of more precise risk assessments taking into account the exposure patterns (route of exposure, duration, and repetition). The kinetics of reversibility of cholinesterase inhibition for aldicarb has been studied through the development of a mathematical model, which describes the reversal of inhibition using data generated during the 1992 study in human volunteers. As previously described, healthy male and female volunteers were administered aldicarb in a study to evaluate its effects on plasma (PChE) and red blood cell (RChE) cholinesterase activity. Cholinesterase activity was measured at three pretreatment time points ( -16 and - 3 hr, and immediately predose) and 1,2, 4, 6, 8, and 21 hr post dose. Measurement of PChE and RChE activity at these time points allowed the determination of the maximum inhibitory effect for each dose level as well as the time required for spontaneous reactivation of the enzyme activity through hydrolysis of the carbamate. In addition, samples were taken for urinalysis, clinical chemistry, and hematology evaluations.
1115
The data presented in this chapter will focus on the male RChE data because these are more robust (a larger number of male than female subjects were observed), and because RChE data are generally considered to have more biological relevance than PChE in regard to potential peripheral and central nervous system effects (Carlock et aI., 1999; V.S. EPA, 2000). A graphical representation of the male RChE data in the volunteer study is shown in Fig. 53.2. The data show a very stable baseline activity level and rapid decrease in RChE activity following exposure to the bolus dose. The RChE inhibition is then subsequently reversed during a time period ranging from minutes to approximately 2 hours depending on the magnitude of the initial inhibition event (i.e., the smaller the inhibition, the faster the recovery time). The mathematical model was developed to quantitatively describe the data from the study. With such a model, one can predict the degree of cholinesterase inhibition which would be expected to occur following exposure to a dose that had not been tested experimentally. Additionally, one can calculate the time needed to completely reverse the effects of cholinesterase inhibition following exposure to a given dose. The approach presented here is based on the knowledge that the clinical determination of the reversal of cholinesterase inhibition following exposure to a carbamate is in effect the measurement of the removal of the carbamate from the body. This is so because the reactivation of the cholinesterase enzyme occurs via hydrolysis of the carbamate; hydrolysis results in the production of inactive metabolites and the full restoration of the functional capability of the enzyme. If the enzyme is not inhibited, then there is no longer free aldicarb in the blood compartment available for binding to the enzyme. Thus a mathematical description of the reactivation of cholinesterase activity following cholinesterase inhibition due to carbamate exposure can be regarded as the inverse of the model for removal of the carbamate. The latter model can be used in risk assessments for human health issues surrounding carbamate exposure because, when the initial exposure level is known, the amount remaining after a given point in time can be calculated. In general terms, the model is f(tUPI,
>2, 1>3)
= 1>d 1 -
1>2 exp[ -1>3(t -
1)]}
where t is time in hours after exposure (t ~ 1), 1>1 is the horizontal asymptote that is approached over time (t -+ 00), 1>1 (l - 1>2) is the minimum value at 1 hour, and 1>3 is a scale parameter related to the rate of recovery. The value lOO1>2 is the percentage reduction from the asymptotic maximum value to the minimum value at time 1 hour. In biological terms, 1>1 is the baseline cholinesterase activity level, 1>2 is a parameter describing the degree of inhibition as a function of dose, and 1>3 describes the rate of recovery of enzymatic activity (Williams et aI., 2000). The model belongs to the general class of nonlinear mixed effects growth curve models containing both fixed and random effects. The fixed effects permit estimation of the average curves describing the cholinesterase activity levels following exposure to a given dose; the random effects account for individual variability observed in a repeated measures study of
1116
CHAPTER 53
Aldicarb Risk Assessment
Activity (mU/mL)
it.:·.S
~:4.?
13. i
--------------.--~.(!! -~~
~
:\
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'
-
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i
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-20
o
·30 Time (hrs)
Figure 53.2
Male red blood cell cholinesterase activity from the human volunteer study on aldicarb.
this type. Although the model was developed and validated empirically, it can be thought of as a mechanistic model in the sense that assumptions regarding what was known about the biological system were used in its development. For example, the degree of cholinesterase inhibition was assumed to be a function of the exposure dose, and it was also assumed that cholinesterase returned to predose activity levels over time. Both of these assumptions were tested by the model and validated. The rate of change (as cholinesterase activity returned to baseline levels following inhibition) predicted from the model was calculated as follows. The rate of change, or the first derivative, of the predicted activity for a particular person is a function of the person's normal cholinesterase activity. The basic model used for activity level is
The models for the
+ U1l
where mi is the typical, nonexposed, cholinesterase activity for person i, fh is the proportion of the preexposure activity, and u 11 is a random effect specific to session I, which a person should reach for the recovery to be considered complete. To test time to recovery, one must specify what percentage of the
preexposure activity is considered the normal state, if f31 is estimated in the model to be 1, and then recovery has completely returned to predose levels. Finally, U I I accounts for random effects associated with person i. For the magnitude of the inhibition,
= exp[f32 + U2I ]di
where f32 is a fixed value for a typical subject, U2I is a random effect specific to session I, and di is the dosage person i ingests. Thus the degree of inhibition is a simple multiple of the dose to which the person is exposed. Finally, for the rate of recovery,
where f33 and f34 are fixed values for each subject. f34 was included in the model to test if there was an effect of dosage on the rate of recovery; the model was found to be very close to 1, indicating that the rate of recovery was independent of doseestimated f34; U31 is a random effect specific to session I. Thus, taking the derivative of the full model gives
This shows that the rate of change in activity after exposure is a function of the normal cholinesterase activity particular to person i (mi), and that the derivative changes with time (t). The
53.5 Use of Pharmacokinetics in Aldicarb Risk Assessments Recovery
Time (hrs) 9
7 ---------------------------------------------------------------------------------------------------------------
5 ---------------------------------------------------------------------------------------------- ------- --------
3 2
1117
considered separate events. If, however, the effect has not completely reversed because a certain amount of the dose remains in the system, then the second dose may be additive with however much of the first dose remains. The steps in establishing whether or not two separate exposures can be considered separate events are as follows: • Determination of the critical toxicological effect • Development of a method to measure the critical toxicological effect • Establishing that the critical toxicological effect does in fact reverse completely • Demonstrating the time to reversal In the case of aldicarb, the existing toxicological database provides clear answers to the first three of these questions, and the model provides the time to reversal.
0.02
0.03
0.04
0.05
0.00
007
0.08
53.5.2 CRITICAL EFFECTS
Dosoge (mg,Kg)
Figure 53.3 Model predicted recovery curves based on the male red blood cell cholinesterase activity from the human volunteer study on aldicarb.
model was validated using two methods for variance estimation, by comparing observed versus expected values directly and through residuals, and was shown to have a good fit to the experimental values. Model-predicted recovery curves are shown in Fig. 53.3. The model demonstrated that there was full recovery following exposure to aldicarb. This was evaluated by examination of the parameter fh in the model. /31 as predicted by the model was virtually identical to 1, and /31 is a multiplier of the baseline value; any deviation from return to baseline would have been indicated by a model-predicted /31 being different from 1. This was predicted by simple examination of the ChE activity graphs following aldicarb exposure, and has now been validated both mathematically and statistically. The model also demonstrated that recovery rate was not affected by administered dose; the rate of recovery for all doses was the same. This again was predicted because first-order kinetics would be presumed for the hydrolysis reaction between the cholinesterase enzyme and the aldicarb substrate. Because the rate of recovery is not affected by dose, the time to recovery is dependent solely on the initial magnitude of inhibition (i.e., the smaller the dose, the lower degree of inhibition, and thus the faster the recovery time). 53.5.1 REPEATED EXPOSURE MODELING
The time to recovery for the critical toxicological effect is important in the evaluation of the effects of multiple exposures to a chemical because if the effect has completely reversed prior to a subsequent exposure, then the two exposures can be
The critical toxicological effect is inhibition (i.e., ChEI), which is not in fact a toxic effect, but a marker of exposure. The toxicological studies described in this review covering carcinogenic, reproductive, neurotoxic, genotoxic, and chronic effects generated no effects other than those which are typically associated with acetylcholinesterase inhibition, as a marker of exposure. The most sensitive indicator of acetylcholinesterase inhibition is measurement of RBC cholinesterase activity. The detailed neurotoxicity studies conducted on aldicarb show that the first indicator following exposure is a decrease in RBC ChE activity. At higher doses, effects on the central and peripheral nervous systems become evident. The studies also demonstrate that the degree of central nervous system (CNS) ChEI for a given dose is less than that measured in RBCs. Thus measurement of RBC ChE activity provides the most sensitive indication of exposure to aldicarb. In addition, RBC ChE activity is readily measurable with a variety of analytical techniques. 53.5.3 REVERSIBILITY
The fact that the reversal of RBC ChE inhibition is complete was demonstrated empirically with a mathematical model. The toxicological database clearly demonstrates no persistence of aldicarb. For example, it would not have been possible to dose every day at such high levels in the 2-year rat study were there any cumulative effects of aldicarb. Also, these animals received multiple doses within each day because they were administered aldicarb ad libitium via the diet and rats eat sporadically during the night. Similarly, the two-generation reproduction, the developmental neurotoxicity, and the subchronic neurotoxicity studies all used repeated daily dosing and gave no evidence of any cumulative effects. Thus it has been adequately demonstrated that the effects of aldicarb exposure are the same either following repeated within-day dosing or following administra-
1118
CHAPTER 53
Aldicarb Risk Assessment
1000
:::2:
900
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+1
::::J
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:s
800
S
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~
:~
t5
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en
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en Q)
___ 0 mg/kg (n=5-6) -V- 0.05 mg/kg (n=6-7) _ 0.10 mg/kg (n=7)
600
.~
"0
..c
o oCO
er:
500
* 400~~~~-L~L---~---L--------~------~----------------~------
-1
20
40
60
Dose (0 min)
90
120
180
240
360
Time (minutes)
Figure 53.4 Phase I-mean RBC cholinesterase activity in cannulated adult male CD rats following a single oral administration of aldicarb: (*) significantly different from control within time point (p < 0.05).
tion over multiple days up to and including the lifespan of the animals. More recently, a pharmacokinetic reversibility study was completed using single and multiple dosing of aldicarb in rats to further validate the reversibility. The purpose of this study was to determine the time course of inhibition and recovery of cholinesterase activity after a single or repeated administration of aldicarb for both plasma and RBC. Aldicarb was orally administered once in Phase I and twice in Phase 11 to cannulated adult male CD® rats at three dose levels (0, 0.05, and 0.10 mg/kg). Red blood cell cholinesterase was inhibited lO minutes after oral administration of aldicarb in both dose groups and had returned to baseline levels within 180 minutes (see Fig. 53.4). In Phase 11, two consecutive administrations at 4-hour intervals, RBC cholinesterase activity was inhibited after the first dose in a manner similar to the single administration study. Cholinesterase returned to baseline levels of activity by 120 minutes for the low-dose group and 240 minutes for the highdose group (see Fig. 53.5). Following the second dose, approximately 4.5 hours after the first, inhibition occurred through 40 and 120 minutes for the low- and high-dose groups, respectively. When comparing the cholinesterase inhibition curves following the first and second dose administration for each dose group, no significant differences were detected. Thus, the
pattern of RBC cholinesterase inhibition following a second administration of aldicarb was not different from the single administration. As was shown in the chronic study, no accumulation was observed in this acute pharmacokinetic reversibility study. To assess the recovery time following aldicarb exposure, it was of interest to know the time to recovery of ChEI at dose levels that a human might possibly be exposed to. The model was used to answer the question: What is the time to recovery for an individual exposed to the V.S. EPA reference dose (0.001 mg/kg/day)? Recovery is defined as the return to 99% of the individual's baseline RBC ChE level. The RID was chosen because this dose would be expected to give the longest recovery time within the range of regulatory relevant dose levels. Return to 99% of baseline was chosen because this is a conservative cutoff criteria insofar as normal RBC ChE levels can vary as much as or more than 10% or in nonexposed individuals. The model predicted that recovery time for someone exposed to the full reference dose is instantaneous. This is because no deviation from baseline ChE activity is predicted at the RID. This is, of course, what one would expect given that the RID was set at a level lO-fold lower than the human NOAEL for the effect. The conclusion to be drawn is that, for acute dietary exposure to aldicarb where all exposures are less than the ref-
53.6 Aldicarb Dermal Exposure Risk Assessment
- . - 0 mg/kg (n=6) -V- 0.05 mg/kg (n=6) _ 0.10 mg/kg (n=7)
1000
~
1119
900
ill
en +1
::J E
:3
800
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~ .;;:
«U
700
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ro
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c
600
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o
o CO a:
500
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*
400U-~~~~-----L------------L--.~~~------L-----~-----L----
-15
120 Dose 1 (0 min) Curve A
240
20 40 60
120
180
240
Dose 2 (270 min) Curve B
Post-dose Time Points (minutes) Figure 53.5 Phase II-mean RBC cholinesterase activity in cannulated adult male CD rats following repeated oral administration of aldicarb: (*) significantly different from control within time point (p < 0.05).
erence dose, each meal should be considered a separate eating event for the dietary risk assessment. Any aldicarb exposure encountered in a given meal will be removed from the body prior to any subsequent exposure. Thus acute risk assessments for aldicarb should be based on the amount received in any individual meal and the exposures from meal to meal should not be added together over the course of a 24-hour day.
53.6 ALDICARB DERMAL EXPOSURE RISK ASSESSMENT Aldicarb is formulated solely as a granular product thus greatly reducing the possibility of occupational exposure. A hazard assessment for potential dermal exposure of agricultural workers has typically been conducted one of two ways. These assessments are based on subchronic toxicity studies conducted by either the oral or the dermal route of exposure; in both cases the studies are normally conducted with the technical material.
When an oral study is used for a dermal risk assessment, the oral NOAEL can be converted to dermal equivalents through the use of a dermal absorption factor. This is typically determined experimentally by using radiolabeled test material applied to the skin and measuring the levels of radioactivity, which has been absorbed through the skin, at varying time intervals. The percentage absorbed can then be used to modify the dose from the oral study. This first approach, however, is not appropriate for aldicarb because the compound is rapidly hydrolyzed in the body and, soon after exposure, the radiolabel is no longer attached to active aldicarb and, thus, cannot be used as a reliable indicator of the amount of active aldicarb in the body. The second approach involves repeated dermal dosing with the technical material and determination of the NOAEL by this route of exposure. The risk assessment is then conducted by applying the safety factor directly to the dermal NOAEL. This method has the advantage of not relying on a dermal equivalent dose. However, it is also not appropriate for aldicarb, because the granular product is specifically designed to minimize expo-
1120
CHAPTER 53
Aldicarb Risk Assessment
sure by adhering the technical material to the granule with a binding agent and removing any excess dust after the product is manufactured. Thus a study with the technical material would be expected to greatly overestimate the amount of material absorbed in an agricultural application situation. Because of the problems with the traditional study designs used for the hazard assessment component of the occupational risk assessments, a 21-day dermal study with the granular product was conducted and used this study as the basis for the risk assessment. The effect of Temik grit (containing 14.75% aldicarb), administered by semioccluded topical application on peripheral (erythrocyte and plasma) and brain cholinesterase activities, was evaluated in CD (Sprague-Dawley) rats. The granular Temik grit was applied to the shaved dorsum, and the site moistened with saline to ensure good contact with the skin. Eight rats per sex per dose were exposed topically 6 hours/day, 5 days/week (Monday through Friday) for three consecutive weeks at 0, 100, 250, and 500 mg/kg/day. Body weights and clinical observations were recorded daily, and feed consumption was recorded twice weekly (Mondays and Fridays). Blood samples (0.25 ml) were taken from the lateral tail vein of each rat 1 hour postdosing on day 1 (Monday) and day 5 (Friday) on each of the three weeks of exposure. On the last day of exposure, after the blood sampling, the animals were sacrificed and the brains were weighed and analyzed for cholinesterase activity. For the males, there were no effects on daily body weights or daily weight changes. Absolute and relative brain weights were equivalent across all groups. Feed consumption was equivalent across groups for all intervals. There were no dose-related clinical signs of toxicity. Plasma cholinesterase levels were significantly reduced at 250 mg/kg/day on days 1,5,8,12, and 19 (but not on day 15 when the level was reduced, but not statistically significantly, to 81.4% of controls). Erythrocyte cholinesterase levels were not statistically different across all groups at all time points evaluated, but the mean activity at 500 mg/kg/day was clearly reduced relative to the control group values on all days evaluated. Brain cholinesterase activity on day 19 at termination of the study was not affected at any treatment level. For the females, there were no effects of treatment on body weights, weight changes, absolute or relative brain weights, or on feed consumption for any group at any time point during the study. There were also no dose-related clinical signs of toxicity. Plasma cholinesterase levels were reduced at 250 and 500 mg/kg/day on all days evaluated (days 1,5,8,12,15, and 19). Erythrocyte cholinesterase levels were reduced at 500 mg/kg/day. Brain cholinesterase activity on day 19 was not affected at any treatment level. In conclusion, the study evaluating Temik 15G grit, administered by occluded topical application, demonstrated that effects on peripheral and brain cholinesterase activity were rapidly reversible, and that there were no cumulative effects on these parameters over time. A NOAEL was established for erythrocyte cholinesterase activity at 250 mg/kg/day, and for brain cholinesterase activity of at least 500 mg/kg/day, for both male and female rats.
The NOAEL for RBC cholinesterase activity of 100 mg/kg/ day Temik grit corresponds to a NOAEL for aldicarb of 15 mg/kg/day because the granule is 15% aldicarb. The NOAEL for aldicarb when administered orally is 0.025 mg/kg/day based on RBC cholinesterase inhibition in the study with human volunteers. Thus the dermal study with the formulated product gives 15/0.025 = 600 extra margin of safety relative to oral dosing. In other words the risk assessment for occupational exposure for the granular product using an appropriate toxicity study provides a more realistic margin of safety. This is based on the fact that the release of aldicarb from the granule is very slow and thus the amount absorbed is extremely low. Therefore, any inhibition which occurs is rapidly reversed within seconds. One could argue that because the dermal study was conducted in the rat, an extra lO-fold margin of safety should be applied to the risk assessment to account for the interspecies extrapolation. This is clearly not necessary for aldicarb as the entire toxicology database demonstrates that animals and humans respond very similarly to the effects of exposure. In fact, in this case, the only expected interspecies difference is a lower dermal absorption through human skin compared to rat skin. However, even if an extra lO-fold interspecies safety factor is applied to the risk assessment, the granular product and dermal exposure provide at least 150 extra margins of safety relative to an assessment, which assumes oral dosing with the technical material. Furthermore, the modeling work supports and strengthens the results of this dermal toxicity study based on a hypothesis for total dermal absorption of the material and the hypothesis that this absorption would be constant over the maximum 8-hour workday.
53.7 CONCLUSION Aldicarb is an economically important and scientifically interesting pest control chemical. Were this material not available, in many cases a suitable alternative is not available. The alternative control strategies would require treatment with a number of different chemicals, resulting in a concomitant increase in potential exposure as well as cost. Aldicarb has high acute toxicity, so it must be carefully managed and used according to label instructions to ensure its safe use for crop protection. The following are the major conclusions of this chapter: 1. The biomarker effect observed in all animal and human testing demonstrated that ChEI is the only effect observed. 2. To assess the recovery time following aldicarb exposure, it is important to know the time to recovery of ChEI following exposure. A model was used to clearly answer the question: What is the time to recovery for an individual exposed to the D.S. EPA reference dose (0.001 mg/kg/day)? The model predicted that recovery time for someone exposed to the full reference dose is instantaneous. This is because no deviation from baseline ChE activity is predicted at the RID. This is expected given that the RID was set at a level10-fold lower than the human NOAEL for the effect. Any aldicarb exposure
References
encountered in a given meal will be removed from the body prior to any subsequent exposure. Thus acute risk assessments for aldicarb should be based on the amount received in any individual meal and the exposures from meal to meal should not be added together over the course of a 24-hour day. 3. Detailed analysis of the pharmacokinetics of reversal of cholinesterase inhibition has shown that the very rapid reversibility of effects following exposure allows safe exposure to low levels of aldicarb (i.e., RID). 4. In the dermal exposure risk assessment section, it was clearly demonstrated that release of aldicarb from the granule formulation is very slow and thus the amount absorbed is extremely low. Therefore, any inhibition which occurs is rapidly reversed within seconds. Furthermore, this is in line with the pharmacokinetic model for oral administration, as stated in the chapter. 5. For the overall risk assessment for aldicarb, we have clearly demonstrated that both the humans and animals are very similar (i.e., toxicity profile and rapid reversibility for RBC cholinesterase inhibition) in their response to this product; thus there should be no requirement for an interspecies 10 x -fold safety factor. In-depth risk assessment methodologies such as these, coupled with appropriate toxicological research, provide greater certainty in the risk assessment and, in the case of aldicarb, show that it can be used safely for crop protection, both for agricultural workers and for consumers.
REFERENCES Alvarez, A. P. (1992). "Pharmacology and Toxicology of Carbamates." In "Clinical and Experimental Toxicology of Organophosphates and Carbamates" (B. Ballantyne and T. C. Marrs, eds.), pp. 40-46. ButterworthHeinemann. Anderson, S., Tyl, R., Gilliam, A., Tobia, A., and Rieth, J. (2000). The toxicokinetics of peripheral cholinesterase inhibition from orally administered aldicarb in adult male CD® rats. Unpublished abstract (submitted to Society of Toxicology, 2001 Meeting). Baron, R. (1994). A carbamate insecticide: A case study of aldicarb. Environ. Health Perspect. 103(Suppl. 11), xxx-xxx. California Environmental Protection Agency (1998). "Summary of Toxicology Data for Aldicarb." Chemical Code # 000575, Tolerance # 00269, SB 950 # 130, File # T981120, Department of Pesticide Regulation, Medical Toxicology Branch, Sacramento. Carlock, L., Chen, W., Gordon, E., Killeen, J., Manley, A., Meyer, L., Mullin, L., Pendino, K., Percy, A., Sargent, D., and Seaman, L. (1999). Regulating and assessing risks of cholinesterase-inhibiting pesticides: Divergent approaches and interpretations. 1. Toxicol. Environ. Health 2, 105160. Cohrssen, J., and Covello, V. (1989). "Risk Analysis: A Guide to Principles and Methods for Analyzing Health and Environmental Risks." United States Council on Environmental Quality, Executive Office of the President (ISBN 0-934213-20-80).
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Dourson, M., Knauf, L., and Swartout, J. (1992). On reference dose and its underlying toxicity data base. Toxicol. Ind. Health 8, 171-189. Food Quality Protection Act of 1996, 104th Congress, 2nd session. Report 104669, Part 2, pp. 1-89. U.S. Gov. Printing Office, Washington, DC. International Agency for Research on Cancer (IARC) (1991). "Occupational Exposures in Insecticide Application, and Some Pesticides," Vol. 53. IARC, Lyon, France. International Programme on Chemical Safety (IPCS) (1991). "Aldicarb," Environmental Health Criteria 121. World Health Organization, Geneva. International Programme on Chemical Safety (IPCS) (1994). "Assessing Human Health Risks of Chemicals: Derivation of Guidance Values for HealthBased Exposure Limits," Environmental Health Criteria 170. World Health Organization, Geneva. Food and Agriculture Orgauization-World Health Organization (FAOIWHO) (1992). Aldicarb. In "Joint FAOIWHO Meeting on Pesticide Residues," Rome, 21-30 September 1992. O'Brien, R. D. (1967). "Insecticides: Action and Metabolism." Academic Press, New York. Phillips, J., Powell, G., Scarborough, A., Barraj, L., and Petersen, B. (2000). Acute dietary risk assessment of aldicarb, a reversible carbamate insecticide. Unpublished abstract (submitted to Society of Toxicology, 2001 Meeting). Renwick A., and Lazarus, N. (1998). Human variability and noncancer risk assessment-an analysis of the default uncertainty factor. Regul. Toxicol. Pharmacal. 27, 3-20. Rieth, l., and Starr, T. (1989). Chronic bioassays: Relevance to quantitative risk assessment of carcinogens. Regul. Toxicol. Pharmacol. 10, 160-173. Rotenberg, M., and Almog, S. (1995). Evaluation of the decarbamylation process of cholinesterase during assay of enzyme activity. Clin. Chim. Acta 240, 107-116. Swartout, l., Price, P., Dourson, M., Carlson-Lynch, H., and Keenan, R. (1998). A probabalistic framework for the reference dose. Risk Anal. 18,271-282. Tobia, A., McCahon, P., and Carmichael, N. (2000). A safety and tolerability study of aldicarb at various dose levels in healthy male and female human volunteers. Unpublished abstract (submitted to Society of Toxicology, 2001 Meeting). Tyl, R., Ross, w., Basham, K., Gilliam, A., Myers, C., Rieth, J., Lunchick, c., and Tobia, A. (2000). Cholinesterase activity in CD® rats following topical application of TEMIK® 15G for one week. Unpublished abstract (submitted to Society of Toxicology, 2001 Meeting). U.S. Environmental Protection Agency (U.S. EPA) (1984). "Pesticide Assessment Guidelines: Subdivision F, Hazard Evaluation, Human and Domestic Animals." U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA) (1987). "Pesticide Assessment Guidelines, Subdivision U, Applicator Exposure Monitoring." NTIS PB87-133286, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA) (1993). "Reference Dose (RID): Description and Use in Health Risk Assessments." Background Document lA, Integrated Risk Information, March 15, 1993, U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA) (1998). "The Pesticide Handlers Exposure Database (PHED), Version l.1-PHED Surrogate Exposure Guide, Estimates of Worker Exposure." U.S. Environmental Protection Agency, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA) (2000). WHO (1987). Williams, R., Rieth, l., and Tobia, A. (2000). Non-linear mixed effects models for cholinesterase activity in humans exposed to aldicarb. Unpublished abstract (submitted to Society of Toxicology, 2001 Meeting).
CHAPTER
54 Imidacloprid: A Neonicotinoid Insecticide Larry P. Sheets Bayer Corporation
54.1 INTRODUCTION Imidacloprid is the principal representative of a new pesticide class, the neonicotinoid insecticides. These insecticides are designed to act on nicotinic receptors to control insect pests and, at the same time, to be relatively nontoxic to vertebrate species. This is accomplished by selecting compounds for commercial development that are highly specific for subtypes of nicotinic receptors that occur in insect tissues. The effect of neonicotinoid insecticides on the central nervous system of vertebrates is further reduced by poor penetration of the blood-brain barrier. The toxicology database supports the success of this strategy for imidacloprid, with signs of nicotinic stimulation (e.g., tremor) evident only at relatively high levels of exposure. By oral administration, imidacloprid is rapidly absorbed, metabolized in the liver, and excreted, primarily via the urine. Results from long-term dietary-exposure studies support rapid metabolism, with little evidence of cumulative toxicity and minimal effects, even at maximum-tolerated doses. Imidacloprid is not mutagenic or carcinogenic. Furthermore, it is not a primary embryotoxicant or a reproductive toxic ant, nor is it teratogenic. Due to its high insecticidal potency and relatively low mammalian toxicity, imidacloprid has a very high margin of safety.
54.2 HISTORICAL OVERVIEW 54.2.1 CHEMISTRY Imidacloprid [1-[(6-chloro-3-pyridinyl)methyl]-N-nitro-2-irnidazolidinimine] is the first representative of the neonicotinoid insecticides that was registered for use and is presently the most important commercial product. The history of the neonicotinoids can be traced to the late 1970s, when chemists at Shell Chemical Company investigated the heterocyclic nitromethylenes as potential insecticides (Schroeder and Flat-
Handbook of Pesticide Toxicology Volume 2. Agents
turn, 1984; Soloway et al., 1978). An excellent review of the discovery and early development of these insecticides has been compiled by Yamamoto and Casida (1999). The term "neonicotinoid" is used to distinguish these chemicals from the nicotinoids (Tomizawa and Yamamoto, 1993), with the neonicotinoids being more highly effective as insecticides and less toxic to vertebrate species. Representatives from this group are also referred to as "chloronicotinyls" to emphasize the importance of the chlorine atom for insecticidal potency. Imidacloprid was discovered in 1984 by chemists at Nihon Bayer Agrochem who were exploring the introduction of a 3-pyridylmethyl group on the nitromethylene heterocycle structure (Shiokawa et aI., 1986). The introduction of this moiety has been shown to greatly increase insecticidal activity and reduce mammalian toxicity (Kagabu et aI., 1992; Zwart et aI., 1992 and 1994), while retaining the many other properties that are important for commercial applications. Since the discovery of imidacloprid, several other chemical analogs with the 6-chloro-3-pyridylmethyl moiety have been developed for commercial use (Fig. 54.1). Included in this group are acetamiprid (Takahashi et aI., 1992; Yamada et aI., 1999), nitenpyram (Minamida et aI., 1993) and, more recently, thiacloprid. The replacement of the chloropyridinyl moiety with a chlorothiazolyl group led to the development of a "second generation" of neonicotinoid insecticides (Maienfisch et aI., 1999). This substitution has been shown to further reduce potency in assays with mammalian receptors but does not appear to reduce toxicity to mammals or to reduce activity at the insect nicotinic receptor (Chao and Casida, 1997; Liu et aI., 1993; Zhang et aI., 2000). Compounds in this group that have been developed for commercial use include clothianidin (TI435) and thiamethoxam (CGA 293'343; Maienfisch et aI., 1999) (Fig. 54.1). Thiamethoxam is the first representative from this group that was registered for use (Wiesner and Kayser, 2000).
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Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
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CHAPTER 54
Imidacloprid NEONICOTINOID INSECTICIDES
CHLORPYRIDINES
Im1<:laclopnd
ACelar01pnd
Nitenpyram
Tluaciop"d
Figure 54.1
Neonicotinoid insecticides.
54.2.2 NICOTINIC ACTIVITY 54.2.2.1 Insects The insecticidal activity of the neonicotinoids is attributed to actions on post-synaptic nicotinic receptors (Buckingham et aI., 1997; Nagata et aI., 1998) which, in insects, are located exclusively in the central nervous system. In insects, multiple sUbtypes of nicotinic receptor have been identified which express different physiological and pharmacologic properties (Gundelfinger and Schulz, 2000; Wiesner and Kayser, 2000). With respect to the neonicotinoids, it has been determined that imidacloprid acts on at least three pharmacologically distinct subtypes of nicotinic receptor in the cockroach (Buckingham et aI., 1997). Further characterization of the nicotinic receptors that exist in insect tissues and the relative activity of neonicotinoids on the various sUbtypes are very active areas of research. The treatment of insect neuronal preparations with a neonicotinoid produces a bi-phasic response, consisting of an initial increase in the frequency of spontaneous discharge that is followed by a complete block to nerve impulse propagation (Schroeder and Flattum, 1984). Signs of intoxication in the American cockroach (Periplaneta americana) following exposure to imidacloprid consist of uncoordinated abdominal quivering, wing flexing, tremor, and violent whole-body shaking, followed by prostration and death (Schroeder and Flattum, 1984). Insecticidal activity is greatly enhanced by synergists that inhibit oxidative degradation (Liu and Casida, 1993), which would appear to support including a synergist in commercial formulations. 54.2.2.2 Mammals Mammalian tissues also contain many subtypes of nicotinic receptor. The various sUbtypes are derived from five homologous subunits, in combinations that are formed from nine a, four {3,
y, 8 and E subunits (Tomizawa et aI., 1999). In mammals, nicotinic receptors are located in many tissues, including autonomic ganglia, skeletal muscle (neuromuscular junction), spinal cord, and a number of brain regions. Differences in binding properties to the various receptor subtypes contribute greatly to the much lower activity of neonicotinoids in vertebrate tissues, as compared to tissues from insects (Yamamoto et aI., 1998). There is an extensive database for differential sensitivity with imidacloprid (Chao and Casida, 1997; Liu and Casida, 1993; Matsuda et aI., 1998; Methfessel, 1992; Nagata et aI., 1999; Tomizawa et aI., 1999) that has been summarized by Tomizawa and Casida (1999). The relative specificity for the nicotinic receptor in insects is used to select compounds for commercial development. The success of this strategy is reflected by very high margins of safety for these insecticides (Leicht, 1993). The acute toxicity (defined by lethal potency) of various neonicotinoid insecticides and related analogs in mammals is most closely related to potency at the a7 nicotinic receptor subtype, with a decreasing relationship reported sequentially at a4, {32, a3, and al nicotinic receptors, respectively (Tomizawa and Casida, 1999). However, acute toxicity in mammals involves complex actions at multiple receptor subtypes, with relative subtype specificity conferred by even minimal structural changes. Furthermore, the actions of neonicotinoids at these receptor subtypes involve a combination of agonist and antagonist effects (Nagata et aI., 1998). Given this combination of actions, known or expected differences in relative specific activity at each nicotinic receptor subtype, and expected differences in distribution to target tissues, the toxic effects in vivo would likely vary among representative compounds. However, there has been no systematic assessment oftoxicity in vivo, with tests conducted under appropriately standardized conditions.
54.3 METABOLISM AND TOXICOKINETICS The information available on the metabolism and toxicokinetics of imidacloprid in the rat is described in additional detail elsewhere (Thyssen and Machemer, 1999). Briefly, there are two major routes of metabolism in mammals. The first involves oxidative cleavage to imidazolidine and 6-chloronicotinic acid. The imidazolidine moiety is then excreted via the urine. The nicotinic moiety is further degraded via glutathioneconjugation to a derivative of mercapturic acid and then to methyl mercaptonicotinic acid. This moiety is also conjugated with glycine to form a hippuric acid conjugate for excretion. The second substantive route in the biotransformation of imidacloprid involves the hydroxylation of the intact molecule in the imidazolidine ring, followed by the elimination of water and the formation of an unsaturated metabolite. In the rat, there are no qualitative differences between males and females after the oral administration of a low dose of 1 mg/kg body weight or a dose of 20 mg/kg body weight. At both dose levels, the same complement of metabolites is present in both sexes, although at the higher dose of 20 mg/kg body weight, orally treated females exhibit a slightly higher renal
54.5 Acute Toxicity
elimination than males. More than 90% of a given dose is eliminated within 24 hours, with total excretion by 48 hours. Eighty percent of the dose is excreted via the urine, with the rest eliminated via the feces. Imidacloprid is absorbed and widely distributed to organs within one hour following oral administration to rats. Wholebody autoradiography indicates that imidacloprid is not distributed to fatty tissues, to tissues in the central nervous system (CNS), or to the mineral components of bone. These results indicate that there is low potential for accumulation and poor penetration of the blood-brain barrier, at least to dose levels of up to 20 mg/kg body weight. Poor penetration of the bloodbrain barrier has also been reported with other neonicotinoids (Yamamoto et aI., 1995). This property reduces their access to receptors in the CNS, such that centrally mediated effects would not be expected at low levels of exposure.
54.4 MAMMALIAN TOXICOLOGY The peer-reviewed literature includes very little information on the toxicity of imidacloprid or other neonicotinoid insecticides in mammals. Work that has been published has generally dealt with a determination of acute lethal potency (e.g., LD50) for a series of structural analogs, without further assessment. One such study reported the presence of tremor in mice that had been treated with an acute oral dose of imidacloprid or one of several other neonicotinoids (Chao and Casida, 1997). This finding provides evidence of nicotinic stimulation at near-lethal or lethal dose levels. A second source is a book edited by Yamamoto and Casida (1999), with chapters that discuss mammalian toxicology data for imidacloprid (Thyssen and Machemer, 1999), nitenpyram (Akayama and Minamida, 1999), and thiamethoxam (Maienfisch et aI., 1999). Finally, there is a published comparison of the findings of neurotoxicity studies that were conducted in industry laboratories with commercial products (Sheets, 2001). This work is summarized in Section 54.11.4, following a review of the findings with imidacloprid. The general absence of published information on the toxicology of imidacloprid and other neonicotinoids in mammals contrasts with the extensive database that has been generated by industry laboratories to support the registration of commercial products. The remainder of this chapter is largely devoted to a review of the toxicology studies that constitute the database that Bayer has generated for imidacloprid. These studies were performed in accordance with regulatory guidelines, including those of the U.S. EPA (FIFRA), the OECD, and the Japanese MAFF, and in compliance with the associated Good Laboratory Practice (GLP) requirements. The compound tested in these studies was technical-grade imidacloprid, with a purity of 94-98% active ingredient. For extended periods of exposure, imidacloprid was generally mixed in the diet and provided for ad libitum consumption. Animals were acquired from commercial vendors as
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purpose-bred animals and were housed under standardized conditions that meet or exceed accepted standards for animal care.
54.5 ACUTE TOXICITY An overview of the results for acute toxicity studies that have been conducted with imidacloprid is provided in Table 54.1. Acute exposure to imidacloprid was determined to produce minimal evidence of toxicity by dermal and inhalation routes of exposure and moderate acute toxicity by oral administration. Imidacloprid is not an irritant and does not produce evidence of dermal sensitization. To assess acute oral toxicity, technical-grade imidacloprid was administered as an aqueous suspension to fasted, young-adult Wistar rats (5/sex/dose). Doses of 50 mg/kg in males and 100 mg/kg in females produced no evidence of exposure. By comparison, higher doses of up to 315 mg/kg in males or females produced clinical signs, without causing mortality. At dose levels greater than 315 mg/kg, the incidence of mortality increased rather abruptly, with 20% mortality in both sexes at a dose of 400 mg/kg and 100% mortality at 500 mg/kg body weight. Clinical signs that were evident following treatment included tremor, gait incoordination, and evidence of decreased motility and activity, as well as nasal and urine staining. Signs of intoxication were evident within 15--40 minutes following oral administration and, with few exceptions (e.g., stains), were reversible within eight to 24 hours following treatment. This outcome is consistent with the rapid distribution and metabolism profile that was summarized in Section 54.3. Treatment-related deaths generally occurred within three to seven hours following treatment.
Table 54.1 Acute Toxicity Studies with Imidacloprid
Q
Animal
Route of
LDsolLCso
species
exposure
(mg/kg BW/mg/m 3 air)
Mouse
Oral
131-168
Rat
Oral
424--475
Rat
Dermal
>5000
Rat
Inhalation AE 4h
>69 b
Rat
Inhalation dust 4h
>5323 c
Rabbit
Dermal
Not an irritant
Rabbit
Eye
Not an irritant
Guinea pig
Dermal
Negative for sensitizationd
a LDso and LCso values represent the results for both sexes. b Aerodynamic droplet size <5 J.lM; 100%; max conc. C
Aerodynamic particle size <5 J.lM; 4--11 %.
d Magnusson and Kligman Test.
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CHAPTER 54 Imidac10prid
54.6 SUB CHRONIC TOXICITY
54.7 CHRONIC TOXICITY AND CARCINOGENICITY
54.6.1 RAT 54.7.1 RAT Imidacloprid was administered through the diet for a period of 13 weeks to young-adult Wistar rats (lO/sex/dietary level) to examine cumulative toxicity, with sustained exposure, and to establish dietary levels for the chronic toxicity/carcinogenicity study. In this study, the test substance was provided for ad libitum consumption at concentrations of 150, 600, or 2400 ppm, which corresponded to average daily doses of 14, 61, or 300 mg/kg body weight for males and 20, 83, or 422 mg/kg body weight for females. Satellite groups of control and highdose animals (lO/sex/level) were retained for four weeks after the 13-week period of exposure to assess reversibility. Measures of cholinesterase activity (brain, plasma, and erythrocyte) were included to verify the expected absence of inhibition. Clinical signs associated with treatment were not evident in males or females at any dietary level. Body weight and food consumption were reduced at the 600 ppm (males only) and 2400 ppm (both sexes) dietary levels. The average body weight for high-dose males and females was approximately 15% less than control. The liver was the principal target organ, with hypertrophy of hepatocytes and sporadic cell necrosis in highdose males only. Liver pathology was mild at termination of the study and was fully reversible within the recovery period. Other effects in high-dose males and females included elevated serum alkaline phosphatase and alanine aminotransferase (ALAT) activities and a slight increase in blood clotting time. There was no inhibition of cholinesterase activity at any dietary level. The NOEL (no-ob served-effect level) for this study was 14 mg/kg/day in males and 83 mg/kg/day in females.
A combined chronic toxicity/carcinogenicity study was performed, with imidacloprid administered through the diet for a period of two years to Wistar rats (SO/sex/dietary level). An additional set of animals (lO/sex/level) was reserved for interim examination after twelve months. The test substance was provided for ad libitum consumption at concentrations of 100, 300, 900, or 1800 ppm, which corresponded to average daily doses of 5.7,17,51, or 103 mg/kg body weight for males and 7.6, 25, 73, or 144 mg/kg body weight for females. Treatment-related clinical signs were not evident in either sex, and there was no effect on survival at any dietary level. Body weight was reduced by 12% in both sexes at the 1800 ppm dietary level and by 5-8% in males and females at the 900 ppm dietary level, and was not affected at lower levels of exposure. At 1800 ppm, serum alkaline phosphatase, creatine kinase, and aspartate aminotransferase (AS AT) activities were elevated and cholesterol was reduced. Microscopic lesions were also apparent in the thyroid at this dietary level, with mineralization of the colloid, fewer colloid aggregation cites, and parafollicular hyperplasia sites. These lesions were ascribed to an enhancement of biological aging processes and were not accompanied by a change in thyroid function (e.g., plasma T3, T4, and TSH levels were normal). Mineralization of the colloid in the thyroid follicles was also evident in males at 300 ppm and in both sexes at 900 ppm. There was no change in liver morphology and no inhibition of cholinesterase activity (brain, plasma, or erythrocyte) at any level. The NOEL in this study was 5.7 mg/kg/day. There was also no evidence of carcinogenicity. 54.7.2 MOUSE
54.6.2 DOG Toxicity was examined in a nonrodent species by administering imidacloprid through the diet for a period of 13 weeks to young-adult, pure-bred beagle dogs (4/sex/dietary level) at dietary concentrations of 200, 600, or 180011200 ppm. The 1800 ppm dietary level produced a sharp reduction in weight gain, relative to controls. Body weight was regained after week 4, upon reducing the high dose to 1200 ppm. Tremor was evident in males and females in the 600 ppm and 180011200 ppm dietary groups. The tremor was more severe and occurred at a higher incidence in high-dose animals, relative to the next lower dose. However, it was noted that tremor was not observed at comparable dietary levels in other dog studies, including the one-year dietary study (see Section 54.7.4). There was no evidence of tissue damage by clinical chemistry, gross necropsy examination, tissue weight, or microscopic examination at any dietary level. The NOEL for this study was 200 ppm in both sexes.
To further assess oncogenic potential, imidacloprid was administered through the diet for a period of 24 months to B6C3F1 mice (SO/sex/dietary level), at concentrations of 100, 330, 1000, or 2000 ppm. An additional set of animals (lO/seX/level) was reserved for interim examination after twelve months. These dietary concentrations resulted in average daily doses of 20,66, 208, or 414 mg/kg body weight for males and 30, 104,274, or 424 mg/kg body weight for females. There were no clinical signs associated with treatment and no effect on survival at any dietary level. Males and females that received the 2000 ppm dietary level had a marked decrease in body weight gain, relative to controls, with correspondingly lower food and water consumption. The difference in body weight reached 29% less than controls, indicating that this level exceeded an MTD. Liver changes were also evident at 2000 ppm but not at lower dietary levels. These consisted of low-grade periacinary hepatocyte hypertrophy, which was considered to represent metabolic adaptation to this xenobiotic. Effects that were evident at 1000 ppm consisted of reduced food
54.9 Developmental Toxicity consumption (females only) and reduced body weight, relative to controls, for males and females (up to 10% and 5%, respectively). There were no changes in serum chemistry, tissue weight, or tissue morphology (by gross and microscopic examination) associated with treatment at any dietary level. The number, type, distribution, and time of occurrence of tumors provided no evidence that imidacloprid has carcinogenic potential. The overall NOEL in this study was 330 ppm in males and females. 54.7.3 CLASSIFICATION FOR CARCINOGENICITY
Based on the collective results of the chronic toxicity and carcinogenicity studies in the rat and mouse, the U.S. EPA has classified imidacloprid in category "E." This classification indicates that the database for imidacloprid supports evidence of noncarcinogenicity for humans. 54.7.4 DOG
To evaluate chronic toxicity in a nonrodent species, imidacloprid was administered through the diet for a period of 52 weeks to young-adult, pure-bred beagle dogs (4/sex/dietary level). The test substance was provided for ad libitum consumption at concentrations of 200, 500, or 1250 ppm. The 1250 ppm dietary concentration was increased to 2500 ppm from week 17 onwards. These levels corresponded to daily doses of 6.1,15, and 41172 mg/kg/day. The 1250 ppm dietary level was associated with a slight, but transient, fall in food consumption in both sexes. A similarly transient effect was evident when the dietary concentration was increased to 2500 ppm during week 17. The tremor that was evident in the subchronic dog study, at dietary concentrations of 600 ppm or greater (see Section 54.7.2), was not evident here at any dietary concentration. Effects at the highest dietary level included a slight increase in plasma cholesterol (females only) and a slight increase in hepatic cytochrome P-450 activity (both sexes). The induction of cytochrome P-450 enzymes was associated with a slight increase in liver weight. Thus, the liver was the principal target organ. The chronic NOEL in the dog was 15 mg/kg/day.
54.8 MUTAGENICITY Imidacloprid has been evaluated for mutagenicity using a full complement of in vitro and in vivo tests that is required for registration. The results from this database indicate that imidacloprid is not mutagenic (Table 54.2). Briefly, the in vitro point mutation tests were negative. This includes the results of chromosomal aberration tests conducted in vitro, which were negative at non-cytotoxic concentrations and showed only slightly positive effects at cytotoxic concentrations. In vivo chromosomal aberration tests were also all negative. Finally, the mitotic
1127
recombination test that is conducted in yeast, the rec assay with Bacillus subtilis, and the unscheduled DNA synthesis (UDS) test were also all negative.
54.9 DEVELOPMENTAL TOXICITY 54.9.1 RAT
The potential for imidacloprid to produce developmental toxicity, induding teratogenicity, was examined in the rat. In this study, mated female Wistar rats (25/dose level) were treated daily, by gavage, on gestation days 6 through 15, with doses of 10, 30, or 100 mg/kg body weight per day. On day 21 postcoitum, the fetuses were delivered by cesarean section and examined for development, including skeletal alterations. The highest dosage of 100 mg/kg/day produced signs of maternal toxicity and a delay in embryo development. The offspring of the high-dose dams had wavy ribs as a reversible finding. While an increased incidence of wavy ribs, relative to controls, was ascribed to treatment, it was noted that the incidence was within the range of historical controls. No fetal malformations were evident at any dose level. The maternal NOEL was 10 mg/kg body weight per day and the fetal NOEL was 30 mg/kg body weight per day. These results indicate that imidacloprid is not a primary embryotoxicant and is not teratogenic. 54.9.2 RABBIT
The potential for imidacloprid to produce developmental toxicity was also examined in the rabbit. In this study, mated Chinchilla rabbits (16/dose level) were treated by gavage on gestation days 6 through 18, with daily doses of 8, 24, or 72 mg/kg body weight. Cesarean section and examination of embryo and fetal development, including fetal skeletal alterations, were conducted on day 28 postcoitum. The highest dosage of 72 mg/kg/day produced severe maternal toxicity, including some deaths. Abortions and complete resorptions, delayed ossification, and reduced fetal weights were also evident at this dose level. The next lower dose produced decreased food consumption and reduced body weight gain, relative to controls, but no effects on the fetus. Thus, embryotoxicity was only evident at a maternally-toxic dose. As with the rat, no fetal malformations were evident at any dose level. The maternal NOEL was 8 mg/kg body weight per day and the fetal NOEL was 24 mg/kg body weight per day. These results indicate that imidacloprid is not a primary embryotoxicant and is not teratogenic.
54.10 REPRODUCTIVE TOXICITY The potential effects of imidacloprid on reproduction and development were examined in a two-generation, two-litter study
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CHAPTER 54
Imidacloprid Table 54.2 Mutagenicity Studies with Imidacloprid Point mutation Salmonella microsome (AMES) test
Negative
Reverse mutation test E. Coli
Negative
HPRT Chinese hamster ovary (CHO)
Negative
Chromosomal aberration in vitro Cytogenetics human Iymphocytes
Slightly positive (At cytotoxic concentrations only)
Sister chromatid exchange (SCE) Chinese hamster ovary (CHO)
Slightly positive (At cytotoxic concentrations only)
Chromosomal aberration in vivo Micronucleus mouse bone marrow
Negative
Sister chromatid exchange Chinese hamster bone marrow
Negative
Cytogenetics Chinese hamster bone marrow
Negative
Cytogenetics Negative
Mouse spermatogonia
Other genotoxicity tests Mitotic recombination yeast
Negative
Rec assay (B. subtilis)
Negative
Unscheduled DNA synthesis Rat hepatocytes
in Wistar rats (30/sexJdietary level in the parental generation). In this study, technical-grade imidacloprid was mixed in the diet for ad libitum consumption, at dietary concentrations of 100, 250, and 700 ppm. The treated feed was provided during a prepairing period of 84 days and throughout pairing, gestation, and lactation for breeding of the FlA and FIB litters. Following weaning of the FIB litters on day 21 postpartum, the Fl-generation parental animals were selected. The diets were fed for 105 days prior to pairing and throughout pairing, gestation, and lactation periods for breeding of the F2A and F2B litters. This study included assessments of gonadal function, estrus cycle, mating behavior, conception, parturition, lactation, weaning, and the growth and development of the offspring of multiple generations, as well as neonatal morbidity, mortality, and behavior. Maternal toxicity was evident at the high dose as a decrease in body weight gain and food consumption, relative to controls. A pronounced reduction in body weight gain and food consumption, relative to controls, occurred during lactation. These effects coincided with the large increase in dietary exposure that occurs during lactation, when food consumption increases rather dramatically to support the offspring. Liver enzymes (cytochrome P-450, O-demethylase, and N -demethylase) were also induced in high-dose maternal animals. In the offspring, toxicity was evident at the high dose as a marked decrease in body weight gain, relative to controls, before weaning on postnatal day 21. There were no effects on reproduction or development at any dietary level. More specifically, there was no
Negative
effect of treatment on mating indices, fertility, gestation, conception, litter size, or mortality, at any dietary level. There was also no evidence of pathology, in the form of malformations, gross lesions, a change in tissue weight, or histopathology, at any exposure level. The NOEL in this study was 6.7 mg/kg body weight per day for the adult and 12.5 mg/kg body weight per day for the offspring.
54.11 NEUROTOXICITY 54.11.1 GENERAL The toxicology database for imidacloprid includes acute and subchronic neurotoxicity screening studies that were performed in accordance with the U.S. EPA (FIFRA) guidelines. These studies were conducted using young-adult male and female rats, following an acute oral dose, administered by gavage, or with 13 weeks of dietary exposure. Both studies included a functional observational battery (FOB) and a computer-automated test (figure-eight maze) to measure spontaneous activity, including habituation. At term (on day 14 after the acute dose or during week 14 of dietary exposure), a subset of the animals (6/sexJdose level) was anesthetized and perfused using an aldehyde fixative, and representative skeletal muscle and neural tissues were collected for microscopic examination.
54.11 Neurotoxicity
54.11.2 ACUTE NEUROTOXICITY
To evaluate acute neurotoxicity, Sprague-Dawley rats (12/sex/ dose level) received a single oral dose of imidacloprid, administered by gavage as an aqueous suspension at doses of 42,151, or 307 mg/kg body weight in males and 20, 42, 151, or 307 mg/kg body weight in females. Animals were evaluated using the FOB and the figure-eight maze one week prior to treatment, again at the time of peak neurobehavioral signs, which was approximately four hours following treatment, and on days 7 and 14 following treatment. The only evidence of systemic toxicity at 42 mg/kg was a slight decrease in the activity of females in the figure-eight maze. Additional effects that were evident at 150 mg/kg included tremor (one female), a slight decrease in body temperature, and red nasal stain. The highest dose produced severe acute toxicity, including lethality (two males and eight females). These deaths occurred within four hours to 24 hours following treatment. At the 4-hour observation period, tremor was apparent in all animals that were still alive and was more severe, relative to the next lower dose. Body temperature was also reduced an average 2.0°C and 5.5°C in males and females, respectively. Additional effects at this lethal dose included evidence of motor incoordination (e.g., incoordinated gait and impaired aerial righting), autonomic signs (e.g., perianal and urine stains), and CNS depression (e.g., minimal activity and a diminished response to stimuli). Clinical signs following acute exposure generally resolved in surviving animals within eight hours to 24 hours following treatment. Urine stain was the only effect that persisted for up to four days after treatment. All findings in the FOB and the figure-eight maze had resolved by day 7, which was the first test occasion following the day 0 time-point. Neuropathology was not evident at the highest dose level. The NOEL was 42 mg/kg for males and 20 mg/kg for females. 54.11.3 SUB CHRONIC NEUROTOXICITY
To assess neurotoxicity with a sustained exposure, imidacloprid was administered via the diet for 13 weeks to young-adult Fischer-344 rats (12/sex/dietary level), at dietary concentrations of 150,1000, and 3000 ppm. These dietary levels resulted in average daily exposures of 9.3,63, and 196 mg/kg for males and 10.5,69, and 213 mg/kg for females. The FOB and automated test of activity were performed one week prior to the initiation of treatment and during weeks 4, 8, and 13 of exposure. There was little evidence of toxicity in this study at any dietary concentration. Effects at dietary levels of 1000 ppm and 3000 ppm were generally limited to decreased food consumption and an associated decrease in body weight gain, relative to controls. The difference in body weight for high-dose males and females averaged 15% and 8% less than control, respectively. Toxicity was not evident by cage-side observation or the automated test of motor activity, at any dietary level. On the last test occasion (week 13), there was a modest increase in the incidence of high-dose males, relative to controls, with a slightly
1129
uncoordinated righting response that was ascribed to treatment. There was no evidence of neuropathology at the high dose. The NOEL for this study was 9.3 mg/kg body weight per day for males and 10.5 mg/kg body weight per day for females. 54.11.4 COMPARISON WITH OTHER NEONICOTINOIDS
The results from the neurotoxicity studies with imidacloprid compare closely with the findings of the acute and subchronic neurotoxicity studies that were conducted in industry laboratories with acetamiprid, clothianidin, thiacloprid, and thiamethoxam (Sheets, 2001). Comparisons involving clothianidin and thiacloprid are facilitated by the fact that those studies were conducted under comparable conditions and in the same laboratory as the studies with imidacloprid. For each of these compounds, the time of peak effects ranged from two hours to six hours following administration by gavage. The most consistent finding at a low dose was decreased activity, which was evident by observation and in the automated test devices. By comparison, the most common effects at higher dose levels were tremor, impaired pupillary function (either dilated or pin-point pupils), incoordinated gait, and hypothermia. In studies that included a lethal dose, deaths occurred within four hours to 24 hours following treatment. Except for some residual staining, recovery generally occurred within eight hours to 24 hours following treatment. Neuropathology was not evident with any of these compounds. The results from subchronic neurotoxicity studies with acetamiprid, clothianidin, thiacloprid, and thiamethoxam are also comparable with the findings with imidacloprid. Each compound produced minimal effects, other than decreased body weight and food consumption, at higher dose levels, and little or no overt evidence of an effect on the nervous system. Typically, there were no clinical signs, FOB findings, or effects on spontaneous activity in the automated devices, and little evidence of cumulative toxicity at any dietary level. Finally, none of these compounds produced neuropathology at the highest dietary level. 54.11.5 DEVELOPMENTAL NEUROTOXICITY
There is little information in the published literature to assist in determining the potential for imidacloprid, or any other neonicotinoid insecticide, to affect the developing nervous system. While the results from the developmental toxicity and multigeneration reproduction toxicology studies with imidacloprid provide no indication of developmental neurotoxicity, these studies are relatively limited in such assessment. A much more rigorous assessment of effects on the developing nervous system is provided by studies that are conducted according to the V.S. EPA guideline for a developmental neurotoxicity study (U.S. EPA, OPPTS 870.6300). This study design includes a complement of automated tests of cognition, auditory startle habituation, motor activity ontogeny, and an
1130
CHAPTER 54
Imidacloprid
extensive neuropathology assessment. A developmental neurotoxicity study has recently been completed with imidacloprid. Publication of this work is planned.
REFERENCES Akayama, A., and Minamida, 1. (1999). Discovery of a new systemic insecticide, nitenpyram and its insecticidal properties. In "Nicotinoid Insecticides and the Nicotinic Acetylcholine Receptor" (1. Yamamoto and J. E. Casida, eds.), pp. 127-148. Springer-Verlag, Tokyo. Buckingham, S. D., Lapied, B., Le Corronc, H., Grolleau, E, and Sattelle, D. B. (1997). Imidac10prid actions on insect neuronal acetylcholine receptors. J. Exp. Bioi. 200, 2685-2692. Chao, S. L., and Casida, J. E. (1997). Interaction of imidacloprid metabolites and analogs with the nicotinic acetylcholine receptor of mouse brain in relation to toxicity. Pest. Biochem. Physiol. 58, 77-88. Gundelfinger, E. D., and Schulz, R. (2000). Insect nicotinic acetylcholine receptors: Genes, structure, physiological and phannacological properties. In "Handbook of Experimental Phannacology, Vo!. 144, Neuronal Nicotine Receptors" (E Clementi, D. Fomasari, and C. Gotti, eds.), pp. 497-521. Springer-Verlag, Tokyo. Kagabu, S., Moriya, K., Shibuya, K., Hattori, Y., Tsuboi, S., and Shiokawa, K. (1992). 1-(6-Halonicotinyl)-2-nitromethylene-imidazolidines as potential new insecticides. Biosci. Biotech. Biochem. 56(2), 362-363. Leicht, W (1993). Imidacloprid: A chloronicotinyl insecticide. Pestic. Outlook 4(3), 17-21. Liu, M.- Y., and Casida, J. E. (1993). High affinity binding of [3HJImidacloprid in the insect acetylcholine receptor. Pestic. Biochem. Physiol. 46, 40-46. Liu, M. - Y., Lanford, J., and Casida, J. E. (1993). Relevance of HJImidacloprid binding site in house fly head acetylcholine receptor to insecticidal activity of 2-nitromethylene- and 2-nitroimino-imidazolidines. Pestic. Biochem. Physiol. 46, 200-206. Maienfisch, P., Brandl, E, Kobel, W, Rindlisbacher, A., and Senn, R. (1999). CGA 293'343: A novel, broad-spectrum neonicotinoid insecticide. In "Nicotinoid Insecticides and the Nicotinic Acetylcholine Receptor" (1. Yamamoto and J. E. Casida, eds.), pp. 177-209. Springer-Verlag, Tokyo. Matsuda, K., Buckingham, S. D., Freeman, J. c., Squire, M. D., Baylis, H. A., and Sattelle, D. B. (1998). Effects of the alpha subunit on imidacloprid sensitivity of recombinant nicotinic acetylcholine receptors. Brit. J. Pharmacol. 123,518-524. Methfessel, C. (1992). Effect of imidacloprid on the nicotinergic acetylcholine receptors of rat muscle. Pjlanzenschutz Nachricten Bayer 45,369-380. Minamida, I., Iwanaga, K., Tabuchi, T., Aoki, I., Fusaka, T., Ishizuka, H., and Okauchi, T. (1993). Synthesis and insecticidal activity of acyclic nitroethene compounds containing a heteroarylmethylamino group. J. Pesticide Sci. 18, 41. Nagata, K., Aoyama, E., Ikeda, T., and Shono, T. (1999). Effects of nitenpyram on the neuronal nicotinic acetylcholine receptor-channel in rat phaeochromocytoma PCI2 cells. J. Pesticide Sci. 24, 143-148. Nagata, K., Song, J. H., Shono, T., and Narahashi, T. (1998). Modulation of the neuronal nicotinic acetylcholine receptor-channel by the nitromethylene heterocycle Imidacloprid. J. Pharmacol. Exper. Ther. 285,731-738. Schroeder, M. E, and Flattum R. E (1984). The mode of action and neurotoxic properties of the nitromethylene heterocycle insecticides. Pest. Biochem. Physiol. 22, 148-160.
e
Sheets, L. P. (2001). Neonicotinoid Insecticides. In "Neurotoxicology Handbook" (E. J. Massaro, ed.), Vo!. 1. Humana Press, Totowa, NJ. In press. Shiokawa, K., Tsuboi, S., Kagabu, S., and Moriya, K. (1986). Jpn. Kokai Tokkyo Koho JP 61-267575. Soloway, S. B., Henry, A. C., Kollmeyer, W D., Padgett, WM., Powell, J. E., Roman, S. A., Tieman, C. H., Corey, R. A., and Home, C. A. (1978). Nitromethylene insecticides. Adv. Pestic. Sci. 4,206-217. Takahashi, H., Mitsui, J., Takakusa, N., Matsuda, M., Yoneda, H., Suzuki, J., Ishimitsu, K., and Kishmoto, T. (1992). NI-25, a new type of systemic and broad spectrum insecticide. In "Brighton Crop Protection Conferences B Pest and Diseases," Vo!. I, pp. 89-96. Thyssen, J., and Machemer, L. (1999). Imidacloprid: Toxicology and metabolism. In "Nicotinoid Insecticides and the Nicotinic Acetylcholine Receptor" (I. Yarnamoto and J. E. Casida, eds.), pp. 213-222. SpringerVerlag, Tokyo. Tomizawa, M., and Casida, J. E (1999). Minor structural changes in nicotinoid insecticides confer differential subtype selectivity for mammalian nicotinic acetylcholine receptors. Br. J. Pharmacol. 127, 115-l22. Tomizawa, M., LatIi, B., and Casida, J. E (1999). Structure and function of insect nicotinic acetylcholine receptors studied with nicotinoid insecticide affinity probes. In "Nicotinoid Insecticides and the Nicotinic Acetylcholine Receptor" (1. Yamamoto and J. E. Casida, eds.), pp. 271-292. SpringerVerlag, Tokyo. Tomizawa, M., Otsuka, H., Miyamoto, T., and Yamamoto, 1. (1995). Phannacological effects of Imidacloprid and its related compounds on the nicotinic acetylcholine receptor with its ion channel from the Torpedo electric organ. J. Pesticide Sci. 20,49-56. Tomizawa, M., and Yamamoto, I. (1993). Structure-activity relationships of nicotinoids and Imidacloprid analogs. J. Pesticide Sci. 18,91-98. Wiesner, P., and Kayser, H. (2000). Characterization of nicotinic acetylcholine receptors from the insects Aphis craccivora, Myzus persicae, and Locusta migratoria by radioligand binding assays: Relation to thiamethoxam action. J. Biochem. Mol. Toxicol. 14,221-230. Yamada, T., Takashi, H., and Hatano, R. (1999). A novel insecticide, Acetamiprid. In "Nicotinoid Insecticides and the Nicotinic Acetylcholine Receptor" (1. Yamamoto andJ. E. Casida, eds.), pp. 149-176. Springer-Verlag, Tokyo. Yamamoto, 1., and Casida, J. E. (1999). "Nicotinoid Insecticides and the Nicotinic Acetylcholine Receptor." Springer-Verlag, Tokyo. Yamamoto, 1., Tomizawa, M., Saito, T., Miyamoto, T., Walcott, E. c., and Sumikawa, K. (1998). Structural factors contributing to insecticidal and selective actions ofneonicotinoids. Arch. Insect Biochem. Physiol. 37, 24-32. Yamamoto, I., Yabuta, G., Tomizawa, M., Saito, T., Miyamoto, T., and Kagabu, S. (1995). Molecular mechanism for selective toxicity of nicotinoids and neonicotinoids. J. Pesticide Sci. 20, 33-40. Zhang, A., Kayser, H., Maienfisch, P., and Casida, J. E. (2000). Insect nicotinic acetylcholine receptor: Conserved neonicotinoid specificity of [3HJImidacloprid binding site. J. Neurochem. 75, 1294-1303. Zwart, R., Oortgiesen, M., and Vijverberg, H. P. M. (1992). The nitromethylene heterocycle I -(pyridin-3-yl-methyl)-2-nitromethylene-imidazolidine distinguishes mammalian from insect nicotinic receptor sUbtypes. Eur. J. Pharmacol. 228, 165-169. Zwart, R., Oortgiesen, M., and Vijverberg, H. P. M. (1994). Nitromethylene heterocycles: Selective agonists of nicotinic receptors in locust neurons compared to mouse NlE-1l5 and BC3HI cells. Pestic. Biochem. Physiol. 48,202-213.
CHAPTER
55 Interactions with the gamma-Aminobutyric Acid A-Receptor: Polychlorocycloalkanes and Recent Congeners Gerald T. Brooks University of Portsmouth
55.1 INTRODUCTION Chlorinated insecticides have been with us for 60 years. With the exception of lindane (gamma-hexachlorocyclohexane, HCH) and endosulfan, which are relatively biodegradable and still find extensive uses, most have already been phased out or are being phased out. Their insecticidal properties were discovered at a time when the study of biochemical toxicology was in its infancy. However, the metabolism of dichlorodiphenyltrichloroethane (DDT) was soon discovered after resistance to it appeared in 1947 and the similarities between the actions of DDT and the natural pyrethrins and cross-resistance to them in insects provided the stimulus that soon led to recognition of DDT action on the sodium channel of nerve membrane. The history of lindane and the cyclodiene-related group (collectively polychlorocycloalkanes, PCCAs) is more complex, partly because of the variety of commercially viable insecticides that arose from the early discoveries. Lindane was soon found to be biodegradable but the strongly residual nature of the cyclodiene insecticides and the discovery that aldrin was converted into its stable epoxide, dieldrin, led to the view that these insecticides were inert. Moreover, apart from lindane, there were no other insecticide classes with recognized similar toxic action at the time and the mode of action was completely unknown and would not be revealed until more than 30 years later (ca. 1982)! A comprehensive account of the salient toxicology of these compounds was given in the first edition of this handbook (Smith, 1991). The present account summarizes research on the cyclodiene and related insecticides, which led to an appreciation of their structure-toxicity relationships and in the end to an understanding of their mode of action as noncompetitive antagonists acting in the chloride ion channel of the gamma-aminobutyric acid A (GABA)-receptor. Developments Handbook of Pesticide Toxicology Volume 2. Agents
subsequent to this discovery make research in this area a subject of continuing fascination. For uniformity, the chemical nomenclature used follows that in the first edition. Other systems are in use, for which see Bedford (1974) and Brooks (1974). Simple acronyms and common names will be used wherever possible for chemical compounds as the full chemical names become cumbersome, especially for some of the skeletal rearrangement products so common in this series.
55.2 DISCOVERY OF POLYCHLOROCYCLOALKANE METABOLISM AS A FACTOR IN TOXICITY 55.2.1 BACKGROUND
According to the account of Lauger et at. (1944), DDT was the first molecule rationally designed as an insecticide, based on the known fumigant properties of chlorobenzene and the anesthetic properties of highly lipophilic chloroform. In contrast, the insecticidal properties of technical-HCH (t-HCH) (Bender, 1935) and the first cyclodiene insecticides (Hyman, 1949; Kearns et al., 1945) were discovered as a result of the commercial interest in new uses for readily available chlorine and for hydrocarbons such as benzene and cyclopentadiene, chlorinated hydrocarbons being of general interest, for example, as dielectrics and fire retardants. Thus, Bender added benzene to liquid chlorine in a field and noticed that the product killed insects. Hyman sought new uses for cyclopentadiene; hexachlorocyclopentadiene ("hex") was known to be stable and, at first surprisingly,
1131
Copyright © 2001 by Academic Press. All rights of reproduction in any fonn reserved.
1132
CHAPTER 55
Interactions with the gamma-Aminobutyric Acid A-Receptor: Polychlorocycloalkanes and Recent Congeners
was found to react easily with cyclopentadiene in a Diels-Alder reaction, which led to chlordene, and later with norbomadiene (NB) to give aldrin. The addition of two chlorines to chlordene gave the chlordane isomers, with greatly increased insecticidal potency, whereas allylic chlorination gave heptachlor. A variant of the synthesis of aldrin, in which hexachloronorbomadiene (HCNB) reacted with cyclopentadiene, gave the isomeric isodrin and both compounds underwent chemical epoxidation to their crystalline epoxides, dieldrin and endrin, respectively. These heavily chlorinated insecticides were at that time considered to be rather inert, whereas t-HCH was long known to be readily dechlorinated to trichlorobenzenes, etc., which was its practical use. The potent insecticidal activity of lindane (gamma-HCH; 1, Fig. 55.1) was not established until 1943, 10 years after Bender's original observation, because lindane comprised only 10-15% of t-HCH and was readily lost during purification, which resulted mainly in crystalline, but inactive alpha- and beta-HCH (Slade, 1945). It is unfortunate that due to the high potency of lindane, t-HCH could be used directly and extensively as a practical insecticide, resulting in contamination of the environment with the remaining inactive isomers; lindane itself is relatively biodegradable and continues to be a valuable insecticide. All of these discoveries predate modem biochemical toxicology. Indeed, resistance to modem insecticides, beginning with DDT in 1947, afforded the initial stimulus for research in this area, which subsequently became known as insect, or insecti-
CI#' ~ Z
Cl
Cl
Cl
Cl
(1) LINDANE
~ (4)
Cl
Cl
~ (3)
(2)
~ ~' (S)
~~w (7) HEOM
(8) HCE
(9)
(12)
(10)
Cl6
~13) Figure 55.1
Chemical structures of compounds mentioned in the text.
cide, toxicology and developed in parallel with but somewhat behind mammalian toxicology. One major mechanism of insect resistance to DDT was eventually found to involve its enzymatic dehydrochlorination to DDE (Stemburg et aI., 1954). When it was discovered that certain nontoxic DDT analogs and some other compounds suppressed resistance when co-applied with DDT, studies of the mechanisms of this synergistic effect became an important aspect of insect toxicology and synergists later became standard tools for the detection of metabolic detoxication. Natural pyrethrins were well known to be strongly synergized by various inactive methylenedioxyphenyl derivatives (e.g., piperonyl butoxide: PBO) but, as esters, these insecticides were considered likely to to be hydrolyzed in vivo and the mechanism of the synergistic effect was not understood. Insect resistance to the cyclodienes became evident in the early 1950s, but from research conducted on housefly resistance after 1957 (Brooks, 1960) it appeared not to involve enzymatic detoxication, in contrast to the situation with DDT. Meanwhile, Ryan and Engel (1957) found that carbon monoxide inhibited the C21-hydroxylation of 17 -hydroxyprogesterone by microsomes from the vertebrate adrenal cortex and showed this inhibition to be light reversible; in Klingenberg (1958) reported that rat liver microsomes contained a similar pigment, subsequently called cytochrome P450 (Cy P450) (Estabrook et aI., 1963), that appeared to be important in the metabolism of steroids and drugs, and, in 1965, this pigment was shown to be present in microsomal preparations from insects (Lewis, 1967; Ray, 1967). 55.2.2 LINDANE, ALDRIN, DIELDRIN, ISODRIN, AND ENDRIN AND ANALOGS A new age dawned in insect toxicology in 1960, when Sun and 10hnson published the results of synergism experiments with several classes of organophosphorus insecticides and some cyclodienes (Sun and 10hnson, 1960). The latter showed small factors of either antagonism (for aldrin, 2, Fig. 55.1) or synergism (for dieldrin, 3; isodrin, 4; and endrin, 5) when used in combination with the methylenedioxyphenyl (MDP) synergist sesoxane (sesamex), representative of the well-known pyrethrin synergist structures. They suggested that their results could be explained by the inhibition of metabolic oxidations in vivo, showed that sesoxane inhibited the epoxidation of aldrin in vivo, and postulated that the long-known synergism of pyrethrins by methylenedioxyphenyl compounds resulted, in fact, from inhibition of their oxidative detoxication. The small factors of antagonism for aldrin and heptachlor (6, Fig. 55.1) suggested that epoxidation, their only reported biotransformation at that time, was a bioactivation (toxication) reaction, so that the precursors were possibly propesticides (in current terminology). At this time, numerous nonepoxide cyclodiene analogs were found to be synergized by sesoxane (Brooks and Harrison, 1963, 1964a), indicating that they had intrinsic toxicity of their own, although they may be pharrnacokinetically less efficient
1133
55.2 Discovery of Polychlorocyc1oalkane Metabolism as a Factorin Toxicity
than the epoxides. Also, epoxidation generally produces another toxicant, so that the level of toxic material in the tissues is maintained, whereas it is attenuated if the conversion is a detoxication reaction. Remarkably, considering the small structural change involved, removal of the unchlorinated methanobridge from dieldrin gave the isomeric cyclohexane-derived epoxides HEOM and HCE (7 and 8, Fig. 55.1), which were inactive (HEOM) and poorly toxic (HCE). With sesoxane, however, HCE became as toxic as dieldrin to houseflies, whereas HEOM toxicity was not improved. HCE was then found to be hydroxylated by mixed-function oxidases (MFOs) in vivo, mostly with epoxide ring retention, whereas HEOM suffered only addition of water to the epoxide ring, which was found to be an enzymic process, not inhibited by sesoxane (Brooks, 1966). The relative efficiencies of these pathways (Fig. 55.2) in vitro are shown in Table 55.1, from which it is evident that, for HCE, oxidation is the main route in microsomes from several species; liver microsomes from birds and the rat hydroxylated HEOM to some extent, whereas those from rabbit and pig liver and houseflies hydrated the epoxide ring too rapidly for oxidation to be observed. Similar products and their conjugates are formed in vivo (Chipman and Walker, 1979), and the availability of two routes, one blockable by MFO-inhibiting synergists in vivo in insects, offered the possibility of selective toxicity in favor of vertebrates, which have epoxide ring hydration available as an escape route. Because the hydration of HEOM could not be significantly inhibited in vivo, its intrinsic toxicity was not demonstrable by the use of any known synergist. However, HEOM had the same toxicity as DDT to tsetse flies and was toxic to some species of mosquitoes, which appeared not to hydrate the epoxide ring efficiently, thus establishing that HEOM was intrinsically toxic (Brooks et aI., 1981). These observations verified the Sun-Johnson hypothesis regarding, the action of MDP synergists, completely altered the perspective regarding the "inertness" of cyclodienes, and provided the first firm evidence for the existence of epoxide hydro-
lases (EHs), which Boyland (1950) had suggested to mediate the ultimate metabolism of aromatic hydrocarbons via labile, nonisolable epoxides. Also, sesoxane was found to stabilize certain of the metabolic ally labile cyclodienes in both dieldrinresistant (R-) and dieldrin-susceptible (S-) houseflies but it synergized them only in S-flies, supporting the view that resistance did not involve metabolic detoxication (Brooks and Harrison, 1964b). Korte and Arent (1965) reported that dieldrin-treated rabbits excreted trans-6,7 -dihydroxy-6,7 -dihydroaldrin (t -DDA; 1, Fig. 55.3) in their urine, indicating the epoxide ring opening of dieldrin to occur in vivo. Dieldrin and the heptachlor epoxides were subsequently found to be hydrated slowly in microsomal preparations from the livers of rabbits and pigs (Brooks and Harrison, 1969b; Brooks et aI., 1970). This challenged the hitherto prevailing view that cyclodienes would accumulate indefinitely in the tissues of treated animals, a challenge reinforced by the pharmacokinetic studies of Ludwig et al. (1964) and Robinson et al. [cited in Brooks (1969)] on mammals, birds, and marine organisms. The continual improvement in techniques for preparing microsomes from liver and other animal tissues afforded opportunities for the rapid examination of the likely phase 1 metabolites and stimulated interspecies comparisons of cyclodiene metabolism (Craven et aI., 1976;
'16~
(1)
Table 55.1 Oxidation Versus Hydration of HCE and Hydration of HEOM by Microsomes from Vertebrate Liver and Houseflies
Pigeon
Quail
Rat
Housefly
Rabbit
Pig
HCEa Oxidationb
1.0
0.7 c
1.3
2.0
1.0
Hydrationb
0
O.02c
0.03
0
0.33
1.0
0.005
0.06
5.74
37.0
1.3
HEOMd
Hydration ratee
46.0
100
aVertebrate liver and housefly abdomen microsomes (+NADPH); incubations for birds 90 min at 42°C; others 30 min at 37°C. bConversion of epoxide (HCE), percent/min at pH 7.4. cll,OOO g supernatant. dlncubations with pig liver and housefly microsomes at 30°C for 30 min (pH 8.4), with rabbit and rat liver microsomes at 37°C (pH 7.4), and bird liver microsomes (El Zorgani et aI., 1970) at 42°C (pH 7.4). eRelative to pig liver (100 is equivalent to 31 ~g HEOM-diol formed/mg microsomal protein/min).
l ~bH H~~,
H20
HEOM (3)
p, B
EPOXIDE HYDROLASE
H, p, R, RA
1
0
H"O
2
~------ NOT READILY BLOCKED---------~ BY INHIBITORS
rJ\ Y(4y
.OH
OH
Figure 55.2 Alternative phase I metabolic pathways for HCE and HEOM and the effect of inhibitors. Epoxide ring hydration (1 gives 2 and 3 gives 4) is enzymic and only one enantiomer of HCE is hydrated: A, all species examined; B, some birds; H, housefly; P, pig; R, rat; Ra, rabbit. Pathways in vitro and in vivo for housefly; for mammals and birds, results are for liver homogenates and microsomes but similar products and their conjugates are formed in vivo.
1134
CHAPTER 55
Interactions with the gamma-Aminobutyric Acid A-Receptor: Polychlorocyc1oalkanes and Recent Congeners
El Zorgani et aI., 1970; Slade et aI., 1975). In contrast to alphaand especially beta-HCH, lindane proved to be quite biodegradable, and its complex metabolism in insects and vertebrates via dehydrochlorination and oxidation to chlorinated phenols and their conjugates is well documented (Brooks, 1974; Smith, 1991; Ullman, 1972). In contrast to the conversion of aldrin and isodrin into their stable 6,7-epoxides, dieldrin and endrin (Giannotti et aI., 1957; Kunze and Laug, 1953), 6,7-dihydroaldrin (9, Fig. 55.1) and 6,7-dihydroisodrin (10) lacking the olefinic double bond, are monohydroxylated in the 6-(7-)position by microsomal oxidases (Brooks, 1966; Brooks and Harrison, 1969a) and in vivo in houseflies. The synergism of these dihydro-compounds by sesoxane in this insect indicates that this is a detoxication, in contrast to the epoxidation reactions. This result also suggests that the dihydro-compounds are intrinsically toxic, although they act more slowly than the epoxides. Notably, these dihydro-compounds have the same synergized toxicity as photoisodrin (11), the complete cage rearrangement product of isodrin, which is as polar as endrin (based on Rp values) and acts more rapidly than the dihydro-compounds. Also, the 5,8oxirane (12), which has a "built-in" epoxide function, is similar in toxicity to dieldrin when synergized and acts more rapidly than the other dihydro-compounds discussed (Brooks, 1966; Brooks and Harrison, 1963), again supporting the view that such compounds are intrinsically active and that epoxidation (or an appropriate increase in polarity) improves the pharmacokinetic properties of these molecules, besides maintaining the total level of toxicant in the tissues. This does not, however, rule out a possibly more efficient binding of the epoxides at the site of action. Insect poisoning by commercial cyclodienes was not at first recognized to be reversible because their persistence in the tissues led to eventual death due to desiccation and starvation, without any recovery. Housefly poisoning by HCE was noted to be reversible, however, and even insects poisoned with HCE/sesoxane combinations would occasionally recover after prolonged periods of knockdown, although their wing musculature appeared to be permanently damaged. A further complication arose when 6,7-dihydroxydihydroaldrin (aldrin-trans-diol; t-DDA; 1, Fig. 55.3) (Wang et aI., 1971) and subsequently the corresponding cis-DDA (Burt, 1973) were found to be rapidly neuroactive when applied to isolated nerve ganglia of the American cockroach (Periplaneta americana), in contrast to dieldrin, which acted significantly more slowly. Wang et al. (1971) then suggested that the slow action of dieldrin was related to a requirement for its conversion into trans-DDA, as the active neurotoxicant liberated at the site of action by dieldrin hydration. Small amounts of these diols were later reported to be dieldrin metabolites in this insect (Nelson and Matsumura, 1973). However, t-DDA caused prostration only slowly when injected into cockroaches, quite different from the rapid action of dieldrin in vivo. The bioactivation hypothesis was further supported by the neuroactivity of t-DDA observed on frog nerve-muscle preparations (Akkermans et aI., 1974, 1975). The findings for cock-
;;c:. ty
COOH ___ _ . . CONJUGATION
eOOH
rFJH+ C~GA::e(:~:i)
PH=D=~:~ ::~HCl s Hmo(V)
Q:;;
i
(1)
~:~) 5(V)\{!) .".~CiS-
O:H
JI( ) . Cl h~:rition v
~
RCv) (1)
I (v)
(m)
environment:
\
mfo
S(v)
\..s mfo
J"
,
r
+
(!~~v
'I
HO
ENDRIN
H
0
!
Pi
cl
H
(l)
'KLEIN KETONE'
mfo
~GLUCURONIDE
OH~R(V)
~ "-:fO
HS-""
/mfo
4;iii'Cl
0
B
Cl~l Cl OIOL
B
Cl
(§)
8
trans-
, H
Cl "OB
\\R~:~'" , /> :~ Cl
~o ?....
'"' lR~V)(i) =M
Cl
\
r
1
:~~"'O
\ 'mf°ll~Cl t Cl j/ )1
C
man (v) R(V) (i)
R(v}
rearranqement[Cl~
DIELDRIN
micerv)
H(v)
(ri)
",,0
".
0
'.A ~
s
f1'r;I'-:/",
r-
~ (Z) 0
.'-Q.'
Figure 55.3 Biotransformation routes of dieldrin, photodieldrin, and endrin. Wavy line indicates only partial ring structure shown. H, housefly; I, some insects; mo, mosquito; m, microorganisms; P, pig; Ra, rabbit; R, rat; S, sheep; v, in vivo; i, in vitro; S ---+, sesoxane inhibits. See text for references.
roach were confirmed (Schroeder et aI., 1977; Shankland and Schroeder, 1973), but, based on the less intense neuroactive effect of the diols and their very slow action in vivo, the diols were concluded to be detoxification products in the cockroach. Both diols are produced as metabolites of dieldrin by rats and mice and appear to be detoxification products in these mammals. These pharmacokinetic studies raise questions about possible internal barriers to the penetration of such molecules and their metabolites to critical sites in the nervous system. Similar problems are apparent throughout the series and are difficult to resolve experimentally. Moreover, a particular metabolite might be a bioactivation product in one species but a detoxification product in another. The tendency for molecular rearrangements in the environment (e.g., from exposure to sunlight) and in vivo has complicated investigations on residues and metabolites. Photoconversion products are frequently more toxic than the parent insecticides and may themselves be further metabolized; for example, photodieldrin (PD; 2, Fig. 55.3) is oxidatively dechlorinated to the pentachloroketone (3, Fig. 55.3; "Klein's ketone";
55.2 Discovery of Polychlorocycloalkane Metabolism as a Factorin Toxicity Klein et al., 1970) in rats and insects (Baldwin and Robinson, 1969; Baldwin et al., 1972; Khan et al., 1970; Matthews and Matsumura, 1969); PD has a much shorter half-life (23 days) than dieldrin (10-13 days) in rat adipose tissue but is two- to four-fold more toxic to rodents and insects (Table 55.2). Dieldrin-treated rats excrete 9-hydroxy-dieldrin (9HD; 4, Fig. 55.3) in the feces and the pentachloroketone in the urine (Richardson et al., 1968), and these are considered to arise by alternative modes of attack from beneath the ring system (Fig. 55.3). The same pentachloroketone (3) was produced, along with varying amounts of 9-HD and cis- and trans-DDA, in American cockroaches, German cockroaches (Blattella germanica), and houseflies (Nelson and Matsumura, 1973). The pentachloroketone (3) was reported to be more toxic than photodieldrin to mosquitoes and houseflies (Khan et al., 1970) but less toxic and slower acting than PD to the German cockroach (Kadous and Matsumura, 1982; Reddy and Khan, 1977) indicating that PD itself is the active toxic ant in this insect. PD acted four-fold more rapidly (LDso, 0.01 ).Lg/insect) than dieldrin (LDso, 0.05 ).Lg/insect) and two-fold more rapidly than the pentachloroketone (LDso, 0.13 ).Lg/insect) observations that suggest it has pharmacokinetic properties more favorab1e for toxicity than the other compounds. 9-HD (4) appeared to be more toxic (LDso, 0.02 ).Lg/insect) than dieldrin to the German cockroach and may contribute to dieldrin's toxicity in this insect; the cis- and trans-DDAs appeared to be relatively nontoxic when injected. From other experiments on the American cockroach, it seems clear that these metabolites can enter the nerve cord from the insect body and are also produced in small amounts by metabolism in the nerve itself. Isodrin was found to be epoxidized to endrin (Fig. 55.3) in houseflies (Brooks, 1960) and subsequently by liver microsomes from rats and rabbits, as a result of mixed-function oxidase (MFO) action (Nakatsugawa et al., 1965; Wong and Terriere, 1965). Endrin incubated with pig or rat liver microsomes in the presence of reduced nicotinamide adenine dinucleotide phosphate (NADPH) gave a monohydroxy-derivative, formation of which was inhibited by sesoxane, indicating MFO involvement (Brooks, 1969). It soon became clear from mammalian studies that the inversion of the unchlorinated norbornene nucleus in isodrin and endrin (as compared with aldrin and dieldrin) exposes this ring to enzymatic hydroxylation in vivo and greatly increases the rate of elimination of these compounds from mammalian tissues, in contrast to their behavior in insect tissues. Endrin is generally more toxic to vertebrates and less toxic to some insects than dieldrin; whereas the latter undergoes 9-hydroxylation syn to the epoxide ring and 9-HD (4, Fig. 55.3) is eliminated by conjugation in mammals, endrin is both syn- (slowly) and anti- (rapidly) hydroxy1ated; the antiderivative (5, Fig. 55.3) is rapidly conjugated and excreted but the syn-isomer (6, Fig. 55.3) is further oxidized to 9-keto-endrin (9-KEN, also called 12-keto-endrin; 7, Fig. 55.3), a remarkable example of the profound influence of stereochemistry on metabolic pathways. Bridge-end (tertiary) hydroxylation also occurs and endrin trans-diol is a minor metabolite. 9-KEN is some five-fold more
1135
toxic than endrin to rats and appears to be the ultimate toxic metabolite of endrin (Bedford et al., 1975a; Hutson et al., 1975). Species differences are evident, since Kadous and Matsumura (1982) reported the order of endrin metabolite toxicity to male German cockroaches as 5-0H > anti-9-0H > 9-keto-, whereas the order on topical application was 9-keto '" syn-9OH > endrin » anti-9-0H to houseflies and 9-syn-OH > 9-keto-> endrin » anti-9-0H to blowflies (Brooks and Mace, 1987). Also in this report, syn-9-hydroxydieldrin (9-HD; 4) was essentially nontoxic to houseflies and blowflies, whereas the order 9-oxadie1drin (9-0D; 13, Fig. 55.1) '" dieldrin > 9-ketodieldrin (9-KD; 8, Fig. 55.3) and 9-oxadie1drin (9-0D) ~ 9-KD > dieldrin, respectively, was found for houseflies and blowflies. Toxic 9-KD is apparently not formed from 9-HD in vivo, possibly because, in contrast to the situation with endrin, steric hindrance prevents enzymic attack on the hydroxyl group. Each set of toxicities lies within a narrow range and the toxicities of 9-KD and 9-0D might be expected to be similar, because 9-0D is an isostere of 9-KD, in which -C=O has been replaced by the more compact 5,8-bridged oxirane structure. These results show that several of the oxidative metabolites of these insecticides retain insect toxicity and may contribute to the toxic effect of the parent insecticides. 55.2.3 HEPTACHLOR, CHLORDENE, DIHYDROHEPTACHLOR, CHLORDANE,ANDISOBENZAN Further chlorination of the feebly toxic chlordene, the DielsAlder adduct of "hex" and cyclopentadiene, gave heptachlor, the dihydroheptachlor isomers (Table 55.3 and Fig. 55.4), and the chlordane isomers (Fig. 55.5). The nontoxic adduct of "hex" and cis-2-butene-1 A-diol, namely 5,6-bis(hydroxymethyl)hexachloronorbornene-2-ene, is the precursor to which isobenzan, endosulfan (Fig. 55.6), bromocyclen (Bromodan®), and chlorbicyclen (Alodan®) (Fig. 55.7; 20 and 21, respectively) are related. The last two compounds were once used to control animal ectoparasites because of their low mammalian toxicity; endosulfan is still used extensively today, whereas isobenzan was discontinued in 1965. A preparation of heptachlor is mentioned in the original Hyman patent (Hyman, 1949) on chlordane. Numerous investigations from 1951 demonstrated the formation of heptachlor exo-epoxide, m.p. 160°C (HE160; Fig. 55.4); the less insecticidal endo-epoxide, m.p. 90°C (HE190) is not formed in vivo but can be obtained indirectly by chemical synthesis. The biotransformations of chlordene and heptachlor involve allylic hydroxylation for chlordene, hydrolysis of allylic chlorine for heptachlor, epoxidation (Miles et al., 1969), and epoxide ring hydration (Brooks, 1966; Fig. 55.4). Microorganisms can degrade heptachlor by removing the allylic chlorine, either reductively or by hydrolysis, so that the degradation routes for chlordene can then be followed (Miles et al., 1969); in some soils, the production of 1-hydroxychlordene (1, Fig. 55.4) is comparable to HE160 production. The hydroxylated metabolites appear to
1136
CHAPTER 55
Interactions with the gamma-Aminobutyric Acid A-Receptor: Polychlorocycloalkanes and Recent Congeners
Table 55.2 Toxicity Data for Some PolychlorocycIohexane Insecticides and Their Transformation Products
Compound
Rodent acute oral LDSOa
Topical 24-h LDsOb
(mg/kg)
housefly
Lindane
90-190
Aldrin
38-60
6,7-Dihydroaldrin
1.5 1.5 40 (3.0)
5,8-0xadihydroaldrin Dieldrin
(~g1g)
30 (1.5) 47
1.0
77 (m) HCE
>400
90 (2.0) 200-400 (m)
HEOM Photodieldrin
>500 10
0.12
7 (m) Didechlorodieldrin (DD)
0.2
0.9 1.4
9-Hydroxydieldrin (9-HD)
>400 (m)
trans-Dihydroaldrin-diol (t-DDA)
1,250 (m)
Isodrin
750 >750
12-17
6,7-Dihydroisodrin
3.0 39 (4.0)
Photoisodrin
>2,000
15 (3.0)
Endrin
5.6
2.0
29 (m) 9-Keto-endrin
1.0
0.95
anti-9-Hydroxy-endrin (AHEN)
2.5-5.5
syn-9- H ydroxy-endrin
1.2
Heptachlor epoxide (HEI60)
60
>100 1.2 1.0
Heptachlor epoxide (HE90)
6.0 2,400-4,600
1-H ydroxychlordene
Inactive 50 (20)
Chlordene Chlordene exo-epoxide
35 (4.0)
1-Hydroxy-chlordene exo-epoxide
Inactive
trans-Chlordane
11
1,100 500-600
eis-Chlordane
4.0 5.0
alpha-Endosulfan
76
beta-Endosulfan
240
9.0
Endosulfan sulfate
76
9.5
Endosulfan diol
>15,000
>500
Endosulfan ether
>15.000
>500
alpha-Hydroxy-endosulfan ether
1,750
>500
Endosulfan lactone
306 (m)
>500
Isobenzan
3-10
1.0
6.0 (m) Bromocyclen (Bromodan®)
13,000
11.5
Chlorbicyclen (Alodan®)
15,000
15.5
Mirex Chlordecone Toxaphene (technical) Toxaphene (component B) a For
600->3,000 125 90-270 75 (m; ip)
rat unless marked (m) for mouse. bparenthetic values in housefly column are toxicities measured with sesoxane (5 ~g), preapplied before the insecticide to inhibit microsomal oxidases. Data from Brooks and Harrison (1964a), Buchel et al. (l966a, 1966b), Jager (1970), Khan et al. (1970), Korte (1967), Maier-Bode (1968), Miles et al. (1969), and Bedford et al. (1975a), Smith (1991).
55.2 Discovery of Polychlorocycloalkane Metabolism as a Factorin Toxicity
be detoxification products in mammals. This is difficult to prove in insects, however. Chlordene (2, Fig. 55.4) and its exo-epoxide (3, Fig. 55.4) have a weak housefly toxicity, which is synergized lO-fold by sesoxane, suggesting that the biotransformations observed in microsomal preparations are detoxications (Brooks, 1966; Brooks and Harrison, 1964a, 1967a, b). Is heptachlor much more toxic than chlordene because the allylic chlorine inhibits hydroxylation in this position and also ensures that heptachlor is converted into the metabolically stable epoxide ?-a question reminiscent of the aldrin/dieldrin situation. The view that heptachlor is intrinsically toxic is supported by the toxicity (Table 55.3) of the alpha- and beta-dihydroheptachlor isomers (Fig. 55.4), formed by the addition of hydrogen chloride to the double bond of chlordene. Their housefly toxicity also is synergized by sesoxane, which suggests that metabolic hydroxylation, which for them replaces the epoxidation of heptachlor, results in detoxication. Beta-dihydroheptachlor (beta-DH; Fig. 55.4; 2, Table 55.3) is particularly interesting because of its low mammalian toxicity (Buchel et aI., 1966a, 1966b). In the presence of NADPH, pig liver microsomes converted alpha, beta-, and gamma-DH into a variety of hydroxylation products, which are illustrated for beta-DH in Fig. 55.4. These were chlorohydrins, obtained by simple hydroxylation of the cyclopentane rings; alcohols, formed by elimination of the single chlorine atom on the cyclopentane ring; dihydroxy-compounds; and a ketone (e.g., 2-keto-dihydrochlordene from beta-DH, which may afford the corresponding alcohol via a ketoreductase reaction). The 2-0H-dihydrochlordene excreted by rats fed beta-DH (Korte, 1967) may arise in this way. Sim-
1137
ilar metabolites were produced in housefly microsomes, although no dihydroxy-compounds were detected. Sesoxane inhibited the hydroxylations, which doubtless explains the synergism against houseflies observed in vivo (Table 55.3). The metabolism of the two chlordane isomers, alpha- (= trans-l ,2-dichlorodihydrochlordene) and beta- (= cis-l,2-dichlorodihydrochlordene), is complex (Fig. 55.5). Either isomer might give heptachlor by dehydrochlorination and hence HE160 and all the metabolites arising therefrom. In fact, the metabolites in rats include l-exo,2-dichlorochlordene (1, Fig. 55.5), oxychlordane (2), l-exo-hydroxy-2-chlorochlordene (3), l-exo-hydroxy-2-chloro-2,3-epoxychlordene (4), l-exohydroxy, 2-endo-chlorodihydrochlordene (chlordene chlorohydrin), 1,2-trans-dihydroxydihydrochlordene, and the metabolites of heptachlor (Brimfield and Street, 1979; Brimfield et aI., 1978; Tashiro and Matsumura, 1977). A similar series of compounds was excreted in the form of unidentified conjugates in the urine of rabbits treated with these chlordane isomers (Balba and Saha, 1978). These biotransformations demonstrate the remarkable versatility of the drug-metabolizing enzymes. In particular, the formation in rats of oxychlordane (Fig. 55.5), analogous to heptachlor epoxide and said to be more toxic than trans-chlordane (Street and Blau, 1972), is a bioactivation due to the unexpected formation of a stable epoxide in vivo, presumably following an enzymatic desaturation that introduced a 2,3-double bond. There was no evidence for epoxide formation from either cis- or trans-chlordane in houseflies, however. Notably, transchlordane was three-fold less toxic than eis-chlordane to this insect, and neither isomer was synergized by sesoxane (Brooks
Table 55.3 Toxicities of Dihydroheptachlor and Chlordane Isomers to Housefly and Mouse exoB, exo-
endo-
Compound
A
(1) a-
fJ-
(2)
Housefly LDso
Housefly LDSO (flg/f1y with
(flg/fly)a
sesoxane)a
Mouse acute oral LDso (mg/kg)
B
C
D
Cl
H
H
H
0.26
0.015
1,285
H
Cl
H
H
0.16
0.015
>9,000
(3) y-
H
H
Cl
H
1.8
0.07
>6,000
(4)b
Cl
H
Cl
H
0.22
0.22
1,100
(5)b
Cl
Cl
H
H
0.08
0.08
(6)
H
Cl
Cl
H
0.04
>600
(7)b
Cl
H
H
Cl
0.04
31
Alodan
0.31
0.05
Dieldrin
0.02
0.02
aTopicaI application; sesoxane applied (5 flgl20 mg fly) before insecticide. b 4, trans-chlordane; 5, eis-chlordane; 7, 8-chlordane. cLDSO for rat. Data compiled from Brooks and Harrison (l964a, 1967b) and Buchel et al. (l966a, b).
500-600
15,000e 75-100
1138
CHAPTER 55
Interactions with the gamma-Aminobutyric Acid A-Receptor: Polychlorocyc1oalkanes and Recent Congeners
tY
P::(H
!~:;;
;::!iJ) (J
/
~7
~O
rrY/°PI(i)
'l'"
~~~)
B
3..
~
~ HE 160
~ H(,t~~Yh ~ Cl
Cl +
td
RCv)
p:j~
PICI)
CHLORDANE
3
lC~LPHA_(TRANS_) BETA_(CIs_)~16 ~: 1
CHLORDANE
PI(I) H(I)
p:jH
rrY
CI
MCv)(I); MCV)
rr-,
1-'.--1 ALPHA- ~ GAMMA- el
BETA-
I
+
DIHYDROHEPTACHLOR ISOMERS
~
~
~CHLOROHYDRINS,
~V) Cl
PCI) 2,3-EPOXIDE 4--~ 'OXYCHLORDANE' R(v) Cl (2)
Y'V
HEPTACHLO~j{V)(I)J
SYNTHESIS CHCl GAS)
RA(V~V)
~~(V)
~CH~~~DENE ' ~ \
R(v)
'"""011;"'1"
Cl
E,\:
-
OH
~:~)
W 6
(1)
R(v)
RA'vy /
,RC?):
rr;,
1
1
M : I (I) I
I
~
RCI)
RA'v)
OH
Cl
OH
;J;;jq) :
/H(vHI)
r
HEPTACHLOR, 3-oH-CHLORDANE
~OH
;:t;J0H
~H(V){I}~~ (2) H(v){l)
~
HE 190
ETC.
AS FOR BETA-ISOMER
OH Figure 55.4 Biotransformations of chlordene, heptachlor, and the dihydroheptachlor isomers. B, bacteria; H, housefly; I, some insects; M, mammals generally; Pi, pig; R, rat; m, microorganisms; Ra, rabbit; E, abiotic conversion; v, in vivo; i, in vitro. All structures contain the fully chlorinated norbomene moiety.
and Harrison, 1964a), indicating that intrinsic toxicities were being measured. Alpha- and beta-DH were as toxic as heptachlor when synergized; synergized gamma-DH was four-fold less toxic than synergized alpha- and beta-DH and as toxic as eis-chlordane (Table 55.3). This suggests that the 2-endochlorine atoms in trans-chlordane and gamma-DH contribute less to toxicity than the exo-chlorines present in alpha- and beta-DH and eis-chlordane. Moreover, an additional chlorine introduced into the 2-exo-position of gamma-DH (Table 55.3) increases its toxicity more than 40-fold, so that the resulting gem-dichloro-compound is as toxic as beta-DH having the single exo-chlorine in this position. Does this extra exo-chlorine simply reduce the possibilities for metabolic detoxification that are more likely for gamma-DH (exo-side of the ring exposed to enzymatic attack) and transchlordane, or do exo-chlorines increase the affinity of these molecules for a critical binding site in the nervous system? That synergized gamma-DH is as toxic as eis-chlordane (unaffected by sesoxane) may suggest that metabolism is the only factor involved and that the exo- or endo-disposition of the chlorines is immaterial. There is also the interesting question of the
(.!.>
Figure 55.5 Biotransformations of the eis- and trans-chlordane isomers. Abbreviations as in Fig. 55.3. All structures contain the fully chlorinated norbomene moiety.
role of symmetry; the most insecticidal compounds in this series are beta-DH, the 2,2-gem-dichloro-analog (gamma-DH), and l-exo, 3-exo-dichlorodihydrochlordene (delta-chlordane; Table 55.4), all having a plane of symmetry, in contrast to the other molecules discussed, for which the enantiomers may differ in toxicity (see later discussion on the heptachlor epoxide enantiomers in Section 55.3.1). Production of isobenzan (Telodrin® ceased in 1965 (Jager, 1970), but this molecule (Fig. 55.6) remains of theoretical interest as a cyclic ether analog of gamma-chlordane, which, like the latter, has high insect and mammalian toxicity. Enzymatic attack on the chlorinated cyclic ether structure of isobenzan analogous to the biotransformations noted for the chlordane isomers results in hydrophilic metabolites such as derivatives of the gamma-hydroxy-acid (1, Fig. 55.6), which afforded the lactone (2) and alcohol (3) on hydrolysis. Alternatively, these might arise directly by oxidative or hydrolytic elimination of chlorine atoms from the cyclopentane ring. Of particular interest because they illustrate the variety of structures having toxicity in this series is the mixture of two interconvertible isomeric ketones (14 and 15, Fig. 55.7), with high insect and mammalian toxicity (housefly LD50, 0.5 j..lg/g; rat LDso, 7 mg/kg) , which can be obtained chemically from 5,6-bis(hydroxymethyl)-HCNB ("endosulfan-diol"). Transannular dehydrochlorination affords an even more toxic cage ketone (16, Fig. 55.7; housefly LDso, 0.25 j..lg/g; rat acute oral LDso, 1.0 mg/kg). These analogs of isobenzan are more compact versions of the various cage molecules formed from dieldrin and provide further evidence that the dichloroethylene moiety of cyclodienes can be replaced by other polar moieties without loss of toxicity and with increased toxicity in some cases.
55.2 Discovery of PolychlorocycIoalkane Metabolism as a Factorin Toxicity
1139
Table 55.4 Insect Toxicity of Aldrin and Dieldrin Relatives, Including Some Molecules with Fewer Chlorine Atoms
x =
y
=
carbon,
except for compounds (3 ) and (4 ) Chlorination in aldrin analog Chemical
2
3
4
lO-syn
lO-anti
HFa
GRb 2.0
(1)
---------
6-Cl
- - - - - - (aldrin)
0.55
(2)
---------
6-Cl
---
---------
6,7 -epoxide: dieldrin
1.0
1.0
(3)
---------
6-Cl
---
---------
6,7-N=N-
3.6
5.3
(4)
---------
---------
6,7-N=N(--+ 0)-
4.45
2.1
lOa
2.1
9.3
6-Cl
---
(5)
H
H
4
lOs
(6)
H
H
4
lOs
lOa (6,7-epoxide: DD)
3.8
8.0
(7)
H
H
4
H
H
0.03
0.8
(8)
2
3
4
H
lOa
0.08
0.65
(9)
2
3
4
lOs
H
0.02
Inactive
a Housefty toxicity compared with dieldrin (1.0) by direct spray.
bGerman cockroach toxicity compared with dieldrin (1.0) by exposure to dry films on paper. Data compiled from Soloway (1965).
55.2.4 ENDOSULFAN (THIODAN) Technical endosulfan is a 7:3 mixture of the alpha- (m.p. lO9°C) and beta- (m.p. 213°C) isomers, the former (Fig. 55.6) having an "extended," dieldrin-like structure (see also Section 55.4.3) and the latter having a cagelike structure resembling endrin stereochemically. The alpha-isomer is more toxic than the beta-isomer to mammals and houseflies; both are oxidized in vivo to endosulfan sulfate (4, Fig. 55.6), which resembles beta-endosulfan stereochemically and has similar toxicity to alpha-endosulfan, so that this conversion is analogous to the aldrin-to-dieldrin one. The cyclic sulfite (and sulfate) ester structures completely alter the behavior of the endosulfans, which disappear quite rapidly from living tissue, partly by hydrolysis to the parent nontoxic endosulfan-diol and metabolites similar to those formed from isobenzan. The sulfate is formed faster from the alpha- than from the beta-isomer in houseflies and is as toxic as beta-endosulfan to these insects (Barnes and Ware, 1965); cyclodiene-resistant flies eliminated these isomers more rapidly than normal (S-) flies, but the tissues contained only the toxic sulfate, which also appears in the body fat of mammals but disappears rapidly when exposure ceases. Endosulfan-treated locusts excreted the sulfate, endosulfan ether, alpha-hydroxy-endosulfan ether (3, Fig. 55.6), and the corresponding lactone (2). Endosulfan-treated mice stored the sulfate transiently in their fat and excreted endosulfan, the sulfate, and the parent diol in feces (Maier-Bode, 1968). It is evident that endosulfan is a relatively nonpersistent compound in mammals (Dorough et aI., 1978) and has generally favorable environmental properties, apart from high fish toxicity, which requires caution in aquatic situations. With the exception of the
toxic sulfate, metabolites of endosulfan isomers are undoubtedly detoxication products.
55.2.5 TOXAPHENE, MIREX, CHLORDECONE (KEPONE) Toxaphene (camphechlor) is a complex mixture of some 177 compounds obtained by chlorinating camphene to a 67-69% chlorine content (Pollock and Kilgore, 1980; Saleh et aI., 1979). The identified compounds are actually chlorinated bornanes arising from the Wagner-Meerwein rearrangement of the camphene skeleton, among which the octachloronorbornanes; 2,2,5-endo, 6-exo-8,8,9,1O-octachloro-norbornane (17, Fig. 55.7) and 2,2,5-endo,6-exo-8,9,9, lO-octa-chloronorbornane (18, Fig. 55.7), are highly potent, with mouse ip LDso values of 2-3 mglkg (Turner et aI., 1977). The less toxic 2,2,5endo,6-exo,8,9,lO-heptachloronorbornane (19; compound B; LDso, 75 mglkg) was potentiated eight-fold by PBO administered prior to the insecticide, suggesting the possibility of oxidative detoxication mechanisms for this compound. Experiments with rat liver preparations confirmed metabolism by MFO, and the formation of glutathione and glucuronide conjugates (Chandurkar and Matsumura, 1979) could be demonstrated (see Smith, 1991). The positioning of the added chlorine substituents in compound B seems to be critical; at the 3-exo-position and in the 10-chloromethyl moiety, an additional chlorine greatly reduces mouse toxicity, as does the combination of 3-exochlorination and 5,6-dehydrochlorination, to give a vinylic chlorine atom. Notably also, the simpler (less bulky) molecules
1140
CHAPTER 55 Interactions with the gamma-Aminobutyric Acid A-Receptor: Polychlorocycloalkanes and Recent Congeners
~o, ~ ~~S02
•
(4)ENDOSULFAN SULFATE
1
HYDROLYSIS
/() ENDOSULFAN 'ETHER'
ENVIRONMENT BIOTIC, ABIOTIC
~
i
Cl
,
mfo ?
CI~I 0 Cl ,0'7),' _LOC_U_ST_)~ I
0"
HYDROLYSIS
(l)
~o
0
(2) ENDOSULFAN 'LACTONE'
Cl ALPHA-ENDOSULFAN
Y1'CH20S03H p......COOH
\ '4
'-\ OXIOOREDUCTASE?
/~.
~CI ISOBENZAN
Figure 55.6 Major transfonnations of alpha-endosulfan and isobenzan. All structures contain the fully chlorinated norbomene moiety. Note that - -0- - indicates the skewed ("trans") position of the second oxygen in the "twist-chair" (asymmetric) configuration of alpha-endosulfan (Schmidt et aI., 1997).
hexachloronorbornene-2,5-diene and heptachloronorborn-2ene used to prepare cyclodiene insecticides lack toxicity, which only appears when halomethyl groups are introduced into the nucleus as in bromocyclen (Bromodan®; 20, Fig. 55.7) and chlorobicyclen (Alodan®, 21). Both are synergized lO-fold by sesoxane in houseflies (Brooks and Harrison, 1964a; Table 55.2) and are quite good insecticides with very favorable mammalian toxicity (rat acute oral LDsos, 13000-15000 mg/kg); that is, they appear to be considerably more selective (insect versus mammal) than the most toxic components of toxaphene. Mirex (22, Fig. 55.7) is the fully chlorinated cage molecule, formed by the self-condensation of two molecules of "hex," and might be expected to be rather resistant to enzymatic attack. Animal tissue levels plateau only slowly on exposure and decrease very slowly when exposure ceases. One chlorine atom is reductively replaced in the environment to give photomirex (8-monohydro-mirex), which appears to behave like mirex in the rat (Chu et aI., 1979; Hallett et aI., 1978). Reductive dechlorination can occur in vivo; 2,8-dihydromirex and 5,10dihydromirex have been identified as rat metabolites. Whereas 2,8-dihydromirex does not appear to be further metabolized, 5,IO-dihydromirex appears to be converted into more polar metabolites, which appear in rat urine (Yarbrough et aI., 1983). Mirex has low mammalian toxicity (rat oral LDso ranging from 600 to > 3000 mg/kg) and its signs of poisoning differ from those produced by the less chlorinated cyclodienes. The metabolism of mirex in houseflies is equally slow, and Shankland (1982) compared its slow insecticidal action with the delayed onset of dieldrin poisoning discussed earlier. The onset of poisoning following topical application of
lethal doses of mirex to the American cockroach occurred only after 3 days. Moreover, when isolated sixth abdominal ganglia were irrigated with suspensions of 5 x 10-4 M mirex for 4 h, there was no change in the patterns of spontaneous activity or elicited postganglionic responses. Ganglia excised from symptomatic cockroaches showed, however, spontaneous after-discharge behavior characteristic of poisoning following dieldrin treatment. Hemicholinium-3, which depletes Ach stars, eliminated the neuroactivity in giant fibers, but the ganglia remained responsive to nicotine, as is found in dieldrin poisoning. Because mirex appears to be highly resistant to biotransformation, Shankland concluded that the delayed action was unlikely to involve a requirement for bioactivation and must arise from the intrinsic properties of this highly chlorinated molecule, such as slow penetration through diffusion barriers in the insect central nervous system. Chlordecone (Kepone, 23, Fig. 55.7) differs from mirex in having a carbonyl group, which is probably responsible for its moderately rapid clearance from animal tissues. In humans and pigs, this is via the alcohol (chlordecol), and a cytosolic ketoreductase, which can effect this reduction, has been found in gerbil and human liver (Molowa et aI., 1986). Bloomquist and Shankland (1983) found that chlordecone produced the same signs of poisoning as mirex in the American cockroach and concluded that chlordecone has the same mode of action as dieldrin, although, like mirex, it acts more slowly. From experiments on the displacement of [3H] picrotoxinin (PTX) binding by mirex and chlordecone from American cockroach head membranes, Tanaka et al. (1984) concluded that chlordecone interacts with the PTX-binding site, as expected, whereas mirex was much less potent in this respect; moreover, dieldrin-resistant
55.3 Structure-Toxicity Relationship and Mode of Action
Cl
II
~
Cl
CIO
CI~CI Cl
~~
11
o U4,15)
1141
10
C
O~
(6)
Cl
Cl ..... ,~~__ Cl CH 2CI
(2() Bromodan®
(8) Toxaphene
~ (23)
Mirex
'0
Chlordecone (Kepone)
Cl
Figure 55.7
Chemical structures of compounds mentioned in the text.
German cockroaches were resistant to chlordecone but not to mirex. Chlordecone is also known to have inhibitory effects on neurotransmitter uptake in mammals and such an action may also contribute to its insect toxicity.
55.3 STRUCTURE-TOXICITY RELATIONSHIP AND MODE OF ACTION 55.3.1 FULLY CHLORINATED CYCLODIENES: SUBSTITUTED HEXACHLORONORBORNENES (HCNB) Soloway (1965) published a comprehensive review on the structure-activity relationships of cyclodiene insecticides at a time when information on their metabolism was just beginning to appear, so his review makes only passing reference to the possible influence of metabolism but provides a
great deal of information about toxicity trends in numerous series of cyclodiene analogs. Initially, he emphasized the similarity between heptachlor epoxide HE160/cis-chlordane and HE90ltrans-chlordane (Figs. 55.4 and 55.5), each pair having two similarly oriented electronegative atoms (i.e., oxygen and chlorine), with toxicity greater in the first (exo,exo) orientation than in the second (exo,endo) orientation of these substituents. Delta-chlordane (Table 55.3) with l-exo,3-exo chlorine substituents is a highly insecticidal symmetrical variant of the orientation found in the HE 160lcis-chlordane pair. Interestingly, delta-chlordane is an analog of alodan in which the two side-chain chlorines have become fixed in the exopositions by the extra carbon atom of the cyclopentane ring and their insect toxicities are of the same order when alodan is synergized by sesoxane (Table 55.3). The second (3-exo) chlorine in delta-chlordane has a severe effect on mammalian toxicity, because this molecule has a much higher rodent toxicity than either alodan, alpha- (1, Table 55.4) or beta-dihydroheptachlor(2) (DH), or the chlordane isomers (4, 5). As noted already, the order of housefly toxicity of the DH-isomers is beta-DH > alpha-
1142
CHAPTER 55
Interactions with the gamma-Aminobutyric Acid A-Receptor: Polychlorocycloalkanes and Recent Congeners
DH > gamma-DH; beta-DH has an exo-chlorine and is also symmetrical, alpha-DH has an exo-chlorine but is asymmetrical, whereas gamma-DH has an endo-chlorine, which, being "hidden" beneath the ring system, may be less accessible to a critical binding site and also leaves the exo-face of the ring more exposed to metabolic attack from the exo-side (compare the metabolism of endrin in Section 55.2.2). Soloway presented insect toxicity data for many derivatives of HCNB of the chlordane, isobenzan, endosulfan, aldrin, and isodrin series, together with lindane, which already appeared to have the same mode of action (Busvine, 1964). He concluded that high insecticidal activity required the presence of two electronegative centers within a narrow range of distance and direction with respect to one another and placed on or across the plane of symmetry defined by the CClz-bridge. Many cyclodienes fulfill these requirements but some, such as dihydroaldrin, dihydroisodrin, and photoisodrin (9, 10, and 11, respectively, Fig. 55.1) (only one electronegative center) and bromodan (20, Fig. 55.7), alpha-DH, exo-chlordene epoxide (Fig. 55.4), and HCE (8, Fig. 55.1) (asymmetrical), do not, yet are indicated to have high intrinsic toxicities when their metabolism is suppressed in vivo. Evidently, the involvement of a second electronegative center such as an epoxide ring in binding to the site of action may increase the affinity of the molecule for this site, by hydrogen bonding, for example. Thus, lack of a second electronegative center may explain the earlier noted slow action of the dihydro-compounds, which, in the absence of an inhibitor of metabolic oxidations, may afford them increased opportunity for both detoxication and binding to inert storage sites. Notably, the cage molecule photoisodrin is more polar than the related dihydroisodrin and dihydroaldrin and acts rapidly, especially when synergized; it may have the more favorable pharmacokinetic properties of the more rapidly acting epoxides, although apparently lacking their second electronegative center (Brooks, 1973; Brooks and Harrison, 1963). There is limited information about the relative toxicities of enantiomeric forms of chiral cyclodienes, which are obviously of interest in this context. The epoxide hydrolases of pig liver microsomes selectively hydrate the same enantiomers of chlordene epoxide, HCE and HE90 (Brooks et aI., 1968). The isolated residual epoxides appeared to have the same order of toxicity to houseflies as their respective racemates, which are not detectably hydrated by this insect (Brooks et aI., 1970). Miyazaki et al. (1978, 1979, 1980) synthesized the pure enantiomers of chlordene, chlordene exo-epoxide, HEl60, 2-chloroheptachlor (Fig. 55.8), and 3-chloroheptachlor and found that their toxicities to the German cockroach (topical LD50, !-1g/g) were in the order (+ )-chlordene (148) > racemic chlordene (>300) > (-)-chlordene (inactive); (-)chlordene epoxide (74) > racemic chlordene epoxide (158) > (+ )-epoxide (inactive); racemic heptachlor (2.64) > (+)heptachlor (3.38) > (- )-heptachlor (5.32); (+ )-HE160 (1.29) > racemate (1.82) > (- )-HE160 (2.98); (+ )-2-chloroheptachlor (20) > racemate (50) > (- )-2-chloroheptachlor (100); 3-chloroheptachlor (enantiomers and racemate inactive).
Miyazaki et al. concluded that (- )-chlordene is intrinsically nontoxic to the German cockroach, observing that the corresponding (+ )-epoxide (nontoxic) formed in vivo is metabolized to the expected oxidative and hydrolytic products (Brooks and Harrison, 1965) at about the same rate as the toxic ( - )-epoxide from observably toxic (+ )-chlordene. However, they also considered (+ )-chlordene to be intrinsically inactive, therefore requiring bioactivation by conversion into the toxic (- )-epoxide in vivo. Unfortunately, these experiments did not include a synergist to suppress oxidative metabolism. Experiments with houseflies showed that both chlordene and dihydrochlordene had low but measurable toxicities to that insect, which were synergized by sesoxane (Brooks, 1966; Brooks and Harrison, 1964a); in fact, synergized chlordene was only fivefold less toxic than synergized chlordene exo-epoxide. The role of epoxidation in the toxicities of the heptachlor enantiomers (or those of 2-chloroheptachlor) has not been reported. The (+)- to (- )-heptachlor toxicity ratio for German cockroach was 1.56; for (+)- to (-)-HE160 it was 2.3 and for (+)- to (-)-2chloroheptachlor it was 5.0, with the more toxic (+ )-antipodes
_-..~o ~ (+~
~ .~. C1 6
(-)-,
Chlordene
Chlordene
~o '
C1 6
~
~ .-
(+)-
Heptachlor
~-epoxide
Cl
,
(-)-
C16
2
Cl
~
Heptachlor exo-epoxide (HE 160)
2-Chloroheptachlor
Figure 55.8 Absolute stereochemical configuration of the enantiomers of chlordene, chlordene epoxide, heptachlor, heptachlor epoxide (HEI60), and 2-chloroheptachlor as established by Miyazaki et al. (I978, 1979, 1980).
55.3 Structure-Toxicity Relationship and Mode of Action and toxic (- )-chlordene epoxide all having the same absolute stereochemistry (Fig. 55.8). Apart from (+ )-chlordene epoxide and the antipodes of 3-chloroheptachlor, the other antipodes are clearly all active but the ratio of 2.3 for the HEl60 antipodes is likely to be the safest measure of comparative intrinsic toxicities in this series because the known stability of this epoxide should avoid or minimize the complication of metabolism in vivo. Thus, although one absolute configuration of HEl60 is favored, both are toxic, which might be expected if the critical binding site is in a symmetrical (or nearly symmetrical) cylinder of about the same diameter as the molecules discussed, so that either antipode can interact reasonably well with such a site in the bore of the structure, now known to be the chloride ionophore of the GABAA -receptor (Section 55.4). Notably, alpha-DH must exist in enantiomeric forms, which, if superimposed, give a symmetrical "composite" molecule that resembles both delta-chlordane and isobenzan (its oxygen isostere). Likewise, superimposition of the enantiomers of both HEl60 and HE90 gives "composites" that are similar to both delta-chlordane and isobenzan. Such symmetrical molecules might be expected to interact particularly well with a closefitting cylindrical binding site. 55.3.2 COMPOUNDS WITH FEWER, OR NO CHLORINE ATOMS 55.3.2.1 Reductive Dechlorination of Cyclodienes Early information (Soloway, 1965) indicated that the unchlorinated methano-bridge of aldrin could be replaced by 9-syn-CI-CH- and that of isodrin by -CH2CH2- or spirocyclopropane, but the overall molecular length could not exceed that delineated by dieldrin or alpha-endosulfan. There were, however, interesting indications that some of the chlorine atoms in the hexachloronorbomene moiety could be replaced by hydrogen (Table 55.4). Species differences were evident; an aldrin analog (7, Table 55.4) having only the two one- and four-bridge chlorines was reported to be nearly as toxic as dieldrin to the German cockroach, although nontoxic to other insects tested. In aldrin, the methano-bridge chlorine atom anti to the chlorinated double bond was found to be more important for toxicity than the syn-chlorine (compare 8 and 9, Table 55.4), and replacement of the two ethylenic chlorines in dieldrin by hydrogen to give didechloro-dieldrin (DD; 6, Table 55.4) increased housefly toxicity four-fold and toxicity to the German cockroach eightfold; the latter insect appears to be particularly sensitive to these compounds. The high toxicity of diazaaldrin and its N-oxide (3 and 4, Table 55.4) should also be noted. Of interest was the possibility that if the increase in toxicity effected by replacement of the ethylenic chlorines in dieldrin proved to be a general phenomenon for cyclodiene insecticides, it might be possible to combine the change to a tetrachloronorbomene moiety with a more labile epoxide ring or other labile system (e.g., cyclic sulfite as in endosulfan) to produce useful insecticides having both oxidative and hydrolytic
1143
detoxication routes that would be more selective and environmentally acceptable. Selective dechlorination of several series of cyclodienes was undertaken to test this possibility (Brooks, 1975,1977,1980,1985; Brooks and Mace, 1987; Brooks et aI., 1981). It transpired that the effect of dechlorination was not uniform but depended on the structure of the molecule as a whole. The most consistent observation for all series was the greater importance of the anti- versus the syn-chlorine atom in the pentachloronorbomene moiety, as noted for aldrin by Soloway (1965). These two changes for dieldrin were combined to give the 1,4, anti-l O-trichloro-analog of dieldrin (DSD; 24, Fig. 55.7), which approached dieldrin in toxicity to houseflies and blowflies and was 7-20-fold more toxic than its 10syn-chloro-isomer with the chlorine atom adjacent to the double bond. Thus, the pentagonal arrangement of chlorine atoms evident in lindane and in the cyclodienes derived from HCNB (Busvine, 1964) is not sacrosanct for cyclodienes. 55.3.2.2 Structural Convergence of Cyclodienes and Their Dechlorinated Analog with Other Cage Convulsants Acting at the Chloride Ionophore Further information arose from comparisons with the naturally occurring convulsant picrotoxinin (PTX; Fig. 55.9), which Hathway et al. (1965) found to have effects similar to those of dieldrin and isobenzan on ammonia metabolism in rat brain. PTX antagonizes the action of GABA by blocking the chloride ion channel associated with its receptor (Kadous et aI., 1983; Takeuchi and Takeuchi, 1966, 1972). Evidence was then reported that cyclodienes and lindane compete with PTX at a commmon binding site in cockroach brain (Matsumura and Ghiasuddin, 1983; Tanaka et aI., 1984), and this was the site of convulsant action of these compounds, a proposal supported by the similarity in the neurophysiological effects of of the cyclodienes and PTX and the cross-resistance to PTX shown by cyclodiene-resistant cockroaches. The structural similarities among PTX, HE160, and lindane led Ozoe and Matsumura (1986) to elaborate the two electronegative center hypothesis of Soloway (1965) in a series of PTX analogs that emphasized the importance of the bulky trans-substituent on the lactone ring as a third requirement for interaction with the PTX binding site. This is the point of convergence (Fig. 55.9) with the cyclodienes (B) and lindane (C), for which the lO-anti-chlorine and the central axial chlorine (or the central equatorial chlorine), respectively, appear to provide the appropriate bulky substituent (Brooks and Mace, 1987) and the highly toxic cage compounds such as t-butylbicyclophosphorothionate (TBPS) and t-butylbicycloorthobenzoate(TBOB), in which the t-butylmoiety is the necessary bulky substituent (Casida et aI., 1985; Palmer and Casida, 1985). The highly insecticidal orthobenzoate EBOB (Fig. 55.10) proved to be the best ligand for the insect GABA-receptor chloride ionophore, and eH]-EBOB has been used subsequently in binding displacement studies with numerous putative channel blockers at the PTX binding site. Attempts to simplify the orthobenzoate ring system of these potent
1144
CHAPTER 55
Interactions with the garnma-Aminobutyric Acid A-Receptor: Polychlorocycloalkanes and Recent Congeners A (trans-J CB
a Cl
A
3
1
Cl S
cl
e
c Ya 8 ' 'Cl
~lc~_",...~.!
~
B
a
Cl
e
Cl
A
PTX
B Cyclodienes
antl-lO-CI
C12
C
Lindane
trans-lsopropenyl
Cl3
D
E
Figure 55.9 Structures of (A) picrotoxinin (PTX) with its bulky beta-isopropenyl group trans- (or anti-) to the lactone ring; (B) heptachlor epoxide; (C) lindane, showing the aaaeee configuration of chlorine atoms essential for toxicity. A, anti-; S, syn-; a, axial; e, equatorial substituents. D shows the highly chlorinated face of a cycIodiene insecticide, as in heptachlor epoxide (B above), and E, the lactone ring system as in PTX, both viewed from the right ("end-on" position). In the corresponding view of lindane (C viewed from below), note that the chlorine equivalent to the syn-IO-chlorine of fully chlorinated cycIodienes is replaced by a hydrogen (double arrow). However, the electrostatically more favored superimposition of lindane on PTX (Calder et aI., 1993) is C viewed from the right, in which the three-axial chlorines of lindane are together equivalent to the trans-bulky isopropenyl substituent and lactone ring of PTX.
cage convulsants and the search for structural changes to confer selective toxicity in favor of mammals versus insects produced numerous insecticidal dithianes, oxathianes, and their sulfoxides and sulfones (Palmer and Casida, 1995; Wacher et aI., 1992). Figure 55.10 further shows the convergence of structural changes between partly (Brooks and Mace, 1987) and totally (Ozoe et aI., 1990) dechlorinated alpha-endosulfan (1, Fig. 55.10) (which surprisingly retains measurable housefly toxicity (LD50, ca. 43 J.l.g/g; 20-fold less toxic than alphaendosulfan, when co-applied with sesoxane), the toxic t-butyl trioxabicyclooctanes (TBOs) such as 2 (Fig. 55.10) (Palmer and Casida, 1985), and the dithianes described by Elliott et al. (1990). Although the 1,3-dioxan (not shown but analogous to structure 7) corresponding to 2 (Fig. 55.10) is weakly insecticidal (Palmer and Casida, 1995), the hybrid molecule (3) of deschloro-alpha-endosulfan (1) and that 1,3-dioxan was more toxic than fully dechlorinated alpha-endosulfan (1), indicating that extra rigidity conferred on the 1,3-dioxan structure can improve insecticidal activity. In the hybrid molecule (3), the norbomene moiety appears to provide the bulky substituent (compare t-butyl in 2), while possibly increasing the conformational rigidity of the dioxan moiety. Furthermore, Ozoe et al. (1993) showed that the extended structure 4
(Fig. 55.10) is highly insecticidal and that compound 5 has intermediate toxicity, which is entirely lost by any chlorination in the norbomene nucleus. Thus, the dioxan (5) loses its housefly toxicity (LD50, 5.5 J.l.g/g) completely when oneor two-bridge chlorines are introduced and there is evidently a crossover point between the two structural types, because the fully chlorinated but not extended dioxan (6) has the same toxicity, which is much reduced in various partially dechlorinated analogs (see Section 55.4.3 for further discussion). Bridge bis-chlorination of fully dechlorinated alpha-endosulfan (Fig. 55.10) greatly reduces its housefly toxicity (Ozoe et aI., 1993), but the 10-anti-chlorine analog retains measurable toxicity and several dechlorinated analogs having this Cl atom in combination with one ethylenic chlorine and the two bridgeend chlorines are highly toxic, showing that the additional chlorines are required for binding this shortened molecule in the critical site (Brooks and Mace, 1987), implying the requirement for a minimum of four chlorines for high toxicity. Nevertheless, at least one of the bridge-end chlorines in cyclodienes can be replaced, because a dieldrin analog (25, Fig. 55.11), which has one bridge-end carbon atom replaced by nitrogen, has appreciable insect toxicity (Gladstone and Wong, 1977).
55.3 Structure-Toxicity Relationship and Mode of Action
1145
anti-
H~CL 4H 0,
H~O, syn-
'V
O'S
~
0-
~
anti-lO, pentachloroisomer (1.0)
arious dechlor endosulfans
~
~9
0-
CL Cl ~O"
P=S TBPS
0 .... ~
0
4H
~
6H~-
, rE0'
S 1\
0
(265)
----.'
(>500)
6H
~~R R=
(1) deschloro-endosulfan (43)
(3) 4-Br-Ph (15)
R= phenyl (Ph) (TBOB) (2) R= 4-Br-Ph (0.8) 4-CHaC-Ph various TBOs 4-CN-Ph
(4) 4-CII;C-Ph (0.33) (5) 4-CN-Ph (5.5)
~ o
6CI
~
(6)
(5.5)
>
U
f:I
~~t'"-·-~ (8)
f,
H axial
~)2
Ph-4-C",CH
(0.24)
Figure 55.10 Convergence between exploration of reductively dechlorinated alpha-endosulfan analog (norbornene type) (Brooks and Mace, 1987; Ozoe et aI., 1990, 1993), the trioxabicyclooctane-derived cage convulsants, and the more recent dioxans and dithianes (Pulman et al., 1996). The bracketed number following a chemical number or structure is the housefly topical LDSO (J.1g/g; measured in the presence of sesoxane or piperonyl butoxide).
Many trioxabicyclooctanes are highly toxic to both mammals and insects but remarkable selectivity can be conferred on some structures by appropriate derivatization; thus, the trimethylsilyl-derivative (26, Fig. 55.11) of the 4-n-butyl analog of EBOB (Fig. 55.10) is highly toxic to houseflies (LDso, 0.43 Ilg/g) but poorly toxic to mice (LDso, > 400 mg/kg). This derivative appears to be oxidatively reconverted into its toxic ethynyl precursor in houseflies, whereas in the mouse this oxidative bioactivation is much less important (Palmer et aI., 1991b). Because all the compounds mentioned and some other classes of cage convulsants are believed to act at the PTX binding site, a considerable array of compact molecules is now available with which to delineate this site. A 4-ethynylphenylsubstituent in the 2-position of 5-t-butyl-l,3-dioxane (7, Fig. 55.10) or 1,3-dithiane (8, Fig. 55.10) was found to be more effective than a 4-bromophenyl-substituent; the trans-(linear) ethynylphenyl dithiane (8) was somewhat more toxic to houseflies than the analogous trans-dioxane (7) and cis- (angular) isomers were generally equally toxic to or less toxic than trans-isomers (Palmer and Casida, 1995; Pulman et aI., 1996). With the possibility of oxidation at sulfur in vivo, which may enforce additional conformational rigidity and also increase binding propensity, the situation becomes more complex (see Section 55.4.3).
55.3.3 LINKS BETWEEN POLYCHLOROCYCLOALKANEAND RECENT HETEROCYCLICS APPARENTLY ACTING AT THE CHLORIDE IONOPHORE
Recently, arylpyrazoles, such as fipronil (27, Fig. 55.11), and various 5-alkyl-2-arylpyrimidines (28) and 1,3-thiazines (29) (Pulman et aI., 1996), in which the planar heterocyclic ring replaces the spacers formed by the TBO and 1,3-dioxane and dithiane structures, have been added to the list of chloride ionophore blockers. Insecticidal activity was also found in triazoles (30, 31) (Boddy et aI., 1996; Von Keyserlingk and Willis, 1992) and pyrimidinones (32) (Whittle et aI., 1995) and a spirosultam (33) (Bloomquist et aI., 1993), demonstrating the diversity of structures that probably act at this site. Cole et al. (1994) examined the inhibition of eH]-alphaendosulfan binding in housefly head membranes by lindane and several cyclodienes and concluded that these insecticides are the only GABA-receptor ionophore blockers that consistently inhibit the binding in these membranes not only of the earlier used ligands such as eSS]-TBPS and eH]-EBOB but also of eH]-alpha-endosulfan. However, a representative dithiane, EBOB, fipronil, and other pyrazoles were less effective in inhibiting [3H]-alpha-endosulfan binding than the chlorinated insecticides, from which it appeared that the latter compete
1146
Cl
CHAPTER 55
Interactions with the gamma-Aminobutyric Acid A-Receptor: Polychlorocycloalkanes and Recent Congeners
~ )~ ~
r'
~"'PH-4-C=CSi (CH3)3
0
(26)
c.l
t
'-.FN ~N ~NQ- PH-4-C=CH ~ (28)
PH-4-C=CH (29)
Cl Figure 55.11 35).
Chemical structures of compounds mentioned in the text (25-
directly for the endosulfan site, whereas the others bind with different inhibition kinetics or at a site more closely coupled to the EBOB than to the endosulfan binding domain. Notably, the channel activator avermectin Ba did not inhibit endosulfan binding. An even more suitable ligand for the chlorinated insecticides is [3H]-BIDN (34, Fig. 55.12) (Holyoke et aI., 1994), a simple norbomene derivative, which has high insect and mammalian toxicity (KOlbl et aI., 1981; Middleton and Bingham, 1982). Several putative affinity probes for the binding site have also been described (Casida and Pulman, 1994). When aryl pyrazoles synthesized as herbicides were found to be insecticidal, their convulsive activity was not immediately recognised to result from GABA-antagonism (Klis et aI., 1991). Co le et al. (1993) reported subsequently that several compounds, including fipronil (27, Fig. 55.11) (Colliot et aI., 1992; Hatton et aI., 1988), which has become a commercially successful insecticide, blocked the GABA-gated chloride ionophore with higher potency for a site in housefly than in mouse brain, offering the possibility of selective toxicity. Fipronil has relatively low acute mammalian toxicity (Section 55.4.3). It inhibits eH]-EBOB binding to housefly head membranes and dieldrin-resistant flies show some resistance to it (Cole et al.,
1993; Colliot et aI., 1992), providing a clue to its mode of action. The cyclodiene insecticides and lindane were found to be potent displacers of eSS]-TBPS binding to GABA-receptors in rat brain and inhibitors of GABA-dependent 36 Cl ion flux into rat brain microsacs, from which it was suggested that these PCCAs act as noncompetitive blockers of GABAAreceptors (Abalis et aI., 1985; Gant et aI., 1987; Lawrence and Casida, 1984). Potency in these assays correlates with toxicity (Casida et aI., 1988) but TBPS is not a potent insecticide and [3S S]_TBPS is unsuitable as a radioligand for insect studies; it appears that the structural features required for binding at the housefly GABA-receptor are different from those for the mammalian one and eH]-EBOB, a highly potent insecticide, was ultimately designed as a superior ligand for insect binding studies (Deng et aI., 1991) and generally provides a good correlation between its displacement by PCCAs and their housefly toxicities. By use of this ligand, it was concluded that PCCA, PTX, dithiane-related compounds, and phenylpyrazoles all have the same mode of insecticidal action, a view supported by the up to 27-fold cross-resistance to EBOB shown by dieldrin-resistant houseflies (Cole et aI., 1993). Moreover, the naturally occurring insecticide avermectin Bla and derived moxidectin (Fisher, 1997), which behave as GABA-agonists, stimulating rather than inhibiting chloride ion influx, are potent noncompetitive inhibitors of EBOB binding. This implies that averrnectin action involves the chloride ionophore but that it is bound at a site different from that involving EBOB and PCCA; nor, in contrast to EBOB, is there cross-resistance to dieldrin, so that the channel modification that confers dieldrin resistance does not apparently involve the avermectin binding site (Deng et aI., 1991). Based on ligand binding studies, Deng et al. (1993) proposed four partly associated sites in the housefly GABA chloride ionophore that are relevant to insecticidal action: site A, interacting with EBOB and its isosteres; B with TBPS and isosteres; C with phenylpyrazo1es; and D with averrnectins. Action at sites A and C gives similar signs of poisoning and crossresistance to dieldrin; PCCA and some TBPS isosteres may act at both A and B. The avermectin site D is coupled in some way with A and C but not to the TBPS site B, which is also distinct from the phenylpyrazole site C. Thus, the reduced affinity for [3H]-EBOB binding observed in dieldrinresistant houseflies is due to its reduced affinity for the PCCA binding site, and the cross-resistance noted for TBOs, lindane, toxaphene, cyclodienes, dithianes, arylsilatranes (35, Fig. 55.11), and PTX suggests that the structural modifications in the EBOB binding site are involved in resistance to all these insecticides (Hawkinson and Casida, 1993) but fortunately do not confer resistance to averrnectins, which have very high toxicity against agricultural and household insect pests, phytophagous mites, and plant and animal nematodes.
55.4 Molecular Mechanism of Action
1147
Chloride ion channel
OUTSIDE
:,--,
.
,"" • • 1
"I
,- -~
,I :
I
LIPID BILAYER
,.. - - ' l
, I
I
-, I
CYTOPLASM
-- -
-
(INSIDE)
binding site for noncompetitive blockers (cyclodienes, lindane, PTX, arylpyrazoles, etc.)
Figure 55.12 Schematic representation of the GABAA -receptor of mammalian brain, showing five transmembrane glycoprotein subunits, each with their four trans-membrane helices (M \-M4), of which the M2 segments (shown as cylinders, and black circles in the plan view) are believed to form the pore of the integral chloride ion channel (MacDonald and Olsen, 1994). A modified subunit carrying cyclodiene resistance in Drosophila (Rdl) shows homology with the mammalian brain beta-subunit (ffrench-Constant et ai., 1991).
55.4 MOLECULAR MECHANISM OF ACTION 55.4.1 TOPOGRAPHY OF THE GAMMA-AMINOBUTYRIC ACID A-RECEPTOR GABA is the principal neurotransmitter of the mammalian and insect central nervous system (CNS) and the insect neuromuscular junction. In mammals, baclofen-sensitive GABABreceptors are coupled to calcium and potassium channels and the action of GABA is mediated by G-proteins. In contrast, GABAA -receptors, of interest here, are members of the super family of ligand-gated ion channels that contain a chloride ionophore (Schofield et aI., 1987). Simplistically, an inhibitory GABA-ergic nerve terminal abutting on the presynaptic terminal of another nerve that releases a neurotransmitter (e.g., acetylcholine, ACh) releases GABA when stimulated. GABA then diffuses to the presynaptic terminal of the other nerve, where it binds to a GABAA -receptor, causing entry of chloride ions and resulting in hyperpolarization of the terminal and inhibition of release of the other neurotransmitter. Thus, postsynaptic stimulation of the other nerve by its transmitter (e.g., ACh) is reduced. This inhibitory mechanism explains the apparent cholinergic effects of dieldrin and lindane on Ameri-
can cockroach ganglia (Shankland and Schroeder, 1973; Uchida et aI., 1978), because disinhibition (blockade of a presynaptic chloride ionophore) of the presynaptic terminal of a cholinergic nerve should result in uninhibited ACh release and consequent hyperstimulation of the postsynaptic terminal, as is observed. The same basic mechanism for disinhibition may, of course, affect nerve terminals involving neurotransmitters other than ACh (Joy, 1982) with a variety of possible effects, depending on the species and on differing nerve architecture. The GABAA -receptor of human brain consists of four or five 50-60 kDa glycoprotein subunits, each of which contains four (M 1-M4) hydrophobic domains (alpha-helices) that traverse the membrane (Fig. 55.12) (MacDonald and Olsen, 1994; Schofield et aI., 1987) and contribute to and stabilize the walls of the chloride ionophore. The five M2-domains are believed to be arranged so as to form the 5.6-A-diameter lumen of the channel, with the side chains of their threonines and serines forming hydrophilic rings that contribute to the induction of ion flow.
55.4.2 MOLECULAR BIOLOGY OF CYCLODIENE RESISTANCE Recently, a cyclodiene resistance-conferring gene, Rdl, from the fruit fly, Drosophila melanogaster, has been cloned and shows homology with the mammalian brain beta-subunit
1148
CHAPTER 55
Interactions with the gamma-Aminobutyric Acid A-Receptor: PolychlorocycIoalkanes and Recent Congeners
(ffrench-Constant et al., 1991). Dieldrin resistance was subsequently found to be associated with the point mutation alanine 302 to serine (Ala302-Ser) within the M2 membranespanning domain, near the site conferring charge selectivity in the closely related nicotinic ACh-receptor (ffrench-Constant et al., 1993a, b). Homooligomeric, wild-type Rdl-receptors expressed in Xenopus oocytes showed the expected electrophysio10gica1 properties of GABA-receptors; channels containing the Ala302 _ Ser mutation, when similarly expressed, were identical with the wild-type ones but had consistently lower sensitivity to dieldrin and PTX, sensitivity to these being reduced about lOO-fold. Notably, Lee et al. (1995) reported no detectable specific [3H]-EBOB binding in resistant D. melanogaster strains carrying the Ala302 -rep1acement, despite high specific binding to membranes from susceptible flies, indicating the involvement of Ala 302 in EBOB binding to GABA-receptors containing Rdl subunits. Similar results were found for the dieldrin resistance mutation Ala302 -glycine found in Drosophila simulans (ffrench-Constant et al., 1993a, b). Furthermore, EBOB blocks chloride ion currents generated by Rdl homomultimers expressed in insect cells and the Ala302 -Ser replacement reduces sensitivity to this block lO-fold (Lee et al., 1995). Examination of the Rdl gene from three different insect orders has revealed that in all cases Ala302 is replaced by either a serine or a glycine (less effectively), indicating that this mechanism is universal. The change confers reduced sensitivity to PTX, lindane, and TBPS, lower channel conductances, extended channel open times and shorter closed times, and a markedly reduced rate of GABA-induced receptor desensitization. From a simple model to represent binding and allosteric changes, it has been suggested that the preceding mutations are the only ones that can directly weaken cyclodiene binding to the desensitized (antagonist favored) conformation of the receptor and simultaneously destabilize the antagonist-favored conformation through an allosteric mechanism, resulting in a powerful dual-resistance mechanism (ffrench-Constant et al., 1995; Zhang et al., 1994). In this model, the antagonist associates with the open channel but binds much more tightly when the channel next changes into the desensitized (closed) state, so that this configuration is stabilized. If homomultimers are present in vivo, which may not, however, be the case (Zhang et al., 1995), the preceding mutation could lead to a resistant ion channel with a ring of up to five serines replacing the five alanines in the wild-type Rdl ion channel, which would greatly increase the polarity of this region (5-CH20H replacing 5-CH3) and considerably alter its affinity for the various toxicants under discussion; even one or two added hydroxyl groups introduced here in a heteromultimer might have a significant effect and also reduce the energy barriers for ion permeation by participating in hydrogen bonding with water (Leonard et al., 1988). Analogies with the nACh receptor are evident and the closed configuration of the mutated chloride channel may remain somewhat ion permeable (Zhang et al., 1994; Revah et al., 1991).
55.4.3 MOLECULAR TOXICOLOGY OF NONCOMPETITIVE CHLORIDE IONOPHORE BLOCKERS
The localization of Ala302 to the PTXlcyclodiene binding site in Rdl prompts some further consideration of the information on the structure-toxicity relationships outlined in earlier sections. The cylinder formed by the amino-acid sequences Leu 303 , Ala302 , Va1 301 in five adjacent M2 domains (helices) provides a quasi-centrosymmetricallipophilic pocket into which cyclodiene insecticides and other noncompetitive chloride ionophore blockers (NCBs) might fit. The molecular dimensions of cyclodienes, taking into account the known range of allowable molecular substituents in these rather compact molecules, are sufficient to block a 5.6-A pore. If the pore is more or less centrosymmetrical, a cyclodiene molecule could fill the pocket and interact with all of its walls simultaneously because similar binding sites are presented around the lumen even if the arrangement is not a homomultimer. This would explain the observed toxicity of both enantiomeric forms of asymmetric molecules such as heptachlor epoxide; the forms may differ somewhat in toxicity, however, if the channel is not completely symmetrical, as is found (Miyazaki et al., 1980). Symmetrical molecules should be particularly effective, because they may be able to offer a symmetrically distributed electronegative center (or centers) to similar binding sites on opposite sides of the channel, as in the case of delta-chlordane and isobenzan, each having two symmetrically substituted chlorines on their fivemembered rings. These molecules may be viewed as symmetrical composites of the enantiomeric forms of alpha-DH and heptachlor epoxide (HE160), respectively, as suggested in Section 55.3.1 In a hypothetical model (Fig. 55.13) in which the HCNB moiety of cyclodienes is presumed bound at the synaptic end of the lipophilic pocket so that its gem-dichloro-bridge is presented to the channel wall, then the second electronegative center in, for example, dieldrin, is directed toward the cytoplasmic end of the pocket with its epoxide ring and unchlorinated methano-bridge fixed in an inward direction toward the channel lumen. This "cytoplasmic" end of the molecule is then close to the critical ring of Ala302 -methyl groups around the channellumen, which when replaced by -CH20H groups inhibits the binding of NCBs. Dieldrin and alpha-endosulfan (extended molecules) appear to provide the limiting acceptable molecular "lengths," as noted earlier, whereas isobenzan, endrin, and betaendosulfan are more compact. On this model, it might be argued also that the anti-1 O-chlorine atom (Fig. 55.9 and Table 55.4) of the dichloromethano-bridge of cyclodienes is better accommodated in the lipophilic pocket than the syn-10-chlorine, which may interfere sterically with the large side-chain alkyl groups of the ring of Leu 303 s that lie at the synaptic end of the pocket, making this syn-chlorine universally unfavorable for toxicity. The same argument might explain the four-fold increase in dieldrin toxicity effected by removal of the ethylenic chlorines (in DD, Table 55.4), because these chlorines might also interfere sterically in this region. This increase in toxicity is not uni-
55.4 Molecular Mechanism of Action
,
CYTOPLASI\
-=-.
I I:
- 5-'
-
\ §..=
-'
•
\ ... 5
",
,,= H 1,\-
.........:;.--H,c ,-
SYNAPSF
Figure 55.13 Dieldrin (A) is oriented in a hypothetical binding site in or near the chloride ion channel lumen (in the region of Leu 303 ?) with chlorine X (I O-anti-chlorine) located in a subsite P that accommodates a bulky substituent. Its epoxide ring then penetrates a three-dimensional region S near Ala302 , which may interact, especially in the closed channel configuration, with the electronegative moieties of various cyclodienes when similarly oriented. If, however, Y (the IO-syn-chlorine) is presented to P, then the epoxide ring cannot so readily interact with zone S (dieldrin orientation B). C indicates the approximate position of the epoxide ring of endrin and also of the sulfur of beta-endosulfan, when either X or Y in these molecules is bound to subsite P; D is the approximate position of the endrin methano-bridge when its IO-antichlorine is located in P, corresponding to dieldrin orientation A. In this model, the dotted cylinder is the region containing a ring of leucine side chains. Note that a substituted benzene ring and some other extensions are permissible in the arrowed directions when the bridging system is unchlorinated (see Fig. 55. IO and related discussion, Section 55.4.3).
versal for cyclodienes, however, and the syn-chlorine atom is still present in DD, so that its adverse effect on toxicity is more than offset by removal of the ethylenic chlorine atoms. Further reductive replacement of the syn-10-chlorine atom in DD to give the trichloro-derivative DSD (24, Fig. 55.7) reduces toxicity to the level of dieldrin again (Brooks and Mace, 1987). The syn-lO-chloro- isomer of DSD is significantly less toxic than DSD or dieldrin, again indicating the greater importance of the 10-anti-chlorine for toxicity. The difference in toxicities between syn- and anti-l O-monodechloro-isomers is less marked for endrin and beta-endosulfan (the endrin-like isomer) but remains evident for alpha-endosulfan. This observation was discussed (Brooks, 1992) in connection with the insect cross-resistance spectrum for lindane/cyclodienes first noted by Busvine (1964), in which lindane, isobenzan, endrin, and the endosulfan isomers retain measurable toxicity to dieldrin-resistant insects (Brooks and Harrison, 1964a; Busvine, 1964). The first three molecules and beta-endosulfan are rather compact compared with dieldrin and alpha-endosulfan; the latter has been considered to be extended
1149
and dieldrin-like (Fig. 55.6), but recent structural studies indicate a more complex situation (see later). If the 10-anti-chlorine of dieldrin (X in Fig. 55.13) corresponds to the bulky anti-substituent found in PTX and must be presented to an appropriate lipophilic pocket in a binding subsite (P, Fig. 55.13) so as to place the epoxide ring in a correct position (in the region of S) with respect to the remainder of the binding site, then the lO-syn-chlorine (Y in Fig. 55.13), if similarly presented, cannot place the epoxide ring in the same position. If this latter position is modified in resistance to prevent interaction with the epoxide ring, dieldrin can no longer bind; however, either bridge-chlorine of endrin or beta-endosulfan can be offered to the bulky substituent binding subsite (P) such that the epoxide ring or sulfite moiety will still be placed in approximately their original positions (near to C), still able to interact with the critical site (S), on account of the more compact "cage" shape of these molecules. Consequently, these molecules may still be able to interact to some extent with the binding site that has been modified to exclude binding with dieldrin. On this model, alpha-endosulfan is anomalous in retaining effectiveness, because the same arguments apply as to dieldrin yet this molecule was actually somewhat more toxic than beta-endosulfan to dieldrin-resistant houseflies (Brooks and Harrison, 1964a). Two further observations may be significant, however. First, endosulfan sulfate (4, Fig. 55.6) formed in vivo from both endosulfan isomers may be the critical toxicant; it has the same structural (cage) configuration as beta-endosulfan (Forman et aI., 1965) and is formed faster from alpha- than from beta-endosulfan in some living organisms. Second, alpha-endosulfan has recently been reported (Schmidt et aI., 1997) to exist in the asymmetrical "twist-chair" conformation, in which the c-o bonds are "trans," not parallel as usually depicted (Fig. 55.6). Molecular models suggest that this twisted configuration may be more flexible, allowing the S=O moiety to occupy several spatial positions between the extremes represented by beta-endosulfan and the extended alpha-structure shown in Fig. 55.6. Consequently, the alphaisomer might be expected to be intrinsically at least as effective as the beta-isomer in terms of their interactions with the resistance-modified binding site, regardless of possible oxidation to the sulfate in vivo. In the case of endrin, the 9-keto(l2-keto-) metabolite is presumed to be the ultimate toxicant in mammals (Hutson et aI., 1975) and may contribute to endrin toxicity in insects (Kadous and Matsumura, 1982); notably, this oxidation places a second, additional, electronegative center at D, near to the upper sub site S (Fig. 55.13), which may improve binding potency toward the resistance-modified binding site. Lindane resembles a very compact cyclodiene and might bind without conflict with a subsite modified for dieldrin resistance or in more than one orientation, and similar arguments apply to isobenzan. Interestingly, isobenzan may be regarded as a "composite" of the HEl60 enantiomers, in which an "in plane" oxygen replaces the epoxide rings. Dieldrin resistance normally confers total resistance to HE160 (Brooks and Harrison, 1964a; Busvine, 1964) so the more compact placement of oxygen in isobenzan, combined with a possible increase in binding affinity
1150
CHAPTER 55
Interactions with the gamma-Aminobutyric Acid A-Receptor: Polychlorocycloalkanes and Recent Congeners
associated with the symmetrical chlorine substituents, appears to overcome both dieldrin and HE160 resistance to some extent. Notably, diaza-aldrin (3, Table 55.4), in which the unchlorinated double bond is replaced by -N=N- and which is probably converted into its N-oxide (4, Table 55.4) in vivo, is much more toxic than dieldrin to some insects (Busvine, 1964; Soloway, 1965) but dieldrin-resistant insects are immune to it. Using the TB PS binding assay in rat brain membranes, it has recently been confirmed that this molecule inhibits the binding competitively and therefore interacts directly with the PTX binding site (Ozoe et aI., 1995). Other modifications of the dichloroethylene moiety of dieldrin are acceptable to the binding site, as in photodieldrin (converted in vivo into Klein's ketone, however; Fig. 55.3) and the ketone analogs (14, 15, 16, Fig. 55.7) derived from isobenzan. From the analogy with PTX, the dichloroethylene moiety, unchlorinated double bond, and ketone derivatives in these various analogs may correspond to the lactone system of PTX. In lindane, the best superimposition with PTXlcyclodienes is apparently that in which two of the axial chlorines substitute for the lactone ring of PTX and the third axial chlorine provides the equivalent of the bulky trans-isopropenyl group of PTX or anti-l0-chlorine of cyclodienes. This last configuration of lindane is favored electrostatically (Calder et aI., 1993), although in it the bridge-end chlorines, which complete the pentagonal arrangement of chlorine atoms seen in cyclodienes (Brooks and Mace, 1987), are replaced by hydrogens. Replacement of the bridge-end chlorines by hydrogen reduces the toxicity of some cyclodienes (Soloway, 1965) and, notably, whether bulky alkyl groups in the alpha- and beta-positions of gammabutyrolactones (cf. PTX) stabilize the open or closed states of the ionophore and hence induce anticonvulsant or convulsant activity depends markedly on the stereochemistry of these substituents relative to the carbonyl group (Holland et aI., 1995; Klunk et aI., 1983; Peterson et aI., 1994). However, the replacement of one bridge-end carbon atom of dieldrin by nitrogen (available for hydrogen bonding) gives azadieldrin (25, Fig. 55.11) without great loss in toxicity compared with dieldrin (Gladstone and Wong, 1977), another indication that some modification in this region of cyclodienes is possible. The preceding discussion indicates that numerous modifications of the HCNB moiety retain binding capacity to the critical site. In this context, it may be noted that BIDN (34, Fig. 55.11), in which the simple norbomene nucleus carries two strongly electron-withdrawing substituents (gem-diCN and gem-di-CF3), is highly toxic to both mammals and insects (K61bl et aI., 1981). The allowable length compatible with toxicity of the fully chlorinated cyclodienes appears to be restricted, however (see earlier). TBPS has approximately the same molecular length as cyclodienes such as dieldrin, but there is the question of how the binding of cyclodienes to the critical site relates to that of the extended unchlorinated molecule EBOB or similar molecules in which an aromatic substituent replaces the bulky 4-alkyl group (Palmer et aI., 1991a). Superimposition of cyclodienes, aryl-TBOs, aryldithianes, arylsilatranes (35, Fig. 55.11), and PTX by CoMFA (comparative
molecular field analysis) (Calder et aI., 1993) supports the view that all act at the same or overlapping sites. Interesting additional information is available from work on the molecular hybrids of dechlorinated alpha-endosulfan (Fig. 55.10) (Ozoe et aI., 1993) and insecticidal dioxans (Palmer and Casida, 1995) mentioned previously (Section 55.3.2.2). In this series (Ozoe et aI., 1993), the fully chlorinated molecule (6, Fig. 55.1 0) cannot be extended but has the same housefly toxicity as the unchlorinated extended molecule (5, Fig. 55.10) related to the compounds reported by Casida. The latter molecule cannot be chlorinated; a single bridge chlorine abolishes its housefly toxicity. Assuming that the hybrid molecule (5) is a rigid form of the corresponding dioxan in which the unchlorinated norbomene moiety serves as the bulky substituent (e.g., t -butyl) (Ozoe et aI., 1990) and occupies a spatial region equivalent to that occupied by the HCNB moiety of cyclodienes, it is evident that the combination of substituted phenyl, dioxan, and norbomene moieties can bind to the receptor and afford potentially excellent insect toxicity, as found in 5-t-butyl-2-(4-ethynylphenyl)-1,3-dioxane (7, Fig. 55.10). However, chlorination may disrupt this binding in the corresponding deschoro-alpha-endosulfan analog with a 4-cyanophenyl substituent by forcing the "extended" molecule into a position it cannot occupy in the ion channel for steric reasons. Thus, the compact unchlorinated norbomene moiety may not by itself have sufficient binding potency in the lipophilic pocket bounded by the Leu 301 , Ala302 , and Va1 303 rings, but an added aromatic ring with an appropriate 4-substituent (particularly ethynyl), which, according to this model, binds additionally in a region of the channel pore beyond Ala302 and near to Va1 301 and Arg 300, may greatly reinforce the interaction. Conversely, the bulky, fully chlorinated cyclodiene interacts well with the closed configuration of the spherical pocket, but extension is impossible in this case because the added benzene ring in the extended HCNB moiety would be forced into steric hindrance with the valine isopropyl groups and the narrow cytoplasmic end of the channel; this contains the large guanidiny1 side chains of Arg 300 and might admit only "sticklike" structures such as the ethynyl moiety. Support is given to this hypothesis by the application (Akamatsu et aI., 1997) of CoMFA analysis to compare alphaendosulfan analogs (their series 2) with the hybrid extended norbomene derivatives (series 1) (Ozoe et aI., 1993; Fig. 55.10) analogous to the 1,3-dioxanes reported by Cas id a (Fig. 55.10). When the simple cyclodienes were closely superimposed on the extended molecules with a 2-(4-cyanophenyl) substituent, the housefly toxicity of some of the extended molecules was not well predicted until, in the superposition, the extended molecules were rotated 15° clockwise about their common bond on the norbomene ring junction (C4a-C8a in Table 55.4). This rotation enabled the separate correlation equations for the two series derived by CoMFA on the basis of close superimposition to be combined satisfactorily into a single equation representing the housefly toxicities of both series. The two series compared could not, however, be brought together in the same way when the measure of biological activity was the dis-
55.4 Molecular Mechanism of Action
placement of eSS]-TBPS binding from rat brain membranes. In compound 5 (Fig. 55.10), the previous rotation turns the aromatic ring toward the center of the ion channel as modeled in Fig. 55.13, away from a region sterically forbidden according to CoMFA, and incidentally may "rock" the unchlorinated norbornene moiety into closer contact with the lipophilic pocket than it can achieve when chlorinated. Conversely, chlorination of the norbornene moiety of these extended molecules would have the reverse effect, forcing the aromatic ring toward the channel wall, into sterically forbidden space. The extending group in TBOs need not necessarily be aromatic, because short alkyl chains with a terminal ethynyl group give insecticidal activity, especially when synergized (Smith et aI., 1993) as do 4-ethyny1cyclohexyl groups in TBOs and dithianes (Weston et aI., 1995). A further interesting feature of the dithianes is their conversion into sulfoxides and sulfones, which doubtless occurs in vivo; the equatorial (trans- with respect to t-butyl; linear) 5-t-butyl-2-(4-Br-phenyl)-1,3-dithiane (8, Fig. 55.10; two-substituent is equatorial 4-Br-Ph) is twofold more toxic than the eis- (2-axial; angular) isomer, and the corresponding 4-ethynylphenyl-isomers have equal toxicity when synergized; for each isomer, conversion into the isomeric monosulfoxides and monosulfone progressively increases toxicity (Palmer and Casida, 1992). The not dissimilar toxicities of the trans- (linear) and eis- (angular) dithianes is intriguing. If, in the pore model (Fig. 55.13), the bulky t-butyl substituent is placed in the lipophilic pocket as for the other bulky substituents discussed previously, then the sulfoxides and sulfones from the isomeric dithianes occupy rather different positions in this lipophilic site; only in the linear isomer do these moieties occupy positions near the two ester oxygens of alpha-endosulfan and the epoxide ring of dieldrin, toward the cytoplasmic end of the channel and near to Ala302 . For the angular isomer, the SO (or S02) moieties are placed at the synaptic end of the lipophilic pocket and lie toward the channel center, in this case, the aromatic ring overlies the trioxabicyclooctane ring of TBOs and its 4-substituent reaches only the aromatic ring of the linear isomer, so that the position of this part of the angular molecule is foreshortened in the cytoplasmic direction relative to the linear one. All of the molecules discussed can be superimposed, and with the cyclodiene molecules oriented in this way, the sulfite moiety of beta-endosulfan (and the S02 of endosulfan sulfate) lies in the region occupied by the aromatic ring of the angular dithiane isomer. It should be noted, however, that the possibility of eis- to trans-rearrangement exists for the dithianes in vivo (Pulman et aI., 1996), in which case, the eis-isomers would merely be precursors of the linear (trans-) molecules, which are more readily accommodated in the model. Among the heterocyclic compounds of recent interest that are believed to act by blocking the GABA-gated chroride ionophore, the experimental spirosultam LY 219048 (33, Fig. 55.11) contains an obvious lipophilic bulky substituent in the form of the cyclohexane ring, in analogy with the compounds discussed previously. Probably because this ring may be susceptible to metabolic attack, the compound is not very toxic
1151
to insects but its toxicity to mice is similar to that of endrin (Bloomquist et aI., 1993). Other insecticidal compounds such as the phenylpyrazoles (e.g., fipronil; 27, Fig. 55.11) have little obvious structural similarity to the compounds discussed in previous sections yet inhibit eH]-EBOB binding to housefly head membranes and are believed to act at the PCCAlPTX binding site (Cole et aI., 1993). Some other compounds containing the common 2,6-dichloro-4-trifluoromethylphenyl moiety combined with a pyrimidinone (e.g., 32, Fig. 55.11) and other small heterocyclic rings were described by Whittle et al. (1995) but these lacked the broad-spectrum insecticidal activity shown by fipronil, which was introduced in 1993 (Colliot et aI., 1992) and now has a wide range of applications for crop protection by foliar, soil or seed treatment. Its mammalian toxicity is generally moderate (rat acute oral LDso, 97 mg/kg; mouse acute oral LDso, 95 mg/kg) and it is readily converted into degradation products, including the corresponding sulfone, in the environment (Tomlin, 1997). The substitution pattern in the phenyl and heterocyclic rings of these compounds places these ring planes at right angles (Whittle et aI., 1995) so that the molecules are "space filling" in these two planes and are relatively rigid, with the pyrazole ring skewed with respect to the benzene ring because of the pyramidal linking nitrogen. The substituents (-NH2, -CN, -SOCF3) are attached to double bonds and are fixed in the plane of the pyrazole ring, although they can rotate about their attaching bonds, which may be particularly important for the critical -SOCF3 group. This group confers insect (housefly LDso, 0.3 J.Lg/g) and vertebrate (mouse ip LDso, 30 mg/kg) toxicity, even in the analogous molecule lacking both the - NH2 and -eN substituents, which, however, clearly optimize the binding properties of fipronil. In these molecules, the sulfur atom requires bioactivation through sulfoxide/sulfone formation to confer toxicity. How do the phenylpyrazoles bind to the chloride ionophore in relation to the other molecules discussed previously? They inhibit EBOB binding in housefly head membranes noncompetitively, which may involve irreversible or slowly reversible inhibition or action at an allosteric but coupled site (Cole et a!., 1993). Whittle et al. (1995) explored a series of compounds, including pyrimidinone 32 (Fig. 55.11) on the basis of their commonality of dipole direction with phenylpyrazoles; the positive end of the dipole lies toward the benzene ring in active compounds. If this ring, with its strongly electron-withdrawing -CF3 moiety, is placed in the channel pore binding site so that the -CF3 is in a similar position to that occupied by the 4-ethynyl group of EBOB, according to the model discussed earlier, that is, in the region of Arg 300 , then the benzene ring with its chlorine atoms and the pyrazole ring with its substituents are well placed to bind with channel components in this quasi-centrosymmetrical ionophore and, notably, the -SOCF3 moiety is then placed near to Ala302 in a region occupied by the endosulfan ester oxygens, the oxygens of TBOs, the oxygens or sulfurs of dioxans and oxidized trans- (linear) dithianes, and the second electronegative center (e.g., the epox-
1152
CHAPTER 55
Interactions with the gamma-Aminobutyric Acid A-Receptor: Polychlorocyc1oalkanes and Recent Congeners
ide ring of dieldrin) of cyclodienes, when these molecules are located in the lipophilic pocket bounded by Leu 303 . These molecules might penetrate the chloride ionophore from either the extracellular (synaptic) mouth or the cytoplasmic mouth or by first penetrating the lipid bilayer and then entering the channel laterally. Current understanding of the nACh-receptor (nAChR) ionophore, which has analogies with the GABAA ionophore, may be relevant. According to Unwin (1995), the M2 cylinders (alpha-helices) (Fig. 55.12) are kinked inward in the region of a ring of leucines when the ionophore is closed; in the open configuration, the cylinders are twisted laterally, moving the leucine side chains away from the pore. By analogy, a cyclodiene (for example) may enter a similar open conformation in the chloride ionophore and then become tightly bound when the pore reverts to the closed conformation, as suggested by ffrench-Constant et al. (1995). Nakanishi et al. (1997) provide an interesting discussion on the question of mode of channel entry, based on research on philanthotoxins (PhTXs) binding to the nAChR. In particular, they noted that an n-butyl side chain introduced into the hydrophilic polyamine chain of PhTX-433 increases potency eight-fold. The n-butyl moiety is so placed as to reach the area of Leu 251 when the molecule is inserted linearly into the AChR ionophore from the cytoplasmic end, and the authors speculate that the increased potency might result from additional hydrophobic binding between the n-butyl side chain and the alkyl groups of Leu 251 ; potency decreases below that of the parent PhTX-433 when this side chain is placed in other positions along the polyamine chain. Hydrophobic binding must also have a significant role in the interaction of the cage convulsants with the chloride ionophore. Here we recall that the hybrid (unchlorinated) molecule 4 (Fig. 55.10) combines the bulky norbornene (hydrophobic binding) and dioxan moieties with the additional binding capacity evidently conferred by the benzene ring with its electronegative 4-ethynyl substituent; it is more toxic (when oxidative metabolism is inhibited) than many fully chlorinated cyclodienes. No doubt work with the affinity probes currently under development (Casida and Pulman, 1994) will resolve some of these questions regarding the exact location of binding sub sites in the chloride ionophore.
55.5 CONCLUSIONS Some 45 years after serious toxicological research on the PCCA insecticides began, there is now a broad understanding of their mode of action and, due to rapid progress in the application of molecular biology, the tantalizing mechanism of insect resistance to them has been illuminated. Furthermore, new structural classes of chemicals have now emerged that appear to act in the same way. Although many chemicals found to interact with the picrotoxinin binding site in the GABAA -receptor chloride ionophore are highly toxic to both mammals and insects, binding studies with radioligands have indicated differences between their GABA-receptors that offer the prospect of selective insect toxicity involving this target.
The commercially successful insecticide fipronil appears to fulfill these expectations and other new chemicals are likely to follow. There is also the possibility of "building in" selectivity by using the "propesticide" approach, which exploits differences between insects and nontarget organisms in their biotransformation routes for chemically derivatized toxic ants and has been applied successfully to alleviate the mammalian toxicity of other classes of insecticides. Meanwhile, many questions remain to be answered that are of fundamental importance in understanding the molecular action of neurotransmitters and insecticides on ion channels. The intense interest in this subject, stimulated by rapid advances in molecular biology, will ensure a prominent use for the PCCA insecticides and the newer chemicals with related actions, as tools in these explorations.
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Structural features and mechanisms of proinsecticidal action and selective toxicity. 1. Agric. Food Chem. 39, 1335-1341. Peterson, E. M., Kun Xu, Holland, K D., McKeon, A c., Rothman, S. M., Ferrendelli, J. A., and Covey, D. E (1994). Alpha-spirocyclopentyl and alpha-spirocyclopropyl-gamma-butyrolactones: Conformationally constrained derivatives of anticonvulsant and convulsant alpha, alphadisubstituted gamma-butyrolactones. J. Med. Chem. 37, 275-286. Pollock, G. A., and Kilgore, W. W. (1980). Toxicities and descriptions of some toxaphene fractions: Isolation and identification of a highly toxic component. 1. Toxico!. Environ. Health 6, 115-125. Pulman, D. A, Smith, I. H., Larkin, J. P., and Casida, J. E. (1996). Heterocyclic insecticides acting at the GABA-gated chloride channel: 5-Alkyl-2arylpyrimidines and 1,3-thiazines. Pestic. Sci. 46,237-245. Ray, J. W (1967). The epoxidation of aldrin by housefly microsomes and its inhibition by carbon dioxide. Biochem. Pharmaco!' 16,99-107. Reddy, G., and Khan, M. A Q. (1977). Metabolism of [I4C]_photodieldrin in house flies. J. Agric. Food Chem. 25, 25-28. Revah, E, Bertrand, D., Galzi, J. L., Devillers-Theiry, A., Mulle, c., Hussy, N., Bertrand, S., Ballivet, M., and Changeux, J. P. (1991). Mutations in the channel domain alter desensitisation of a neuronal nicotine receptor. Nature (London) 353,846-849. Richardson, A., Baldwin, M. K., and Robinson, J. (1968). Metabolites of dieldrin in the urine and faeces of rats. Chem. Ind. (London) 588-590. Ryan, K J., and Engel, L. L. (1957). Hydroxylation of steroids at carbon-21. J. Bio!. Chem. 225, 103-114. Saleh, M. A, Skinner, R. E, and Casida, J. E. (1979). Comparative metabolism of 2,2,5-endo,6-exo,8,9,IO-heptachloronorbomene and toxaphene in six mammalian species and chickens. J. Agric. Food Chem. 27,731-737. Schmidt, W E, Hapeman, C. J., Fettinger, J. c., Rice, C. P., and Bilboulian, S. (1997). Structure and asymmetry in the isomeric conversion of beta- to alpha-endosulfan. J. Agric. Food Chem. 45, 1023-1026. Schofield, P. R.., Darlison, M. G., Fujita, N., Burt, D. R., Stephenson, E A, Rodriguez, H., Rhee, L. M., Ramachandran, J., Reale, v., Glencorse, T. A, Seeburg, P. H., and Barnard, E. A (1987). Sequence and functional expression of the GABAA -receptor shows a ligand-gated receptor super-family. Nature (London) 328, 221-227. Schroeder, M. E., Shankland, D. E., and Hollingworth, R. M. (1977). The effects of dieldrin and isomeric diols on synaptic transmission in the American cockroach and their relevance to the dieldrin poisoning syndrome. Pestic. Biochem. Physio!. 7,403-415. Shankland, D. L. (1982) Neurotoxic action of chlorinated hydrocarbon insecticides. Neurobehav. Toxico!. Terato!' 4, 805-811. Shankland, D. L., and Schroeder, M. E. (1973). Pharmacological evidence for a discrete neurotoxic action of dieldrin (HEOD) in the American cockroach, Perplaneta americana (L.). Pestic. Biochem. Physio!. 3, 77-86. Slade, R. E. (1945). The gamma-isomer of hexachlorocyclohexane (gammexane). Chem.Ind. (London) 40, 314-319. Slade, M., Brooks, G. T., Hetnarski, H., and Wilkinson, C. E (1975). Inhibition of the enzymatic hydration of the epoxide HEOM in insects. Pestic. Biochem. Physio!. 5, 35-46. Smith, A G. (1991). Chlorinated hydrocarbon insecticides. In "Handbook of Pesticide Toxicology" (W J. Hayes, Jr., E. R. Laws, Jr., eds.), Vol. 2, pp. 731-915. Academic Press, New York. Smith, I. H., Budd, T. c., Sills, J. H., and Casida, J. E. (1993). Insecticidal 1-(alkynyl alkyl)-3-cyano-2,6,7-trioxabicyclo[2.2.1] octanes. J. Agric. Food Chem. 41, 1114-1117. Soloway, S. B. (1965). Correlation between biological activity and molecular structure of the cyclodiene insecticides. Adv. Pest Control Res. 6, 85-126. Stemburg, J., Keams, C. W, and Moorefield, H. (1954). DDTdehydrochlorinase, an enzyme found in DDT-resistant flies. J. Agric. Food Chem. 2, 1125. Street, J. C., and Blau, S. E. (1972). Oxychlordane. Accumulation in rat adipose tissue on feeding chlordane isomers or technical chlordane. J. Agric. Food Chem. 20,395-397. Sun, Y. P., and Johnson, E. R. (1960). Synergistic and antagonistic actions of insecticide-synergist combinations and their mode of action. J Agric. Food Chem. 8, 261-266.
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CHAPTER 55
Interactions with the gamma-Aminobutyric Acid A-Receptor: Polychlorocycloalkanes and Recent Congeners
Takeuchi, A" and Takeuchi, N. (1966). On the permeability of the presynaptic terminal of the crayfish neuromuscular junction during synaptic inhibition and the action of GABA. J. Physiol. 183,433-449. Takeuchi, A., and Takeuchi, N. (1972). Actions of transmitter substances on the neuromuscular functions of vertebrates and invertebrates. Adv. Biophys. 3, 45-95. Tanaka, K., Scott, J. G., and Matsumura, F. (1984). Picrotoxinin receptor in the central nervous system of the American cockroach: Its role in the action of cyclodiene-type molecules. Pestic. Biochem. Physiol. 22, 117-127. Tashiro, S., and Matsumura, F. (1977). Metabolic routes of cis- and transchlordane in rats. J. Agric. Food Chem. 25, 872-880. Tomlin, C. D. S., ed. (1997). "Pesticidehonval," 4th ed., PPC, 545-547. British Crop Protection Council, Farnham, Surrey, UK. Turner, W. v., Engel, J. L., and Casida, J. E. (1977). Toxaphene components and related compounds: preparation and toxicity of some hepta-, octaand nonachloronorbornanes, hexa- and heptachlorobornenes, and a hexachlorobornadiene. J. Agric. Food Chem. 25, 1394-1401. Uchida, M., Fujita, T., Kurihara, N., and Nakajima, M. (1978). Toxicities of gamma-BHC and related compounds. In "Pesticide and Venom Neurotoxicity" (D. L. Shankland, R. M. Hollingworth, and T. Smyth, Jr., eds.), pp. 133151. Plenum, New York. Ullman and Wong et al. (1972). Unwin, N. (1995). Acetylcholine receptor channel imaged in the open state Nature (London) 373, 37-43. Von Keyserlingk, H. c., and Willis, R. J. (1992). The Gaba-activated chloride channel in insects as target for insecticide action-A physiological study. In "Insecticides: Mechansim of Action and Resistance" (D. Otto and B. Weber, eds.), pp. 205-236. Intercept, Andover, UK.
Wacher, V. J., Toia, R. F., and Casida, J. E. (1992). 2-Aryl-S-tert-butyl-1,3dithianes and their S-oxidation products: Structure-activity relationships of potent insecticides acting at the GABA-gated chloride channel. J. Agric. Food Chem. 40, 497-505. Wang, C. M., Narahashi, T., and Yamada, M. (1971). The neurotoxic action of dieldrin and its derivatives in the cockroach. Pestic. Biochem. Physiol. 1, 84-91. Weil, E. D., Colson, J. G., Hoch, P. E., and Gruber, R. H. (1969). Toxic chlorinated methanoisobenzofuran derivatives. J. Heteroeyc. Chem. 6,643-649. Weston, J. B., Larkin, J. P., Pulman, D. A., Holden, 1., and Casida, J. E. (1995). Insecticidal isomers of 4-tert-butyl-1-(4-ethynylcyclohexyl)-2,6,7trioxabicyclo[2.2.1J octane and S-tert-butyl-2-(4-ethynylcyclohexyl)-1,3dithiane. Pestic. Sci. 44, 69-74. Whittle, A. J., Fitzjohn, S., Mullier, G., Pearson, D. P. J., Perrior, T. R., Taylor, R., and Salmon, R. (1995). The use of computer-generated electrostatic surface maps for the design of new GABA-ergic insecticides. Pestic. Sci.
44,29-31. Wong and Terriere (1965). Yarbrough, J. D., Grimley, J. M., Karl, P. 1., Chambers, J. E., Case, R. S., and Alley, E. G. (1983). Tissue disposition, metabolism and excretion of cisand trans-S,lO-dihydrogen mirex. Drug Metab. Dispos. 11,611-614. Zhang, H.-G., ffrench-Constant, R. H., and Jackson, M. B. (1994). A unique amino acid of the Drosophila GABA-receptor with influence on drug sensitivity by two mechanisms. J. Physiol. 479,65-75. Zhang, H.-G., Lee, H.-J., Rocheleau, T., ffrench-Constant, R. H., and Jackson, M. B. (1995). Subunit composition determines picrotoxinin and bicuculline sensitivity of Drosophila GABA-receptors. Mol. Phannacol. 48, 835-840.
CHAPTER
56 The Avermectins: Insecticidal and Antiparasitic Agents Jim Stevens and Charles B. Breckenridge Syngente Crop Protection
56.2 CHEMISTRY AND FORMULATIONS
56.1 INTRODUCTION The avermectins are macrocyclic lactones isolated from the fermentation broth of the soil actinomycete, Streptomyces avermitilis. Included in this avermectin group are abamectin and emamectin benzoate, which are used as insecticides, and ivermectin, which is sold for parasite control in human and veterinary medicine. Because the avermectins act as GABAA receptor agonists in vertebrates, their general safety for use as pest control agents in mammals depends on an intact bloodbrain barrier in juvenile and adult animals and an intact bloodplacental barrier in utero. Inherent in the integrity of these barriers is the substance P-glycoprotein. Intact P-glycoprotein barriers are present in human adults (male and female, pregnant and nonpregnant), newboms, and children. This fact is supported by significant clinical evidence from in excess of 50 million people; a genetically varied population throughout the world, including thousands of pregnant women, have been administered therapeutic doses (0.15-0.20 mg/kg) of ivermectin. However, there is also experimental evidence that certain laboratory animals, such as genetically polymorphic CF-l mice and rat pups early postnatally, do not possess intact P-glycoprotein blood-brain barriers. Unfortunately, in the process of establishing the hazard profile for the mectins, both of these models were used. This has resulted in very low NOELs (no observed effect levels) based on neurotoxicity in the CF-l polymorphic mouse. The World Health Organization's JMPR and the U.S. Environmental Protection Agency have concluded that the CF-l polymorphic mouse is not an appropriate model for human risk assessment for the avermectins; the JMPR has recognized the flawed nature of the rat pup model for testing of these avermectins. However, a consistent understanding of the inappropriateness of applying results from standard animal models used for toxicity testing to human risk assessment is yet to be fully appreciated. Handbook of Pesticide Toxicology Volume 2. Agents
Abamectin belongs to a general class of closely related macrocyclic lac tones either produced directly by the actinomycete Streptomyces avermitilis or generated through semisynthetic modifications (Fisher and Mrozik, 1989). The structure for the natural avermectins is given in Fig. 56.1. The basic structural motif of the avermectins is evident in the natural product avermectin B 1a, which is the principal constituent of the insecticide abamectin. As used in pesticides, abamectin consists of 80% or more of avermectin Bla and 20% or less of avermectin BIb and is called avermectin BI (Fisher and Mrozik, 1989). Their structures are shown in Fig. 56.2. Chemical modification of avermectin Bla has yielded a number of semisynthetic materials. Emamectin (4/1-epimethylamino-4/1 -deoxyavermectin B la) benzoate is shown in Fig. 56.3. Emamectin and ivermectin differ from avermectin B I by having only a single bond at the C22C23 position (instead of a double bond). The major manufacturers, trade names, and formulations for abamectin, emamectin benzoate, and ivermectin are given in Table 56.1.
56.3 USES Abamectin and emamectin benzoate (Novartis) are used as insecticides and ivermectin (Merck) is sold for parasite control in human and veterinary medicine. Abamectin and emamectin migrate into treated leaves, exhibit oral activity against insect pests, and display rapid breakdown in sunlight; all of these features favor their use in integrated pest management (Bloomquist, 1999). Abamectin is used primarily to control mites, and emamectin benzoate has been designed primarily for control of lepidopterian species in vegetable, cotton, and tobacco. Ivermectin has found great favor in both the pharmaceutical and the veterinary product marketplaces. In human medicine,
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Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
1158
CHAPTER 56 The Avermectins: Insecticidal and Antiparasitic Agents ivermectin has been used as an anthelmintic in the treatment of infection of intestinal threadworm, river blindness (onchocerciasis), and lymphatic filariasis. Its uses in veterinary medicine have been as anthelmintic and antiparasitic agents, including treatment of heartworm, hookworm, threadworm, and whipworm (FAOIWHO, 1992; 1993; Greene et aI., 1989).
Component A: Component B: Component 1: Component 2: Component a: Component b:
Figure 56.1
RS;: CH3 RS;: H X '"' -CH=CHX = -CH2CHOH R26 ;: C2HS R26 = CH3
56.3.1 MODE OF ACTION OF THE AVERMECTINS
Structure of the natural avermectins.
H HC:0 0 0:o 3
HO ....
3
H
3
H COO,
H
3
HaC ....
x .. -CH=CH-
: Abamectin
X .. -C~-C~- : lvermectin Figure 56.2
Structures of ivermectin and abamectin.
-.X
H3
HO .... H
cJ....o}oUHa HaC H
3
0
0,
H H3C'"
Figure 56.3
The avermectins were first demonstrated to possess anthelmintic activity in 1979 (Burg et aI., 1979). The avermectins act as chloride channel-blocking insecticides, causing hyperexcitability and convulsions. Arena et al. (1995) demonstrated that in insects stimulation of glutamate (inhibitory) chloride channels is the most sensitive target site for the avermectins. The glutamate-gated chloride channels of insect and nematode skeletal muscle are especially important as they mediate avermectin-induced muscle paralysis in these organisms. These effects are mediated via a specific, high-affinity (10- 10 M) binding site (Turner and Schaeffer, 1989). In vertebrates, the effects occur via poisoning of the central nervous system (CNS) through reactions at the receptor for the inhibitory neurotransmitter y-aminobutyric acid (GABA); see Fig. 56.4. The avermectins open the GABAA receptor chloride channel by binding to the GABA recognition site (receptor protein) and act as partial agonists (Abalis et aI., 1986). Chloride ions then flow into the postsynaptic neuron. This chloride permeability increase can significantly hyperpolarize (make more negative) the membrane potential, which has a dampening effect on nerve impulse firing. There is also a reversible dosedependent increase in chloride ion permeability in response to very low doses of avermectins. In GABA-insensitive neurons with no inhibitory innervation, the avermectins induce an irreversible increase in chloride ion conductance through interacting with voltage-dependent chloride channels. Avermectin intoxication in mammals begins with hyperexcitability, tremors, and incoordination and later develops into ataxia and coma-like sedation. This is similar to the mode of action of ethanol and barbiturates (Eldefawi and
Structure of emamectin benzoate.
Table 56.1 Chemical Structures, Trade Names, Major Manufactures, and Formulations for the Avermectins
Chemical Abamectin
Trade
Major
names
nanufactures
Agri-mek
Novartis
Emamectin Proclaim Ivomec Mectizan Stromectal
0
0~
0.15Ib/gal 0.15Ib/gal
Novartis
benzoate Denim Ivermectin
Vesicles
Formulation
Zephyr
Presynaptic Nerve Terminals
5% granule 0.16Ib/gal
Merck & Co.
Tablets Injectables
Convulsant • • block Avermeetlns • Ictlvate
G
GABA..>l
Postsynaptic Cell
GABA Receptor I Cl Channel Complex
"1/
Block of Aetion Potential
Inhibitory Postsynaptic Potential
Figure 56.4 Depiction of the mechanism of action for the avermectins in the brain (after Bloomquist, 1999).
56.3 Uses
Eldefawi, 1987) and benzodiazepine sedatives (Williams and Yarbrough, 1979). However, the avermectins are less specific in their action and can affect a variety of other ligand- and voltage-gated chloride channels. The general safety of the avermectins depends on the presence of an intact P-glycoprotein blood-brain barrier.
56.3.2 IMPORTANCE OF THE P-GLYCOPROTEIN BLOOD-BRAIN BARRIER
In mammals, the circulatory system is the most important transport system in the body. The microvasculature, which includes an estimated 40 billion capillaries in the human, serves as a site of substance exchange and is vital for the transportation of chemicals throughout the body (Aigner et aI., 1997). These capillaries consist of a single layer of endothelial cells (continuous, discontinuous, and fenestrated) surrounded by a basement membrane. The capillaries with continuous endothelial cells are only present in a few organs, including the intestine, bile duct, liver, pancreas, kidney, adrenal, testes, placenta, and brain (Thiebaut et aI., 1987). In addition to this structural feature, two other pathways exist at some sites in the brain and the placenta for the further elimination of xenobiotics and intracellular metabolites. These pathways are (1) biotransformation and (2) direct transport by transmembrane pumps such as P-glycoprotein (Fisher and Sikic, 1995; Gottesman and Pastan, 1993). P-glycoproteins are large-membrane proteins (150-180 kDa) consisting of two identical subunits each with a single adenosine 5' -triphosphate (ATP)-binding site and several transmembrane domains (Juliano and Ling, 1976). They are highly expressed in endothelial cells at areas that have a barrier function (e.g., the blood-brain barrier and the blood-placental barrier). P-glycoprotein barriers have also been identified in the adrenal gland, colon, testes, and the gravid uterus (Tiirikainen and Krusius, 1991). In addition, its localization along the apical surfaces of the intestines, proximal tubule of the kidney, and the bile ducts of the liver imply that P-glycoprotein is probably involved in secretory functions in humans. P-glycoprotein is a member of a highly conserved multigene family with isoforms identified in a wide variety of mammalian species, including humans, rats, mice, hamsters, pigs, guinea pigs, rhesus monkeys, orangutans, cows, and chickens (Saunders, 1977). Further, P-glycoprotein has been identified in fruit flies (Wu et aI., 1991) and in tobacco hornworms and budworms (Lanning et aI., 1996). In addition, P-glycoprotein has been found in mussels and sponges (Kurelec and Pivcevic, 1991), in yeast (McGrath and Varshavsky, 1989), in parasites (Descoteaux et aI., 1992), and in nematodes (Lincke et aI., 1992). The fact that P-glycoprotein is conserved across phylogenic lines suggests that it is an ancient protein associated with fundamental cellular functions.
1159
56.3.2.1 Implications of an Incomplete P-glycoprotein Blood-Brain Barrier in Hazard Testing
Unlike other transporters, P-glycoprotein transports a variety of chemically unrelated compounds. These compounds are commonly large lipophilic molecules that contain at least one aromatic ring and a positively charged nitrogen atom. Initially, P-glycoprotein interactions were thought to be limited to only the natural products such as anthracyclines, vinca alkaloids, actinomycin D, epipodophyllotoxins, taxol, and taxotere. However, it is now known that P-glycoprotein also transports steroid hormones, peptide antibiotics, immunosuppressive agents, and calcium channel blockers (Ueda et aI., 1997). Recently, pesticides have also been demonstrated to interact with P-glycoprotein. Such agents include abamectin (Didier and Loor, 1996), ivermectin (Schinkel et aI., 1995), 2-acetylaminofluorene and pentachlorophenol (Toomey and Epel, 1995), thiodicarb, and chlorpyrifos (Lanning et aI., 1996). P-glycoprotein is encoded as three isoforms. Mouse P-glycoproteins are known as mdrla (also called mdr3 or Pgyl), mdrlb (also called mdrl or Pgy2), and mdr2 (Pgy3) (Borst et aI., 1993). The isoform mdrla is primarily found in the intestinal brush border epithelium, the microvessel endothelial cells in the brain and testis, and the microvillus border of the trophoblast and Hofbauer cells of the placenta (MacFarland et aI., 1994; Nakamura et aI., 1997). The isoform mdrlb has been reported in the adrenal gland, kidney, and gravid uterus. Both mdrla and mdrlb play roles in the multidrug resistance phenotype and are able to pump xenobiotics out of the cells. The isoforms mdr 1a and mdr 1b of the mouse are comparable to the human MDRl gene in this regard (Mauad et aI., 1994). The isoform mdr2 has been shown necessary for bile production and is probably the phosphatidyl choline transporter. Mouse mdr2 is comparable to the human MDR3 (also called MDR2) gene and does not confer multidrug resistance to cells. P-glycoprotein transport processes have been conserved across species, probably because such a transporter system is essential for adaptation and survival (Saunders, 1977). It is therefore probably not surprising that a population genotypically recessive for P-glycoprotein has not been identified. However, it has been possible to develop a "knockout" mouse for the mdrla gene (Schinkel et aI., 1994; 1995). In addition, polymorphism for mdrla P-glycoprotein gene expression has been reported for the CF-l mouse (Umbenhauer et aI., 1997) and the collie dog (Lankas et aI., 1997). 56.3.2.2 Impact of an Incomplete P-glycoprotein Blood-Brain Barrier in Animal Model
Animals with a recessive mdrla (- / -) genotype do not have an intact blood-brain or blood-placental barrier because they are deficient in P-glycoprotein expression. Studies using knockout mice homozygous for disruption of mdr1a (Schinkel et al., 1994; 1995) have clearly demonstrated that the presence/absence of P-glycoprotein is a major determinant of drug entry in the brain. Studies with CF-1 mice polymorphic for
1160
CHAPTER 56
The Avermectins: Insecticidal and Antiparasitic Agents
P-glycoprotein (Umbenhauer et aI., 1997) have shown this same response. Mice with disrupted P-glycoprotein and CF-1 mice without P-glycoprotein were shown to be significantly more susceptible to the effects of neurotoxicants. Brain levels of ivermectin in the knockout mice that do not express P-glycoprotein [mdrla (- / -) genotype] were elevated approximately 90-fold over the wild type [mdr 1a (+/ +) genotype] and sensitivity to ivermectin was increased (Schinkel et aI., 1994; 1995). Further, CF-1 mice show a unique developmental response to avermectins due to the polymorphic nature of the P-glycoprotein gene. These mdr 1a ( - / -) animals could be present in a population of CF-1 mice in a range from 0 to 100%, depending on the genotype of the parental animals (Umbenhauer et aI., 1997). Therefore, experiments carried out with P-glycoprotein substrates in the heterogeneous population of the CF-1 mouse must be interpreted with caution and may be unsuitable for risk assessment. Besides the CF-1 mouse model, there are other unique features noted in the standard hazard testing models. The SpragueDawley rat pup does not establish a complete P-glycoprotein blood-brain barrier until appropriately 3 weeks postpartum, making it highly vulnerable to neurotoxic effects (Lankas et aI., 1989).
56.4 HAZARD IDENTIFICATION AND DOSE RESPONSE As previously indicated, in vertebrates, the avermectins increase membrane permeability to chloride ions and act as GABAA agonists; this is similar to the mode of action of the benzodiazepine sedatives (Turner and Schaeffer, 1989). Their toxicity follows this mode of action in overdose scenarios or in animal models with compromised mdrla P-glycoprotein barriers. The acute toxicology profiles for ivermectin and abamectin (EPA, 1999a; Lankas and Gordon, 1989) and emamectin benzoate (EPA, 1999b) are shown in Table 56.2. Table 56.2 Acute Oral Toxicity Studies on Ivennectin, Abamectin (Lankas and Gordon, 1989; EPA, 1999a), and Emamectin Benzoate (EPA, 1999b) LD50 (mg/kg) Species SD rat SD rat, neonatoa
Ivennectin 50 2
11
CF-l mouseb
25
Beagle dogs
80 > 24
Emamectin
Monkey Ivennectin
14-24
Humans
Abamectin
Therapeutic dose
Ivennectin 0.2 mg/kg
Peak plasma levels
20 ng/ml
Minimum effect level
2 mg/kg
2 mg/kg
Peak plasma levels
IlOng/ml
76 ng/ml
Signs of toxicity
Emesis
Emesis
8 mg/kg
8 mg/kg
Toxic effect level
6.6-8.6 mg/kga
Peak plasma levels
270 ng/ml
150 ng/ml
Unknown
Signs of toxicity
Emesis
Emesis
Emesis, mydriasis,
24 mg/kg
24 mg/kg
Peak plasma levels
680 ng/ml
390 ng/ml
Signs of toxicity
Emesis,
Emesis,
mydriasis,
mydriasis,
sedation
sedation
sedation Overdose level
aOverdose in humans.
These three avermectin-derived materials responded quite similarly in the different laboratory models. A comparison of the response of ivermectin and abamectin in the monkey as well as the response noted in the human with ivermectin is presented in Table 56.3. Signs of overdosing noted at 24 mg/kg of ivermectin or abamectin in the monkey were the same as observed in the human overdose at approximately 9 mg/kg. These signs were essentially identical to those observed in 10 adults who accidentally ingested tablets or solutions intended for veterinarian use (Greene, 1991). Subchronic dietary exposure to the three avermectins in the rat, mouse, and dog yielded similar results, as shown in Table 56.4. Although there are slight differences between the NOELs for abamectin and emamectin in the CD-1 mouse (probably the result of dose selection), the responses noted in the rat and dog were similar for the three avermectins. The aver-
Table 56.4 90-Day Dietary Toxicity Studies with Ivennectin (FAOIWHO, 1994), Abamectin (EPA, 1999a), and Emamectin Benzoate (EPA, 1999b)
76-88 Exposure (mg/kg/day)
1.5 220
CD-l mouse
Rhesus monkeys
Abamectin
Table 56.3 Acute Toxicity and Plasma Concentrations of Ivennectin and Abamectin (Lankas and Gordon, 1989)
107-120 22-31
Ivennectin
Abamectin
Emamectin
mg/kg
mg/kg
mg/kg
Study > 24
ap-glycoprotein-deficient blood-brain barrier is seen in neonatal rats (Lankas et aI., 1989). bCF-l mice tested were polymorphic for P-glycoprotein (Umbenhauer et aI., 1997).
SDrat
0.8
0.4
1.4
0.4
2.5
0.5
CD mouse
Nsa
NS
8
4
5.4
0.5
Beagle dog
2.0
0.5
> 1.0
0.5
0.5
0.25
aNo study available.
56.4 Hazard identification and dose response Table 56.5 Chronic Dietary Toxicity Studies with Ivermectin (FAOIWHO, 1994), Abamectin (EPA, 1999a), and Emamectin Benzoate (EPA, 1999b)
Exposure (mg/kg/day) Ivermectin
Abamectin
Emamectin
mglkg
mg/kg
mg/kg
Study SDrat
0.8
0.4
2.0
1.5
2.5
0.25
8
4
5.0 (M)
2.5
(105 weaks)a CD-1 mouse (18 months)
7.5 (F) 1.0
Beagle dog
0.5
0.5
0.25
0.5
0.25
(12 months) aRat studies with ivermectin and abamectin were only 53 weeks in duration.
mectins are equally well tolerated following chronic dietary administration, as shown in Table 56.5. The NOELs found in the chronic dog and mouse oncogenicity studies were comparable for ivermectin, abamectin, and emamectin. The NOEL in the chronic rat study was higher for abamectin than for emamectin or ivermectin. This apparent difference between abamectin and emamectin was most likely due to differences in dose selection between the studies. The avermectins are not genotoxic as has been demonstrated in a variety of standard tests for mutagenicity, clastogenicity, and unscheduled deoxyribonucleic acid (DNA) synthesis, as presented in Table 56.6. The maternotoxicity and developmentallfetotoxicity NOELs and LOELs for these three avermectins are shown in Table 56.7. In the CF-l mouse, SD rat, and rabbit developmental toxicity studies with ivermectin, cleft palate and clubbed feet (rabbit only) were observed at maternally toxic doses (Lankas and
1161
Gordon, 1989). Similar findings were noted in the CF-l mouse and rabbit studies with abamectin. Neither of these effects was noted with emamectin (EPA, 1999b). Sedation was observed in overdosed rabbit dams. Severe neurotoxicity (tremors, convulsion, and coma) was observed in some of the polymorphic CF-l mice with a compromised blood-brain barrier and blood-placental barrier (Umbenhauer et aI., 1997). These effects were also observed with ivermectin administered to the rnrdla knockout mouse (Schinkel et aI., 1994; 1995). Furthermore, the incidence of cleft palate correlated with the maternal mortality in a CF-l mouse study (Lankas et aI., 1997). The incidence of cleft palate was also linked to the polymorphism of mdrla in the CF-l mouse (Umbenhauer et aI., 1997). Developmental toxicity studies have also been conducted with the 8,9-Z-isomer of abamectin in the CF-l mouse to further evaluate the phenomenon of the linkage of developmental toxicity to the blood-placental barrier (Table 56.8). The NOEL for maternal and developmental toxicity was 0.1 mg/kg/dayand 0.05 mg/kg/day, respectively. In young adult CF-l mice, which were genotyped for their P-glycoprotein expression, the brain concentrations of the isomer 8 h after treatment were 60 times higher in (-/-) males and females than in the (+/ +) male and female CF-l mice. Brain concentrations of the delta-8,9-isomer of avermectin B 1a in ( -/-) CF-l fetuses were higher than in ( -/ +) fetuses, which, in turn, were higher than in (+ / +) fetuses. To study the development of P-glycoprotein in the placenta in the CF-l mouse, normal homozygous (+ / +) female Table 56.7 Developmental Toxicity Studies with Ivermectin (FAOIWHO, Abamectin (EPA, 1999a), and Emamectin Benzoate (EPA, 1999b)
1994),
Dose (mg/kg/day) Ivermectin Study
LOEL
Abamectin
NOEL
LOEL
NOEL
Emamectin LOEL
NOEL
Matemotoxicity Table 56.6 Genotoxicity Studies with Ivermectin (FAOIWHO, 1994), Abamectin (EPA, 1999a), and Emamectin Benzoate (EPA, 1999b)
CF-l mousea.b
0.2
0.1
6.0
3.0
0.075 2.0
0.05
Rabbit
1.0
6.0
3.0
SDrat
10.0
5.0
2.0
1.6
4.0
2.0
Mectin Fetotoxicity Tests Mutation
Ivermectin Abamectin Emamectin
Ames (+/- activation)
Negative
Mouse lymphoma
Negative
v- 79 Chinese hamster lung
Negative
Negative
CF-1 mousea Rabbit
3.0
SDrat Negative
Negative
Mouse bone marrow in vivo
Negative
Negative
Chinese hamster ovary
0.4
1.5
2.0
0.2 1.0
6.0C
1O.0c
2.0
1.6
4.0
Developmental
(+ / - activation) Clastogenicity
0.8 C
CF-1 mousea
0.4
0.2
0.4
0.2
Rabbit
3.0
1.5
1.0
6.0C
Negative
SDrat
10.0
5.0
2.0 2.0d
1.6
4.0
Negative
aCF-1 animals tested were polymorphic for P-glycoprotein (Umbenhauer et aI.,
in vitro
Other
Alkaline elution/rat hepatocyte Unscheduled DNA synthesis Negative in human fibroblasts
Negative
1997). bNot evaluated with emamectin benzoate.
cNo adverse effects at the highest dose tested. dNo cleft palates seen at the highest dose tested.
1162
CHAPTER 56 The Avennectins: Insecticidal and Antiparasitic Agents
Table 56.8 Genotyping Study of CF-1 Mice Treated with
~8,9-isomer
(Wise et aI., 1997)
Control +/-F; +/-M Number of fetuses
108
5 mg/kg
+/- F; +/- M
+/-F;+/+M
+/- F; +/+M
+/- F;+/+M
105
141
125
127
o
0
18
80
9
12
12
12
o
0
6
11
31
examined Number of fetuses with cleft palate Number of litters
8
examined Number of litters with cleft palate Pups with - / - genotype
19
50
NFb
NF
Pups with + / - genotype
15
NF
NF
41
29
Pups with + / + genotype
32
NF
39
31
NF
- / - with cleft palate
0%
0%
NF
NF
97%
+ / - with cleft palate
0%
NF
NF
39%
45%
+ / + with cleft palate
0%
NF
0%
0%
NF
aGenotype for P-glycoprotein: +, functional gene; -, defective gene. bNF, genotype not found.
and homozygous (+ / +) males were mated and the concentration of P-glycoprotein was measured in the placenta (Lankas et aI., 1989). The human population is known to be homozygous positive for this gene. The human fetus is therefore protected in utero due to a good placental-blood barrier with proper expression of P-glycoprotein (MacFarland et aI., 1994; Nakamura et aI., 1997). Furthermore, this protein has been identified in the capillaries of the brain of the human fetus as early as the third trimester (28 weeks) (Van Kalken et aI., 1992). The level of expression at 28 weeks is already the same as that of an
Table 56.9 Multigeneration Reproductive Toxicity Studies in SD Rats with the Avermectins (EPA, 1999a; 1999b; Lankas and Gordon, 1989) Ivermectin
Mg/kg/day Effect
LOEL
NOEL
0.4
0.2
Abamectin LOEL
NOEL
0.4
0.12
Emamectin LOEL
NOEL
3.6
0.6
Neonatal
Neonatal
Clinical
mortality
mortality
signs
adult.
Ivermectin has been extensively used worldwide at the high doses administered to humans (in monitored clinical trials as well as in more general therapeutic applications) (FAOIWHO, 1991). If a subpopulation of humans without P-glycoprotein existed, it would have been readily identified. Because humans and other primates have not been shown to have subpopulations deficient in P-glycoprotein, the CF-l mouse developmental toxicity data are not particularly relevant for use in human risk assessments to the avermectins (FAOIWHO, 1997). The critical levels and effects for multigeneration reproduction studies with the avermectins are shown in Table 56.9. The LOEL (lowest observed effect level) and NOEL values for ivermectin and abamectin were quite similar; the LOEL and NOEL for emamectin benzoate were somewhat higher. Early postpartum rat pup mortality has been observed with all three avermectins (EPA, 1999a; 1999b; Lankas et aI., 1989). Lankas et al. (1989) observed a significant increase in mortality between days 7 and 14 postpartum in treated dams nursing treated and control pups. In contrast, the mortality, growth, and development of treated and control pups nursing from con-
trol dams were similar. Because toxicity was only observed in control and treated pups cross-fostered to treated dams, it was concluded that neonatal toxicity of ivermectin in rats was a function of postnatal lactation exposure only and not due to in utero exposure. Further, these investigators administered purified, tritium-Iabeled avermectin B1a (ivermectin B1a) and sampled plasma and milk from dams treated orally with 2.5 mg/kg/day of radiolabeled ivermectin Bra for 61 days. The pattern for pup mortality, milk concentration, pup liver, plasma, and brain concentration of ivermectin, and percentage of adult levels for P-glycoprotein in the pup brain barrier are presented in Table 56.10. In contrast to rodents, the blood-brain barrier is formed prenatally in many species, including humans (Betz and Goldstein, 1981; Bohr and Mollgard, 1974; Jette et aI., 1995; Lankas et aI., 1989; Saunders, 1977; Van Kalken et aI., 1992). Furthermore, the blood-placental barrier is also intact in human infants at birth (MacFarland et aI., 1994; Nakamura et aI., 1997). Rats also differ from humans by having an increased utilization of fats at the time of parturition (Amano, 1967; Chiu et aI.,
56.5 Humans: Experience with Ivermectin
1163
Table 56.10 Pup Mortality, Milk and Pup Tissue Toxicokinetics, and Brain P-Glycoprotein Level Following a Daily Dose of 2.5 mg/kg/day of Tritiated Ivermectin for 61 Days (Lankas et aI., 1989) Day dose
Pup
(mg/kg)
mortality
Day I
-0.22/day"
Maternal milk
Pup plasma
Pup liver
Pup
level
level
level
brain level
p-gp level (%)b
(I-lg/g)
(I-lg/g)
(I-lg/g)
(I-lg/g)
0.094
1.640
0.100
2.324
0.276
3.918
0.251
1.482
0.804
6.106
0.318
1.052
0.893
6.648
0.264
Day 2
6.5
Day 4 Day 5
Rat pup brain
19.3/day
Day 6
5.7
Day 7 Day 8
16.9/day
Day 10
4.4
Day 11
6.9
Day 14 Day 15
19 0.86/day
Day 17
37
Day 20
89
aCorrected for background by subtracting the pup mortality observed in the control group. bCukierski (1995); P-glycoprotein expressed as percent of adult level.
1986; Scow et aI., 1964). This results in a greater release of lipophilic compounds such as abamectin and ivermectin from body fat into milk. In addition, rat milk has a much higher fat content than human milk and this leads to an increased transfer of lipophilic xenobiotic compounds to the neonatal animal compared to what would be anticipated in nursing humans. Ogbuokiri et al. (1993) reported that after a single oral dose up to 12 mg ivermectin administered to lactating women who were not breastfeeding the peak concentration in plasma was seen at 4 h posttreatment. The peak concentration in milk occurred at the same time point, but was 2-3 times lower than what was seen in plasma. These findings contrast with those found in rats where the concentration of ivermectin observed in rat milk was about threefold higher than that in plasma. Based on the uniqueness of the time profile of P-glycoprotein development and milk concentration of lipophilic toxicants in rats, it can be safely concluded that these pup deaths cannot be extrapolated to humans.
56.5 HUMANS: EXPERIENCE WITH IVERMECTIN Ivermectin has been used clinically for over a decade for the control of Onchocerca volvulus and other parasites in veterinary and human medicine (FAOIWHO, 1993). Onchoceriasis is endemic in large areas of Africa and Latin America. It is estimated that nearly 20 million people are infected and another 85 million are at risk. Large and diverse populations have been treated with ivermectin in Africa (Bumham, 1993; Chijioke and Okonkwo, 1992; Chippaux et al., 1993; De Sole et aI., 1989a; 1989b; Doumbo et al., 1992; Gardon et aI., 1997;
Ogunba and Gemade, 1992; Pacque et al., 1990; 1991; Whitworth et aI., 1991), Polynesia (Cartel et aI., 1992), and Latin America (Collins et aI., 1992). Typical human doses range from 0.1 to 0.2 mg/kg (FAOIWHO, 1993). The major effect noted following the administration of ivermectin is a severe inflammatory response, called the Mazzotti reaction. The Mazzotti reaction is secondary to the efficacy of ivermectin in killing the microfiliae, which dislodge from their site of infestation and are subsequently transported in the blood and body fluids (Ackerman et aI., 1990). This acute exacerbated immune response can be characterized by pruritis, erythema, edema, vesicle formation, papule formation, and scaling. Adenitis, fever, and hypotension may occur, and severe inflammatory changes may be noted in both the anterior and the posterior segments of the eye. The World Health Organization reviewed reports on the response to treatment for over 26,000 patients administered ivermectin for parasite control (FAOIWHO, 1993). Single oral doses up to 0.2 mg/kg (bw) produced no major effects except for those resulting from the eradication of the parasite infestation (the Mazzotti reaction). The effects observed in over 200,000 patients treated with ivermectin are summarized in Table 56.1l. Although the primary effect noted following the administration of ivermectin was the Mazzotti reaction, there were two cases of serious neurological response in two patients out a population of 17,877 treated with ivermectin (Gardon et aI., 1997). Headache was a common side effect noted, but no association between headache and treatment was observed in a doubleblind study on 7148 people conducted by Bumham (Bumham, 1993). During the first year of treatment, pain, edema, itch-
1164
CHAPTER 56
The Avermectins: Insecticidal and Antiparasitic Agents
Table 56.11 Observations from Patients Treated with Ivennectin Population treated 14,911
Dose Incidence of 52 (0.35%)
(~g/kg)
Main effects observed
Reference
130--200
37 (0.25%) cases of severe
De Sole et al. (1989a)
symptomatic postural hypotension 13 (0.09%) cases of severe fever 2 (0.01 %) cases of dyspnea 118,925
835 (0.7%)
150
230 (0.19%) cases of headache
Ogunba and Gemade (1992)
210 (0.17%) cases of general pains 150 (0.12%) cases ofpruritis 120 (0.10%) cases of edema 80 (0.06%) cases offever 20 (0.02 %) cases of dizziness 15 (0.01 %) cases of vomiting 10 (0.01 %) cases of diarrhea 7,566
992 (13.1%)
150--200
Primarily Mazzotti reaction
Ogunba and Gemade (1992)
460 cases of headache 50,929
93 (1.83%)
150--200
49 cases of severe symptomatic
Chijioke and Okonkwo (1992)
postural hypotension 34 cases of severe fever 3 cases of severe dyspnea 3 cases of severe pain 7,699
100 (1.3%)
150
Primarily Mazzotti reaction
ing, and rash were found statistically associated with treatment. These reactions diminished in the second year and disappeared by the third year. Hence, in this large human study, patients treated with ivermectin did not exhibit any of the expected neurological side effects that would have occurred if the bloodbrain barrier had been compromised. The populations treated have included not only adults, but also children of all ages and, inadvertently, pregnant women. Epidemiological follow-up of more than 1000 pregnant women treated with ivermectin (primarily in the first trimester) did not yield any indication of an increase in the incidence of miscarriage, stillbirths, or congenital malformations (Burnham, 1993; Chippaux et aI., 1993; Pacque et aI., 1990). Based on this extensive human database, there should be little concern that neurotoxicity or birth defects might occur in humans exposed to the avermectins at doses less than 0.2 mg/kg.
56.6 RISK CHARACTERIZATION Previous joint meetings in 1992 (FAOIWHO, 1992) and 1994 (FAOIWHO, 1994) had established the acceptable daily intake (ADI) for abamectin at 0.0002 mg/kg bw using the NOEL (pup toxicity) of 0.12 mg/kg/day derived from the multigeneration reproduction study conducted in Sprague-Dawley rats to which a 500-fold uncertainty factor was applied (Table 56.12). This 500 x factor was based on the standard 10 x for interspecies differences and lOx for interindividual differences plus an extra 5 x due to concern over the teratogenicity of the abamectin's
De Sole et al. (1989b)
delta-8,9-isomer in the CF-l strain of mouse. In 1997, the World Health Organization's JMPR reexamined the basis for setting the ADI for abamectin and declared that the CF-l mouse was not suitable for human risk assessment because of its heterozygous (+/ - ) or homozygous ( - / - ) genetics for P-glycoprotein(FAOIWHO, 1997; Lankas eta!., 1997; Umbenhauer et aI., 1997). This same rationale, that is, the absence of P-glycoprotein blood-brain barrier in rat pups, postnatally (Betz and Goldstein, 1981; Lankas et al., 1989; Terao et al., 1996), led the JMPR to readjusted the ADI for abamectin to 0.002 mg/day (bw) by reducing the uncertainty factor from 500to 50-fold for the ADI from the rat multigeneration study. The EPA also declared that the CF-l mouse was unsuitable for human risk assessment but failed to consider the compromised unprotected postpartum period unique to rat pups and a lOO-fold uncertainty factor was applied to the NOEL for the multigeneration reproduction study conducted with abamectin (EPA, 1999a). Temporary tolerances were also set for emamectin benzoate in 1999 (EPA, 1999b); unfortunately, the EPA appears to have not applied the same criteria with regard to the inappropriateness of the CF-l mouse, polymorphic to P-glycoprotein, for human risk assessment. Not only was the NOEL for a IS-day neurotoxicity study in the CF-l mouse used, but 300-fold uncertainty factor was also applied to the NOEL derived from this study in order to calculate the reference dose (RID). Under appropriate testing conditions, these avermectins are not developmental toxins or reproductive toxins; neither are the
1165
References
Table 56.12 Chronic ADIIRtD Values Established for the Avermectins ADIor Study/
Uncertainty
RID
factor
mglkg/day
Avermectin
Organization
incidence used
Ivermectin
JECFA
Developmental
(FAOIWHO, 1992)
Abamectin
JMPR (FAOIWHO,1994)
10000intraspecies
toxicity (CF-I mouse) Multigeneration
10000intraspecies
reproduction (SD rat)
0.001
10---interspecies 0.0002
10---interspecies 5-teratogenic concerns
JMPR (FAOIWHO,1997) EPA,1999a
Multigeneration
10000intraspecies
reproduction (SD rat) Multigeneration
10---intraspecies
reproduction (SD rat) Emamectin
EPA,1999b
15-day neurotoxicity study (CF-1 mouse)
benzoate
0.002
5-interspecies 0.0012
10---interspecies 10---intraspecies
0.00025
10---interspecies 3-short duration of study used
genotoxic or carcinogenic. Further, the hazard profiles for the three averrnectins evaluated are qualitatively and often quantitatively similar. Because of these facts, it would appear to be appropriate to use critical values from the avermectin used extensively clinically for the risk characterization of this class of chemical. Therefore, for acute risk characterization, a clinical dose of 0.2 mg/kg could be used as the NOEL (FAOIWHO, 1993). Conservatively, a 10 x interindividual uncertainty factor could be applied, as well as the 3 x -uncertainty factor proposed by the EPA, that is, 30 x, for an acute RID of 0.0067 mg/kg. Likewise, using the common mechanism approach for the averrnectins, the chronic RID based on the NOEL for the I-year dog study (the most sensitive species in the chronic studies) would be suitable. The NOEL for the dog with ivermectin and abamectin (Lankas and Gordon, 1989) and emamectin benzoate (EPA, 1999b), that is, 0.5, 0.25, and 0.25 mg/kg/day, respectively, should be used instead of relying on the data from studies of shorter duration. The chronic RID of emamectin, based on either study, would be 0.25 mg/kg/day divided by lOOx and 3 x, or 0.00083 mg/kg/day, which is essentially the same as the JMPR AD! for abamectin (FAOIWHO, 1997).
56.7 CONCLUSIONS The use of inappropriate animal models for characterizing human risk unfortunately is not well recognized. The hazard assessment of the avermectins provides a better understanding of just two instances in which our surrogate models for humans fail. First, CF-l mice are more sensitive to the avermectins due to its heterozygous expression of P-glycoprotein. Second, the increased postnatal pup mortality in rats is due to a lack of expression of P-glycoprotein at birth and a high milk concentration of lipophilic toxicants unique to this species. The clinical use of ivermectin as well as special studies provides reassUf-
ance that humans are homozygous positive (unimodal) for the P-glycoprotein gene. Further, humans express P-glycoprotein fully at birth. Therefore, as long as animal models continue to be used to characterize potential risks to the human population, it will be critical to appreciate the appropriateness or inappropriateness of the genetics that drive the biological responses in animal surrogates.
REFERENCES Ackerman, S. J., Kephart, G. M., Francis, H., Awadzi, K., Gleich, G. J., and Ottesen, E. A. (1990). Eosinophil degranulation: An immunologic determinant in the pathogenesis of the Mazzotti reaction in human onchoceriasis. J.Immunol. 144,3961-3969. Abalis, I. M., Eldefawi, A. T., and Eldefawi, M. E. (1986). Actions of avermectin B la on the gamma-aminobutyric acid A receptor and chloride channels in rat brain. J. Biochem. Toxicol. 1, 69-82. Aigner, A., Wolf, S., and Gassen, H. G. (1997). Transport and detoxication: Principles, approaches, and perspectives for research on the blood-brain barrier. Angew. Chem., Int. Ed. Engl. 36,24-41. Amano Y. (1967). Changes of the levels of blood glucose during pregnancy in the rat. Jpn. J. Pharmacol. 17, 105-114. Arena, J. P., Lui, K. K., Paress, P. S., Frazier, E. G., Cully, D. E, Mrozik, H., and Schaeffer, J. M. (1995). The mechanism of action of avermectins in Caenorhabditis elegans: Correlation between activation of glutamatesensitive chloride current, membrane binding, and biological activity. J. Parasitol. 82,286-291. Betz, L., and Goldstein, G. W. (1981). Developmental changes in metabolism and transport properties of capillaries isolated from rat brain. J. Physiol. 312,365-376. Bloomquist, J. R. (1999). "Insecticides: Chemistries and Characteristics." Virginia Polytechnic Institute and State University, Blacksburg, VA. Available at http://ipmworld.umn.edulchapterslbloomq.htm. Bohr, v., and Mollgard, K. (1974). Tight junctions in human fetal choroid plexus visualized by freeze etching. Brain Res. 81, 314-318. Borst, P., Schinkel, A. H., Smit, J. J. M., Wagennar, E., van Deemter, L., Smith, A. J., Eijdems, E. W. H. M., Baas, E, and Zaman, G. J. R. (1993). Classical and novel forms of multidrug resistance and the physiological functions ofP-glycoproteins in mammals. Pharmacol. Ther. 60, 289-299.
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CHAPTER 56
The Avermectins: Insecticidal and Antiparasitic Agents
Burg, R. w., Miller, B. M., Baker, E. E., Bumbaum, J., Cunie, S. A., Hartman, R., Kong, Y.-L., Monaghan, R. L., Olson, G., Putter, I., Tunac, J. B., Wallick, H., Stapley, E. 0., Oiwa, R., and Omura, S. (1979). Avermectins, new family of potent anthelmintic agents; producing organism and fermentation. Antimicrab. Agents Chemother. 15,361-367. Bumham, G. M. (1993). Adverse reactions to invermectin treatment for onchocerciasis: Results of a placebo-controlled, double-blind trial in Malawi. Trans. R. Soc. Trap. Med. Hyg. 87, 313-317. Cartel, J. L., Nguyen, N. L., Moulia-Pelat, J. P., Plichart, R., Martin, P. M. v., and Spiegel, A (1992). Mass chemoprophylaxis of lymphatic filariasis with a single dose of ivermectin in a Polynesian community with a high Wuchereria bancrofti infection rate. Trans. R. Soc. Trop. Med. Hyg. 86, 537-540. Chijioke, C. P., and Okonkwo, P. o. (1992). Adverse events following mass ivermectin therapy for onchocerciasis. Trans. R. Soc. Trap. Med. Hyg. 86, 284-286. Chippaux, J. P., Gardon-Wendel, N., Gardon, J., and Emould, J. C. (1993). Absence of any adverse effects of inadvertent ivermectin treatment during pregnancy. Trans. R. Soc. Trap. Med. Hyg. 87, 118. Chiu, S. H., Sestokas, E., Taub, R., Buhs, R. P., Gleen, M., Sestokas, R., Vandenheuval, W. J., Arison, B. H., and Jacob, T. A. (1986). Metabolic disposition of ivermectin in tissues of cattle, sheep, and rats. Drug Metab. Dispos. 14, 590-600. Collins, R. c., Gonzales-Peralta, c., Castro, J., Zea-Flores, G., Cupp, M. S., Richards, F 0., Jr., and Cupp, E. W. (1992). Ivermectin: Reduction in prevalence and infection intensity of Onchocerca volvulus following biannual treatments in five Guatemalan communities. Am. J. Trop. Med. Hyg. 47, 156-169. Cukierski, M. A (1995). "Exploratory study of P-glycoprotein development in Rat Fetuses and Pups:' Unpublished report, Merck Project Number TT #94-739-0, Merck & Co., West Point, PA. Descoteaux, S., Ayala, P., Orozco, E., and Samuelson, J. (1992). Primary sequences of two P-glycoprotein genes of Entamoeba histolytica. Mol. Biochem. Parasitol. 54, 201-212. De Sole, G., Awadzi, K., Remme, J., Dadzie, K. Y., Giese, J., Karam, M. FM., and Opuku, N. O. (1989a). A community trial of ivermectin in the onchocerciasis focus of Asubende, Ghana. H. Adverse reactions. Trap. Med. Parasitol. 40, 375-382. De Sole, G., Remme, J., Awadzi, K., Accorsi, S., Alley, E. S., Ba, 0., Dadzie, K. Y., Giese, J., Karam, M., and Keita, F M. (1989b). Adverse reactions after large-scale treatment of onchocerciasis with ivermectin: Combined results from eight community trials. Bull. World Health Org. 67,707719. Didier, A., and Loor, F (1996). The abamectin derivative ivermectin is a potent P-glycoprotein inhibitor. Anti-Cancer Drugs 7,745-751. Doumbo, 0., Soula, G., Kodio, B., and Perrenoud, M. (1992). Invermectine et Grossesses en Traitement de Masse au Mali. Bull. Soc. Pathol. Exp. 88, 247-251. Eldefawi, A. T., and Eldefawi, M. E. (1987). Receptors for g-aminobutyric acid and voltage-dependent chloride channels as targets for drugs and toxicants. FASEB J. 1, 262-271. Endicott, J. A., and Ling, V. (1989). The biochemistry of P-glycoprotein mediated multidrug resistance. Annu. Rev. Biochem. 58, 137-171. FAOIWHO (1991). "Pharmaceuticals: Ivermectin." Available at http://www. inchem.org/documents/pims/pharmlivermect.htm. FAOIWHO (1992). "Toxicological Evaluation of Certain Veterinary Drug Residues in Food." Report of the 36th meeting of the Joint FAOIWHO Expert Committee on Food Additives (JECFA), WHO Food Additive Series 27, pp. 10-18. FAOIWHO (1993). "Toxicological Evaluation of Certain Veterinary Drug Residues in Food." Report of the 40th Meeting of the Joint FAOIWHO Expert Committee on Food Additives (JECFA), WHO Food Additive Series 31, pp. 23-36. FAOIWHO (1994). Pesticide Residues in Food-I 994. "Report of the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and the WHO Expert Group on Pesticide Residues:' FAO Plant Production and Protection Paper 127, pp. 15-17.
FAOIWHO (1997). Pesticide Residues in Food-1997. "Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and the WHO Expert Group on Pesticide Residues." September 22October I, 1997, pp. 22-34. Fisher, G. A., and Sikic, B. I. (1995). Clinical studies with modulators of multidrug resistance. Drug Resist. Clin. Oncol. Hematol. 9, 363-382. Fisher, M. H., and Mrozik, H. (1989). Chemistry. In "Ivermectin and Abamectin" (w. c. Campbell, Ed.), pp. 1-23. Springer-VerJag. New York. Gardon, J., Gardon-Wendel, N., Ngangue, D., Kamgno, J., Chippaux, J. P., and Boussinesq, M. (1997). Serious reactions after mass treatment of onchocerciasis with ivermectin in an area endemic for Loa loa infection. Lancet 350, 18-22. Gottesman, M. M., and Pastan, I. (1993). Biochemistry of multidrug resistance mediated by the multidrug transporter. Annu. Rev. Biochem. 62, 385-427. Greene, B. M. (1991). Expert report on the safety of ivermectin. In "Ivermectin-Report to JECFA," Vol. I. Unpublished Report, MSD Research Laboratories, Lauterbach, Germany. Greene, B. M., Brown, K. R., and Taylor, H. R. (1989). Use of ivermectin in humans. In "Ivermectin and Abamectin" (w. C. Campbell, Ed.), pp. 311323. Springer-VerJag, New York. Jette, L., Murphy, G. F, Lec1erc, J. M., and Beliveau, R. (1995). Interaction of drugs with P-glycoprotein in brain capillaries. Biochem. Pharmacol. 50, 1701-1709. Juliano, R. L., and Ling, V. (1976). A surface glycoprotein modulating drug permeability in Chinese hamster ovary cell mutants. Biochim. Biophys. Acta 455, 152-162. Kurelec, B., and Pivcevic, B. (1991). Evidence for a multi-xenobiotic resistance mechanism in the mussel Mytilus gallopravincialis. Aquat. Toxicol. 19,291-302. Lankas, G. R., and Gordon, L. R. (1989). Toxicology. In "Ivermectin and Abamectin" (w. C. Campbell, Ed.), pp. 89-112. Springer-VerJag, New York. Lankas, G. R., Cartwright, M. E., and Umbenhauer, D. (1997). P-glycoprotein deficiency in a subpopulation of CF-I mice enhances avermectin-induced neurotoxicity. Toxicol. Appl. Pharmacol. 143, 357-365. Lankas, G. R., Minsker, D. H., and Robertson, R. T. (1989). Effects of ivermcctin on reproduction and neonatal toxicity in rats. Food Chem. Toxicol. 27,523-529. Lanning, C. L., Fine, R. L., Corcoran, J. J., Ayad, H. A, Rose, R. L., and AbouDonia, M. B. (1996). Tobacco budworm P-glycoprotein: Biochemical characterization and its involvement in pesticide resistance. Biochim. Biophys. Acta 1291, 155-162. Lincke, C. R. I., van Groenigen, M., and Borst, P. (1992). The P-glycoprotein gene family of Caenorhabditis elegans: Cloning and characterization of genomic and complementary DNA sequences. J. Mol. BioI. 228,701-711. MacFarland, A, Abramovich, D. R., Ewen, S. W. B., and Pearson, C. K. (1994). Stage-specific distribution of P-glycoprotein in first-trimester and full-term human placenta Histochem. 1. 26,417-423. Mauad, T. H, van Nieuwkerk, C. M. J., Dingemans, K. P., Smit, J. J. M., van den Bergh Weerman, M. A., Verkruisen, R. P., Groen, A. K., Oude Elferink, R. P. J., van der Valk, M. A, Borst, P., and Offerhaus, G. J. A. (1994). Mice with homozygous disruption of the mdr2 P-glycoprotein gene: A novel animal model for studies of nonsuppurative inflammatory cholangitis and hepatocarcino-genesis. Am. J. Pathol. 145, 1237-1245. McGrath, J.P., and Varshavsky, A. (1989). The yeast STE6 gene encodes a homologue of the mammalian multidrug resistance P-glycoprotein. Nature 340,400-404. Nakamura, Y., Ikeda, S.-I., Furukawa, T., Sumizawa, T., Tani, A., Akiyama, S.-I., Nagata, Y. (1997). Function of P-glycoprotein expressed in placenta and mole. Biochem. Biophys. Res. Comun. 235, 849-853. Ogbuokiri, J. E., Ozumba, B. C., and Okonkwo, P. O. (1993). Ivermectin levels in human breast milk. Eur. J. Clin. Pharmacol. 45,389-390. Ogunba, R. O. and Gemade, F I. I. (1992). Preliminary observations on the distribution of ivermectin in Nigeria for control of river blindness. Ann. Trop. Med. Parasitol. 86,649-655.
References
Pacque, M" Munoz, B., Greene, B. M., and Taylor, H. R. (1991). Communitybased treatment of onchocerciasis with ivermectin: Safety, efficacy, and acceptability of yearly treatment. J. Infect. Dis. 163,381-385. Pacque, M., Munoz, B., Poetschke, G., Foose, 1., Greene, B. M., and Taylor, H. R. (1990). Pregnancy outcome after inadvertent ivermectin treatment during community-based distribution. Lancet 338, 486-489. Saunders, N. R. (1977). Ontogeny of the blood-brain barrier. Exp. Eye Res. Suppl. 523-550. Schinkel, A. H., Smit, 1. 1. M., van Tellingen, 0., Beijnen, 1. H., Wagennar, E., van Deemter, L., Mol, C. A. A. M., van der Valk, M. A., Robanus-Maandag, E. c., te Riele, H. P. 1., Berns, A. 1. M., and Borst, P. (1994). Disruption of the mouse mdrla P-glycoprotein gene leads to a deficiency in the bloodbrain barrier and to increased sensitivity of drugs. Cell 77, 491-502. Schinkel, A. H., Wagennar, E., van Deemter, L., Mol, C. A. A. M., and Borst, P. (1995). Absence of the mdr1a P-glycoprotein in mice affects tissue distribution and pharmacokinetics of dexamethasone, digoxin and cyc1osporin A. J. Clin. Invest. 96, 1698-1705. Scow, R. 0., Chernick, S. S., and Brinley, M. S. (1964). Hyperliperdemia and ketosis in the pregnant rat. Am. 1. Physiol. 206, 796-804. Terao, T., Hisanaga, E., Sai, Y., Tamai, 1., and Tsuji, A. (1996). Active secretion of drugs from the small intestinal epithelium in rats by P-glycoprotein functioning as an absorption barrier. J. Pharm. Pharmaco!' 48, 1083-1089. Thiebaut, F., Tsuruo, T., Hamada, H., Gottesman, M. M., Pastan, 1., and Willingham, M. C. (1987). Cellular localization of the multidrug-resistance gene product P-glycoprotein in normal human tissues. Proc. Natl. Acad. Sci. U.s.A. 84, 7735-7738. Tiirikainen, M., and Krusius, T. (1991). Multidrug resistance. Ann. Med. 23, 509-520. Toomey, B. H., and Epel, D. (1995). A multi-xenobiotic transporter in Urechis caupo embryos: Protection from pesticides? Marine Environ. Res. 39, 299300.
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Turner, M. 1., and Schaeffer, 1. M. (1989). Mode of action of ivermectin. In "Ivermectin and Abamectin" (w. c. Campbell. Ed.), pp. 73-87. SpringerVerlag, New York. Ueda, K., Taguchi, Y., and Morishoma, M. (1997). How does P-glycoprotein recognize its substrate? Cancer BioI. 8, 151-159. Umbenhauer, D. R., Lankas, G. R., Pippert, T. R., Wise, D., Cartwright, M. E., Hall, S. 1., and Beare, C. M. (1997). Identification of a P-glycoproteindeficient subpopulation in the CF-l mouse strain using a restriction fragment length polymorphism. Toxico!. Appl. Pharmacol. 146, 88-94. U.S. Environmental Protection Agency (EPA) (1999a). Avermectin; pesticide tolerance for emergency exemptions: Final rule. Fed. Reg. 64, 1684316850. U.S. Environmental Protection Agency (EPA) (1999b). Emamectin benzoate; pesticide tolerance: Final rule. Fed. Reg. 64,27192-27200. Van Kalken, C. K., Giaccone, G., van der Valk, P., Kuiper, C. M., Hadisaputro, M. M. N., Bosma, S. A. A., Scheper, R. 1., Meijer, C. 1. L. M., and Pineda, H. M. (1992). Multidrug resistance gene (P-glycoprotein) expression in the human fetus. Am. J. Patho!. 141,963-1072. Whitworth, 1. A. G., Morgan, D., Maude, G. H., Downham, M .D., and Taylor, D. W. (1991). A community trial ofivermectin for onchocerciasis in Sierra Leone: Adverse reactions after the first five treatments rounds. Trans. R. Soc. Trop. Med. Hyg. 85, 501-505. Williams, M., and Yarbrough, G. G. (1979). Enhancement of the in vitro binding and some of the pharmacological properties of diazepam by a novel antihelmintic agent, avermectin BIb. Eur. J. Pharmacol. 56,1273-1276. Wise, L. D., Lankas, G. R., Umbenhauer, D. R, Pippert, T. R, and Cartwright, M. E. (1997). CF-1 mouse sensitivity to abamectin-induced cleft palate correlates with fetal/placental P-glycoprotein genotype. Teratology 55, 41. Wu, C. T., Budding, M., Griffin, M. S., and Croop, 1. M. (1991). Isolation and characterization of Drosophila multidrug resistance gene homologs. Mol. Cell. Bio!. 11,3940--3948.
CHAPTER
57 Inhibitors and Uncouplers of Mitochondrial Oxidative Phosphorylation Robert M. Hollingworth Michigan State University
57.1 INTRODUCTION TO OXIDATIVE PHOSPHORYLATION: FUNCTIONS AND DYSFUNCTIONS 57.1.1 GENERAL CONCEPTS Oxidative phosphorylation (oxphos) is the primary process by which the energy derived from the catabolism of carbohydrates, fats, and proteins is used to synthesize ATP in virtually every cell of eukaryotic organisms. Since ATP is the universal source of chemical energy in the cell, any events that significantly disrupt its production or enhance its degradation will have widespread, multiple, and possibly severe physiological consequences. It is therefore not surprising that the complex, specialized machinery of oxphos is the target for a large variety of natural and synthetic toxicants, among which are a number of pesticides. Compounds that have a primary action on oxphos have a long history of use to control pests. Today, compounds that disrupt oxphos in the target species are widely used as fungicides and have important uses as insecticides and acaricides. They are of much lower significance as herbicides. Naturally occurring oxphos poisons such as rotenone have been used by native peoples for centuries. Oxphos was also the target of some of the earliest synthetic organic pesticides such as 2,4-dinitro-o-cresol which was in use to control tussock moth caterpillars in the 1890s. In last two decades, a large number of newer pesticides have been discovered that affect oxphos and many are now on market with others in the later stages of development. One problem in covering these new materials is that there is little detailed, publicly available toxicological information available, partly because of their newness and partly because of their safer nature. Many of these newer compounds have very favorable toxicological characteristics and may never generate the intense toxicological scrutiny given to some of their predecessors. On the other hand, many older compounds Handbook of Pesticide Toxicology Volume 2. Agents
in this class with a more extensive toxicological literature have been, or are in the process of being, replaced because they present risks that, by current standards, are unacceptable (e.g., arsenicals, dinitrophenols, and some uses of organotins). In other cases (e.g., rotenone) they have become limited in use because of their low intrinsic activity compared to their modem counterparts. Many of these older compounds were covered in depth in the first edition of this handbook and, because of their declining importance, are given less attention here. Finally, several older compounds affecting mitochondria that continue to be widely used are covered in other chapters of this edition [e.g., pentachlorophenol (Chapter 65), the bipiridyl herbicides such as paraquat (Chapter 70), and the fumigant phosphine and the metal phosphides that generate it (Chapter 86). Pesticides (or their active metabolites) can be divided into several broad classes according to the potency and specificity with which they affect oxphos and related mitochondrial functions:
1. Compounds which are potent disrupters of oxphos in vitro and the consequences of oxphos disruption can plausibly explain many or all of their toxic effects. This chapter focuses primarily on these compounds. 2. Compounds which are of low potency in affecting mitochondria but which may achieve sufficient concentration in vivo at high doses to impact oxphos. There is a continuum between compounds in classes 1 and 2. 3. Compounds for which other specific targets are primary in causing toxicity (e.g., neurotoxic insecticides), but for which oxphos is a possible secondary target which could be involved in some types of adverse responses. 4. Compounds which are generally reactive and have multiple sites of action, one of which is likely to be oxphos. Many older fungicides, and most fumigants, are multisite inhibitors, with a relatively nonspecific mode of action. They
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Pesticides Affecting Oxidative Phosphorylation
probably impact oxphos and respiration as part of their primary actions but also have many other deleterious effects on cell functions. Examples include methyl isothiocyanate and its generators (dazomet and metam-sodium), thiophosgene generators (captan, captafol, folpet), electrophilic alkylating agents (chloropicrin, ethylene dibromide, and methyl bromide), the carbon disulfide generator, sodium tetrathiocarbonate, sulfhydryl reactive compounds (acrolein, chlorothalonil, dithianon, quintozene, and some dithiocarbamate fungicides such as maneb, mancozeb, zineb, ziram), fungicidal cationic surfactants which act as general membrane disruptants (dodine, guazatine, and iminoctadine), and metal salts and derivatives (e.g., copper-, arsenic-, and mercury-containing pesticides). 5. Compounds which have no direct effect on oxphos but which can affect oxphos indirectly through the consequences of their primary action. These compounds include lipid, nucleic acidic, or protein biosynthesis inhibitors, and extramitochondrial generators of reactive oxygen species which impact oxphos functions as well as other cellular events. Generally much less is known about the specific role of mitochondria in the toxic action of the groups 2 through 5 than those in group 1. In particular, data obtained by incubating mitochondria with relatively high concentrations of pesticides in vitro may demonstrate effects on oxphos, but establishing that these have significance in vivo is much more difficult and often is not attempted. Nevertheless such actions may be toxicologically significant and this is considered briefly in Section 57.2. The criterion for inclusion in this chapter is that a compound must be known to have its most important primary toxic effect on oxphos in vertebrates or, if this unclear, the compound has a primary effect on mitochondrial oxphos in target species with a probability of similar action in vertebrates. Most of the sites at which these newer pesticides act are highly conserved across all eukaryotes, and frequently it has been shown that there is little difference in the sensitivity of the target system in a fungus or insect compared to that in a vertebrate. It is not unreasonable therefore to assume initially that these and vital oxphos sites in nontarget species are likely to be important in generating at least some of their toxic effects in non-Karget species also. The data regarding the general properties, uses, and toxicological profiles of individual compounds are drawn from a variety of sources. Many are cited individually as each compound is described. However, two sources contributed to many of the compound descriptions, The Pesticide Manual (Tomlin, 2000) and The Farm Chemicals Handbook (Meister, 2000), and these are acknowledged here as a general resource for these data.
lular energetics and, particularly, the production of ATP by oxphos. This process is now quite well understood. The mitochondrion is bounded by two membranes. The outer one is relatively permeable and allows free exchange of small molecular weight solutes with the external cellular environment. The inner membrane is not readily permeable to ions, is highly specialized, and is unusual in that it contains large amounts of the phosopholipid, cardiolipin, but little cholesterol. The control of its permeability is the key to energy conservation during oxphos. The individual catalytic components responsible for oxphos are located in and span across the inner membrane. Also spanning the inner membrane are a number of proteins which act as translocators for ions and for the precursors and products of oxphos such as tricarboxylic acid cycle substrates, and inorganic phosphate, ADP and ATP. Within the inner membrane is a matrix which contains a number of enzymes responsible for feeding the products of intermediary metabolism of carbohydrates, fats, and proteins to oxphos such as the tricarboxylic acid cycle, fatty acid oxidation, and the enzymes of the urea cycle which is responsible for elimination of wastes from the cell. The mitochondrial DNA which codes for a number of subunits of components of the respiratory chain is also located in this matrix. In addition to oxphos the mitochondrion has a number of other significant cellular functions including an important role in Ca2+ regulation within the cell, thermoregulation in some situations, and as a regulator of apoptosis and cell death. Despite these common roles, mitochondria vary widely in number, form, and activity between different tissues. These differences in functional significance and sensitivity to disruption play an important role in determining which tissues are likely to be injured by exposure to mitochondrial poisons. Information on the structure, functions, and bioenergetics of mitochondria are presented in greater detail in several books (Ernster, 1992; Nicholls and Ferguson, 1992; Scheffier, 1999; Tyler, 1992). 57.1.3 OXIDATIVE PHOSPHORYLATION
Oxidative phosphorylation (Fig. 57.1) consists of two closely coupled processes; electron transport and the phosphorylation
MATRIX Succinate
57.1.2 THE MITOCHONDRION AND THE MACHINERY OF OXIDATIVE PHOSPHORYLATION
Mitochondria are organelles that occur in virtually every eukaryotic cell. Much of the earliest work in understanding mitochondrial functions focused on their central role in cel-
CYTOPLASM
Figure 57.1
Schematic overview of mitochondrial oxidative phosphorylation.
57.1 Introduction to Oxidative Phosphorylation: Functions and Dysfunctions of ADP. In electron transport, the oxidation of intermediates from carbohydrates, fats, and proteins creates NADH and reduced ftavoproteins. These reduced carriers are reoxidized by the transfer of electrons down a chain of redox carriers culminating in the reduction of molecular oxygen and its incorporation into water. As the electrons pass from a higher energy state to a lower one, the free energy change at three points is sufficient to drive the pumping of protons outward across the inner mitochondrial membrane which is proton-impermeable. This creates an energy gradient across the membrane, termed the proton electrochemical gradient (~.uH+) or, when expressed in electrical potential units, the proton-motive force U..,.p). This consists of both an electrical component, the mitochondrial membrane potential (~I¥m) due to the charge separation, and a chemical component (~pH) due to the unequal distribution of protons across the membrane. In mitochondria most of ~p (typically about 200-220 mV) can be attributed to ~ I¥m (typicallyabout 150-180mV) (Nicholls and Budd, 2000). The value of ~pH is usually in the range of 0.5-1.0 pH units. The passage of these protons back across the inner membrane discharges the electrochemical gradient and occurs primarily by passage through channels in the membrane-spanning mitochondrial ATP synthase. The passage of protons through this enzyme drives the phosphorylation of ADP to ATP. The proton gradient is also used to power other transport processes across the membrane such as ionic regulation and amino acid uptake. The profound importance of oxphos can be judged, not only from the severe toxicological effects of compounds that disrupt it, but also from the fact that a normally active human synthesizes (and utilizes) approximately 40 kg of ATP daily. Under conditions of high muscular activity, this rate may increase by 10 to 20 fold. In addition to muscular tissues, the nervous system is also a heavy user of ATP. The human brain, which contributes 2% to the body weight, produces 20% of the ATP, mainly to power ion pumps. Since there are no appreciable other energy storage forms and ATP has a half life in cells of only a few seconds, continual resynthesis is crucial. An interruption of ATP biosynthesis of only a few minutes leads to permanent brain damage. Oxphos is the major, but not the sole, source of ATP synthesis in the cell. Glycolysis provides less than 10% of the ATP under conditions of full oxidation of glucose. However, since glycolysis produces ATP much more rapidly than oxphos, it is the major short term source of energy for skeletal muscle contraction in vertebrates. This cannot continue long since it incurs an "oxygen debt" due to the large amounts of lactic acid produced by glycolysis under anaerobic conditions, which must be reoxidized later by lactate dehydrogenase in well oxygenated tissues. When mitochondria are compromised and cannot transfer electrons to oxygen, this reoxidation cannot occur, glycolysis will slow, and lactate will accumulate in the tissues. Under extreme circumstances this can cause lactic acidosis, a potentially life-threatening condition. Finally it is worth noting that mitochondria tend to congregate in the cell around areas with a high rate of ATP utilization
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which is consumed very rapidly after synthesis. As shown by Aw and Jones (1985) there is a gradient of ATP away from the mitochondrion. In a condition in which ATP synthesis is partially compromised it is reasonable to assume that those functions furthest from the mitochondria will suffer the greatest deficit in ATP availability. Knowledge of the structure and mechanism of the components of oxphos has shown remarkable advances in the last decade and has recently been reviewed by Saraste (1999) and Scheffter (1999). 57.1.3.1 Electron Transport
The mitochondrial electron transport chain is illustrated in Fig. 57.2. The primary source of electrons is NADH which donates them to the chain through mitochondrial complex I (NADH: ubiquinone oxidoreductase). In sequence these electrons pass to complex III (ubiquinol : cytochrome c oxidoreductase) and then to complex IV (cytochrome oxidase) in which molecular oxygen is reduced and protonated to form water. In this process ubiquinone (Q) and cytochrome c act as mobile electron carriers within the inner mitochondrial membrane or the intermembrane space. A second, and significant, route by which electrons feed into the chain is by enzymes such as succinate dehydrogenase which reduce FAD. This passes electrons into the chain through complex 11 (succinate: ubiquinone oxidoreductase). Glycerol-3-phospate dehydrogenase (involved in shuttling extramitochondrial NADH across the impermeable inner membrane) and ,B-hydroxybutyrate dehydrogenase (not shown in Fig. 57.2) are similar FAD-linked enzymes located in the inner membrane with significance in some mitochondria. During the passage of electrons down this chain, protons are pumped outward across the inner membrane by complexes I, Ill, and IV. Thus substrates which yield NADH ultimately activate all three proton pumps whereas those that yield FADH only activate pumps in complexes III and IV, yielding proportionally less energy and ATP.
NADH FMN
~
(Fe-S)n Matrix
~
COUOOH2 ..(;:e-s~ PSST-"""
(COOH. CoO
N-2 Cytoplasm
I Rotenone, Pyridaben etc.
Figure 57.2 Major elements of mitochondrial complex I (NADH : ubiquinone oxidoreductase) and the site of action of inhibitors.
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CHAPTER 57 Pesticides Affecting Oxidative Phosphorylation
Complex I (NADH: Ubiquinone Oxidoreductase) This is the largest and least well understood of the four complexes forming the respiratory chain. In vertebrates it consist of at least 42 polypeptide subunits which are thought to be arranged in an L-shaped configuration (Fig. 57.2). The shorter arm of the "L" extends into the matrix and is the locus of the NADH binding site. The longer arm is integrated into the inner membrane and contains the ubiquinone binding site or sites. Current models propose the presence of two Q binding sites acting in a Q cycle analogous to that established for complex III and invoke a ubi semiquinone radical ion as an intermediate resulting from a one electron transfer to ubiquinone (e.g., see Degli Esposti and Ghelli, 1994). In between these binding sites is a series of electron acceptors and donors including a flavoprotein and several (perhaps seven) iron-sulfur centers whose exact relationship remains to be determined. Four protons are translocated across the membrane for the reduction of each ubiquinone molecule to ubiquinol. The iron-sulfur complex designated N2 is believed to be the final carrier that passes electrons to ubiquinone. This is the region at which most of the known potent inhibitors act, including rotenone and a series of newer acaricide-insecticides such as pyri daben. Complex 11 (Succinate: Ubiquinone Oxidoreductase) In contrast to complex I, complex 11 is relatively simple consisting of only four subunits (Fig. 57.3). Although no high resolution structure for this complex is yet available, the structure of the closely related fumurate reductase of E. coli has recently been resolved at 3.3 angstroms (Iverson et aI., 1999). Despite its apparent simplicity, important features of the structure and mechanism of internal electron transfer of complex 11 remain to be established. Two of the four subunits extend into the matrix. One of these carries the catalytic site for the conversion of succinate to fumarate. The two electrons released in this oxidation are collected by an FAD cofactor and passed on through three iron-sulfur clusters in the second unit. Together, these two units Succinate
Fumarate
~ ~
FAD I
Carboxamides
Matrix
FADH2
FP IP (Fe-S)
Cytoplasm
Figure 57.3 Major elements of mitochondrial complex II (succinate: ubiquinone oxidoreductase) and the site of action of inhibitors.
represent the enzyme succinate dehydrogenase. The other two subunits act as membrane anchors for the succinate dehydrogenase. They share a b-type cytochrome of unknown function. Two quinone binding sites are present, one near the inner face and the other near the outer face of the membrane in an arrangement reminiscent to that of complex Ill, although in the case of complex 11 no protons are pumped outward as occurs in the Q cycle of complex Ill. Complex 11 is the locus of action for a group of fungicides collectively termed carboxamides.
Complex III (Ubiquinol: Cytochrome c Oxidoreductase) Complex III (which is often termed cytochrome c reductase or the bCl complex) consists of 11 peptide subunits in vertebrates. The understanding of its structure has been remarkably advanced by high resolution x-ray crystallography (Iwata et aI., 1998; Xia et aI., 1997), including definition of the binding sites of the (E)-,B-methoxyacrylate inhibitor, myxothiazol, and antimycin A (lwata et aI., 1998). However, some details of electron flow through the complex and the proton pumping mechanism are still unknown. Much of the complex protrudes into the matrix. The electron transfer chain from ubiguinol involves, in sequence, an iron-sulfur complex (named after Rieske), cytochrome b, and cytochrome Cl, arranged as shown diagrammatically in Fig. 57.4. The reaction is completed by the transfer of an electron from cytochrome Cl to the mobile carrier, cytochrome c. During these electron transfers, four protons are transferred from the matrix to the cytoplasmic side of the inner membrane. Two reaction centers for ubiquinone are present. One designated Qi (or QN) is located toward the matrix (inner, negative) face of the inner membrane, while the other (Qo or Qp) is located near the cytoplasmic (outer, positive) face. Together they are postulated to operate in the "Q cycle" in which the electron pair received from ubiquinol is split, one being passed on to the Rieske iron-sulfur center and thence to cytochromes Cl. The other electron is passed back to ubiquinone via cytochrome b yielding ubiquinol. A point of toxicological significance is that during the operations of the Q cycle, ubi semiquinone radical ions (Q-) are created as partial reaction products. By interactions with molecular oxygen, these can act as a source for reactive oxygen species (ROS) such as superoxide, hydroxyl radicals, hydrogen peroxide, and peroxynitrile radicals which may cause lipid peroxidation and other damaging cellular oxidations. Most inhibitors interact with either the Qo site (e.g., myxothiazol) or the Qi site (e.g., antimycin A). This complex is inhibited by a few insecticides, but, more significantly, it is the site of action of some very important new fungicides strobilurins that bind at the myxothiazol site. Complex IV Cytochromes Oxidase The final complex in the respiratory chain is cytochrome C oxidase, otherwise know as cytochrome oxidase. During the action of complex IV, cytochrome C is reoxidized and the electrons from the respiratory chain are used to reduce molecular oxygen resulting in the formation of water. During this process two protons are pumped outward from the matrix for each atom of oxygen reduced. The complex structure of the 13 subunit bovine cytochrome
57.1 Introduction to Oxidative Phosphorylation: Functions and Dysfunctions Matrix
2W
2 QH """
Q,
Center [
Qi
(
@ Q,
(
Center
""
1173
Phosphine is discussed elsewhere in this work (Chapter 86) and complex IV is not considered further in this chapter.
l [
57.1.4 THE SYNTHESIS OF ATP
----",Q
~Q
~ 2Q-' .JI 2QH2
57.1.4.1 Complex V (ATP Synthase)
Cytoplasm
Figure 57.4 Major elements of mitochondrial complex III (ubiquinol: cytochrome c oxidoreductase) and the site of action of inhibitors.
oxidase has been determined with high resolution (Tsukihara et al., 1996). It contains two hemes (cytochrome a and cytochrome a3), two copper centers, and two centers in which magnesium and zinc respectively are coordinated and which are probably involved in stabilizing the complex rather than in its redox reactions. The terminal cytochrome a3 operates in conjunction with one of the copper atoms to cleave molecular oxygen. A point of toxicological significance (Scheffler, 1999) is that during the reduction of oxygen by complex IV, reactive partially oxidized species such as the superoxide radical are formed. Leakage of such ROS from complex IV could contribute to oxidative stress in the cell, although leakage of electrons from higher potential sites earlier in the respiratory chain is perhaps a more general source of ROS. Much more detail might be given on the intricate operations of this oxidase, but since it is not known to be an important site of action for current pesticides, the reader is referred to Scheffler (1999) for additional discussion. Two exceptions to this generalization regarding lack of pesticidal significance are hydrogen cyanide, which is still used as a fumigant, and a second fumigant, phosphine (generated from such precursors as aluminum phosphide), which is also known to inhibit cytochrome oxidase.
ADP. -+-*x''J I Pi
ATP
Cytoplasm
Figure 57.5 Major elements of mitochondrial complex V (ATP synthase) and the site of action of inhibitors.
The linked processes of discharge of the electrochemical gradient and the synthesis of ATP is carried out by complex V (ATP synthase). This remarkable structure (considerably simplified in Fig. 57.5) has been shown to be a molecular-sized motor consisting of three major components (Abrahams et al., 1994; Boyer, 1997; Stock et al., 1999; see Scheffter, 1999 and Saraste, 1999 for overviews). A rotor system consisting of9-12 "e" subunits is embedded in the inner membrane. Although the mechanism of operation of the rotor system is still debated, it is likely that protons pass through the rotor and at the same time cause it to spin at rates up to 100 or more revolutions per second. This membrane-located portion of the complex is the Fo component of the ATP synthase. Attached to the rotor is an axle (y subunit) which is eccentric or bent and which contacts the third component, a head consisting of six subunits (alternating ex and fJ) which is attached to the inner side of the membrane (not shown in diagram). As the rotor spins and turns the axle, the axle alternately squeezes and relaxes the fJ units causing a configurational change that drives the combination of ADP and inorganic phosphate to synthesize ATP. As the fJ unit relaxes, its binding site opens to accept ADP and Pi. As the unit is squeezed by the rotating axle, ATP is expelled. Thus three ATP molecules are synthesized for each full rotation of the axle. The head portion of the complex is the FI component of the synthase. At least three, but perhaps four, protons are needed to drive the synthesis of one ATP molecule, paralleling the stoichiometry of three to four" e" subunits for each ATP catalytic site. Since the ATP synthase machinery (as with the electron transport chain) is reversible, it can hydrolyze ATP back to ADP under appropriate circumstances, at the same time pumping protons back out of the matrix. When operating in this mode it is often referred to as the mitochondrial Mg-dependent ATPase or FIFa-ATPase. The isolated FI portion of the complex can also act as an ATPase but without the capability to translocate protons. Several groups of chemicals inhibit this machinery, primarily by interference with the rotary mechanism (Fa component). These include antibiotics such as oligomycin, carbodiimides or carbodiimide generators such as the pesticide diafenthiuron, and the organotin biocides. The existence of this proton circuit comprising the pumping of protons out of the matrix by the respiratory chain and their reentry through the ATP synthase means that in healthy mitochondria with an intact inner membrane there is a close coupling between the consumption of ATP and the rate of respiration. As ATP is consumed, ADP builds up, the rate of ATP synthase rises, protons reenter the matrix, the value of the electrochemical gradient tends to decline, and electron flow through
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Pesticides Affecting Oxidative Phosphorylation
the respiratory chain increases to maintain it. Substrate oxidation and ADP phosphorylation are therefore coupled together and respond rapidly and efficiently to the varying energy demands of the cell.
57.1.5 SIGNIFICANCE OF QUINONE BINDING SITES FOR PESTICIDE ACTION It is notable that most of the pesticides that inhibit mitochondrial electron transport do so by binding in or around the ubiquinone binding sites of complexes I, 11, or Ill. Complex TV, which contains no quinone binding site, is inhibited only by compounds which complex the metal redox centers such as cyanide or azide. The nature of these quinone binding sites and their roles as targets for pesticides and other inhibitors have been reviewed by Rich (1996), Rich and Fisher (1999), and Berry et al. (1999). The abounding variety of natural and synthetic inhibitors that act on these site in complexes I and III indicates that structural requirements for inhibitors are not highly restrictive. As pointed out by Degli Esposti (1998) and Miyoshi (1998), the main requirements for many inhibitors of complex I are a cyclic head attached to a lipophilic chain that approximates the structure of ubiquinone.
Cytoplasm
Matrix
+ +
A·-<········ ·········A-
w~( + +
~W HA
.:> HA
Electron Transport Chain
OH"
OH"
57.1.6 MECHANISMS BY WHICH CHEMICALS DISRUPT OXPHOS Pesticides may act in several different ways to disrupt oxphos and hence cause damage to cells and tissues. First, they may inhibit the operations of the electron transport chain and the pumping of protons across the inner membrane so that the electrochemical potential gradient is not maintained. Second, they may prevent this gradient from being coupled to the synthesis of ATP. This uncoupling action generally occurs through by a compound's ability to increase the permeability of the inner membrane to protons, or other ions, and to discharge the energy stored in the gradient wastefully. This mechanism is shown in its simplest form in Fig. 57.6, where a lipophilic weak acid shuttles protons across the membrane from the outer side to the inner side, driven only by the simple physicochemical processes of association at the lower pH on the outer side of the membrane, dissociation at the more alkaline inner face, and diffusion along its chemical and electrical potential gradients. The net result is for such a compound to act as a shuttle which continually and rapidly transports protons back across the membrane and thereby discharges the electrochemical gradient formed by the activity of the electron transport chain. The third mechanism for interference with oxphos is to block the machinery of the ATP synthase and bring it to a halt. A fourth mechanism by which pesticides may interfere with mitochondrial functions and lead to tissue injury is by diverting electrons from the electron transport chain which wastes energy unproductively and can create ROS such as superoxide anion radicals and hydroxyl radicals leading to oxidative damage to the mitochondrion and other cellular constituents in the process
Figure 57.6 Basic mechanism of action of protonophoric uncouplers (HA) and suggested mechanism of uncoupling by triorganotin pesticides (R3SnOH).
often termed "oxidative stress" (e.g., see Kowaltkowski and Vercesi, 1999). Oxphos can also be impacted by compounds that inhibit the transport mechanisms that convey ATP precursors (ADP and inorganic phosphate) into the matrix and ATP itself out into the cytoplasm. Finally, arsenical compounds that yield arsenate ions in vivo have a special mechanism termed arsenolysis that prevents ATP biosynthesis. The first three of these mechanisms by which pesticides disrupt oxidative phosphorylation have somewhat different effects on cellular energetics and integrity depending on the degree to which the ATPase activity of complex V is stimulated, and whether the mitochondrial membrane potential (~Wm) can be preserved. These ideas are well reviewed by Nicholls and Budd (2000) and Wallace and Starkov (2000).
57.1.6.1 Uncoupling of Ox phos Uncouplers, generally acting as protonophores, discharge the electrochemical gradient which prevents further synthesis of ATP and the operation of other processes linked to this energy gradient such as ion pumping and the production of ROS. In addition the loss of the gradient also causes complex V (ATP synthase) to run in reverse as an ATPase which rapidly eliminates residual ATP from the cell. Even though glycolysis may continue to provide some ATP in the uncoupled cell, this too is
57.1 Introduction to Oxidative Phosphorylation: Functions and Dysfunctions
subject to the rapid ATPase activity. In almost all cells this represents an energy catastrophe and the consequences are rapid and drastic. 57.1.6.2 Inhibition of Electron Transport Whether acting at complexes I, 11, Ill, or IV, electron transport inhibitors have essentially the same effect in that electron flow through the chain is prevented. Inhibition of the respiratory chain reduces or eliminates its ability to pump protons and to maintain Llp,H+, ATP synthesis, and related reactions. Although both complexes I and 11 can independently reduce ubiquinone, the inhibition of either complex alone eventually blocks the respiratory chain through feedback effects on the citric acid cycle which provides NADH for complex I and succinate for complex 11. As Llp,H+ falls, the action of complex V tends to reverse which destroys extramitochondrial ATP, but in so doing pumps protons back out from the matrix side and so tends to maintain the Llp,H+ value close to that of uninhibited mitochondria. However, in contrast to uncouplers, the permeability of the inner membrane to protons is not increased, so that the protons now being pumped from the matrix cannot readily return and ATP is eliminated from the cell more slowly than is the case with uncouplers. Glycolysis may then, at least in the short term, be able to maintain a level of ATP in the cell that prevents catastrophic consequences depending on the relative kinetics of ATP synthesis and destruction. In the longer term, ATP synthesis may be unsustainable and again drastic effects on cellular integrity result. Since there is reserve capacity in some elements of the respiratory chain, a fairly high percentage inhibition of activity may be necessary before electron transport is limited [e.g., in brain mitochondria, complex I activity can be inhibited by 70 to 75% before it becomes the limiting factor in electron transport through the chain of respiratory carriers (Davey and Clark, 1996), and in human osteosarcoma-derived cells, it can be inhibited by 35-40% before respiration declines (Barrientos and Moraes, 1999)].
1175
57.1.6.4 Redox Cycling and Generation ofROS An additional group of compounds that can severely impact oxphos and mitochondrial integrity are those which have redox potentials within the span of that of the electron transport chain. In the simplest mode this allows them to accept electrons from carriers in the chain (e.g., complex I or complex Ill), thus diverting electron flow from the useful conservation of energy. Such alternative acceptors are likely to be particularly dangerous if they can be reoxidized efficiently, for example, by reaction with oxygen or tissue thiols, or by returning the electrons to the electron transport chain at a point of lower redox potential (e.g., complex IV). In either case this establishes a futile redox cycle in which the compound continually diverts electrons from the chain and bypasses some or all of the energy conservation sites (Fig. 57.7; compound). In the case ofthose compound that are reoxidized by oxygen or tissue thiols, the additional production of reactive free radicals, particularly the superoxide anion radical and other ROS, presents a second threat to the integrity and efficiency of oxphos and general cellular integrity (compounds B and C). Depending on the kinetics of their reduction and reoxidation, they may greatly increase oxygen consumption and discharge the electrochemical gradient and generate heat. This resembles the effects of uncouplers, since the respiratory chain continues to accept electrons from NADH, but this is not coupled to the generation of ATP. Few pesticides are known to be capable of this type of mitochondrial toxicity, but paraquat and related herbicides (see Chapter 70) are highly efficient electron acceptors which generate large amounts of ROS as a primary mechanism of action. Naphthoquinones and nitrosamines which can be formed as pesticide metabolites are also known to be capable of redox cycling. The acaricide acequinocyl (Section 57.5.2.2) is a naphthoquinone derivative, but it is has not been reported to be involved in redox cycling reactions with mitochondria. Several other pesticides (e.g., anthraquinone, dithianon, and quinoclamine) are naphthoquinones and might be involved in such cycling. If it were to occur, this would severely impact oxidative phosphorylation and mitochondrial functions in a
57.1.6.3 Inhibition of ATP Synthase In addition to preventing the synthesis of ATP from the electrochemical gradient, inhibitors of ATP synthase also block the ATPase activity of this complex. This allows glycolysis to continue to provide ATP to the cell. Since electron transport is not eliminated and LlP,H+ is not discharged (it may in fact be somewhat increased), the mitochondrial membrane potential is maintained so that other processes driven by this potential such as ion transport across the inner membrane and the generation of ROS continue unabated. The length of time for which the cell can maintain its integrity under these circumstances will depend on the balance between its capability to produce ATP by glycolysis and its rate of ATP consumption. Different cell types vary considerably in this regard (Nicholls and Budd, 2000), but even those with a vigorous glycolytic capability are likely to fail eventually in the face of continued ATP synthase inhibition.
-
A(ox)
A(red)
,J®
:::x:\~
CoQ
~~(C(OX)XOi C(red)
j/ Cyl
A(red)
~
A(ox)
C
O
/
2
_
'\0. e
O2
H2 0
Figure 57.7 Diagrammatic representation ofredox cycling and the generation of reactive oxygen species by the electron transport chain.
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CHAPTER 57
Pesticides Affecting Oxidative Phosphorylation
manner similar to that shown with menadione (Henry et aI., 1995) and adriamycin. This type of action is reviewed in greater detail by Wallace and Starkov (2000). 57.1.6.5 Inhibition of ADP/ATP Exchange
In support of oxphos, the adenine nucleotide transporter (ANT), that conveys ADP into the matrix and ATP outward from it across the impermeable inner membrane, is critical to maintain the synthesis and supply of ATP to the cell. Inhibitors of this transport system are known and have toxic effects. Bongkrekic acid and atractylosides block the ANT, acting specifically at the ATP and ADP binding sites, respectively (Fiore et aI., 1998). Recently, a new fungicide, MON 65500 (N -allyl-4,5dimethyl-2-( trimethylsil yl)thiophene-3-carboxamide) with the proposed common name of silthiofam (Beale et aI., 1998), has been shown to block the ATP transporter in fungal mitochondria as its primary fungicidal action (Joseph-Horne et aI., 2000). It is unclear whether it has the same capability in vertebrate mitochondria, but since this compound shows very low toxicity to many other fungi and to vertebrates (the acute oral LDso for the rat is >5000 mg/kg), it maybe relatively selective for the ANT of the target species. 57.1.6.6 Arsenolysis
Arsenates (lead or calcium) and arsenites were once widely used as pesticides. Their acute toxicity, concerns regarding their carcinogenicity and environmental accumulation, and lack of high efficacy have led to their general demise in this use. Their mechanisms of toxicity are multiple, but one effect, specific to the arsenate ion, involves effects on oxphos. In this case arsenate is able to replace phosphate during conversion of ADP to ATP. The resulting ADP-arsenate ester is unstable and rapidly hydrolyzes to ADP and arsenate (Moore et aI., 1983). Thus, although the mitochondrial electrochemical gradient is discharged and ADP is consumed, no energy conservation occurs. Although loosely referred to as uncoupling, since the consequences parallel those of uncoupling in increasing respiration and the wastage of energy as heat, this is not really accurate. Respiration and phosphorylation remain coupled and effectively transfer energy from substrates to an esterified ADP product. The biochemical perturbation arises from the instability of this product. The stimulation of mitochondrial respiration by arsenate is also distinguishable from that caused by uncouplers since it is blocked by the ATP synthase inhibitor oligomycin (Welle and Slater, 1967). 57.1.7 TOXICOLOGIC CONSEQUENCES OF
DISRUPTING OXIDATIVE PHOSPHORYLATION 57.1.7.1 General Effects
The mitochondrion is an extremely intricate machine. The disruption of oxphos leads to a web of complex interrelated consequences. A few examples of these interlocking processes,
which often involve positive and negative feedback circuits, include the fact that the membrane potential is responsible for driving ATP synthesis but also is directly involved in regulating ion accumulation such a Ca2+, in driving the reduction of NADPH by NADH, and in generating ROS through components of the electron transport chain. These ROS, unless effectively neutralized, are capable of destroying critical components of the electron transport chain (e.g., see Zhang et aI., 1990). The presence of increased levels of ROS also leads to damage to the mitochondrial inner membrane which, in turn, impacts both calcium regulation and the ability of the mitochondrion to maintain the electrochcmical gradient across it. At another level, the NADPH produced by mitachondrial is essential for the reduction of glutathione as an important antagonist of ROS-induced cellular injury. Calcium dysregulation also leads to the increased generation of ROS and finally may trigger events which lead to apoptosis. The many interactions between these events can be hard to fully comprehend but they are reviewed by Jabs (1999) and Nicholls and Budd (2000). Beyond the direct impact on the mitochondrion itself, the failure of ATP biosynthesis leads to many problems for the cell. Fatty acid, nucleic acid, protein, and sterol biosynthesis all require large amounts ATP and are likely to fail rapidly in an energy-compromised cell. Thus, a severe reduction in the availability of energy to the cell can lead to a variety of effects these range from a blockage in cell division, to impaired maintenance of the cellular cytoskeleton and intracellular transport due to effects on ATP-dependent tubulin polymerization, to dysfunction in ionic regulation through effects on ion pumps that are ATPdependent, and beyond. However, a reduction in ATP-dependent cellular functions is not the only deleterious effect of a compromised oxphos. As noted, the electron transport chain has the capability of producing large amounts of ROS, particularly free radicals [see Nicholls and Budd (2000) for an excellent review]. Even when operating in a healthy cell, up to 2% of the oxygen consumed by the respiratory chain is converted to superoxide radicals and thence to other ROS through the "leakage" from the electron transport chain (Boveris, 1984; Kowaltkowski and Vercesi, 1999). The superoxide is produced when electrons are transferred from intermediate sites such as ftavoproteins, Fe-S complexes, or ubiquisemiquinone radicals directly to oxygen rather than passed down the usual carriers to cytochrome oxidase. This electron leakage may arise in several locations including a site upstream of the rotenone inhibition site in complex I (Cadenas et aI., 1977; Hensley et aI., 1998; Herrero and Barja, 1997). A variety of studies have shown that the partial inhibition of complex I increases the level of ROS, free radical production, and subsequent lipid peroxidation in submitochondrial particles (Hasegawa et aI., 1990; Takeshige and Minakami, 1979; Turrens and Boveris, 1980), isolated mitochondria (Hensley et aI., 1998; Pitkanen and Robinson, 1996), and cells in culture (Barrientos and Moraes, 1999). The critical role of ROS in the lethal action of rotenone, a strong complex I inhibitor, in cultured cells was demonstrated by Seaton et al. (1997) who showed that a variety of antioxidants and free
57.1 Introduction to Oxidative Phosphorylation: Functions and Dysfunctions radical scavengers antagonized the ability of rotenone to induce apoptosis in rat PC12 cells. This enhanced oxidative stress induced by partial inhibition of complex I, rather than ATP depletion, has been cited as the critical event in rotenone's ability to cause apoptosis (Seaton et aI., 1998), and in support of the hypothesis that a deficiency in complex I activity (either genetic or chemically induced) is a cause of the oxidative cell death in the dopaminergic tracts of the substantia nigra that underlies Parkinson's disease. Barrientos and Moraes (1999) provide a particularly illuminating and complete analysis and comparison of the two potential causes of cell death caused by partial inhibition of complex I. Using rotenone and other tools, they assessed the relationship between complex I inhibition, respiration, ROS production, lipid peroxidation, the membrane potential, and the occurrence of apoptosis. Their conclusion is that the key event in causing apoptosis is an increased production of ROS rather than a decrease in ATP levels in the cells. Complex III also generates large amounts of ROS under appropriate circumstances. The site of electron transfer here is probably at the Qo site (Fig. 57.4) where the highly reactive ubisemiquinone radical can transfer an electron to molecular oxygen, creating a superoxide anion. Different types of complex III inhibitors have different effects on this ROS generation. Antimycin-A type inhibitors increase superoxide production by blocking the normal pathway for electron transfer from Qo to Qj. On the other hand, myxathiazol (and by extension other structurally related commercial fungicides based on strobilurin) decrease ROS production by preventing the formation of the ubi semiquinone radical. The complex and often contradictory results regarding the generation of ROS by mitochondria and the effects of inhibitors, including pesticides such as rotenone and the carboxamide fungicide, carboxin, on this process are reviewed by McLennan and Degli Esposti (2000). Under different experimental conditions such inhibitors may either increase or decrease the production of ROS. On the other hand, mitochondrial uncouplers, which reduce the mitochondrial membrane potential that drives ROS production, generally decrease the rate of ROS production by mitochondria. An effective complex of antioxidative defenses is needed to prevent significant injury to the cell. When these defenses, such as superoxide dismutase, peroxidases, and tissue thiols, such as GSH, are inactivated or depleted, the general phenomenon of oxidative stress can occur which is characterized by lipid oxidation, membrane destruction, and DNA and protein oxidations (Kowaltkowski and Vercesi, 1999).
57.1.7.2 Apoptosis and Necrosis A wave of recent interest in the machinery of a poptosis (programmed cell death) has done much to shed light on the complex web of events which follows interference with oxidative phoshorylation, and at the same time has highlighted the central role of the mitochondrion in controlling apoptosis as well as in cellular energetics. Apoptosis may be regarded as the orderly destruction and resorption of cells (whether during organ development or due to cellular injury) which occur constantly in
1177
the body and is characterized by blebbing of the plasma membrane, shrinkage of the cell, condensation of the chromatin and digestion of the DNA in a "ladderlike" fashion, cell shrinkage, controlled digestion of the plasma membrane to form small apoptotic bodies, and resorption of these by neigh boring cells without an inflammatory response. This stands in contrast to cellular necrosis (accidental cell death) in which more drastic damage to cellular functions and bioenergetics leads to a catastrophic collapse of all cellular functions, swelling of mitochondria and the cell in general, and rupture of the cell membrane with release of its contents and the concomitant possibility of an inflammatory response. In fact these two processes probably lie at the extreme ends of a continuum of modes of cell death. While necrosis is always a pathological event, apoptosis has important physiological functions such as the shaping of tissues and organs and the removal of injured cells. Inhibition of apoptosis can lead to anatomical malformations, autoimmune disease, and cancer. On the other hand excessive apoptosis removes healthy, vital cells and may be involved in degenerative diseases such as Parkinsonism (Kroemer et aI., 1998) and immunotoxicity. One critical determinant of whether apoptosis or necrosis occurs in an injured cell is the status of cellular energetics as defined by levels of ATP. Key events in apoptosis depend on the availability of a minimal level of ATP. If the ATP level drops too far, apoptosis cannot occur and necrosis is the likely result. Thus compounds that cause apoptosis are often found to cause necrosis when present at higher concentrations. Clearly, compounds which affect ATP levels may lead to either type of cell death and resulting organ injury. A detailed discussion of the biochemistry of apoptosis and the possible role of mitochondria and oxphos lies beyond the scope of this chapter and can be obtained from recent reviews (Brown et aI., 1999; Hengartner, 2000; Jabs, 1999; Kroemer et aI., 1998; Robertson and Orrenius, 2000). However, an outline of what is known is provided here to help to illuminate the critical role of mitochondria and oxidative phosphorylation in this process. Apoptosis may be initiated by humoral factors acting on cellular receptors or by actions within the mitochondrion. In the latter case, the key event is thought to be the opening of a very large pore, the permeability transition pore, spanning both the outer and inner mitochondrial membranes, in a process termed the mitochondrial permeability transition. This immediately causes drastic changes in mitochondrial function including discharge of the potential across the inner membrane, which drives ATP biosynthesis and ionic regulation, and the loss of the soluble cytochrome c and other apoptogenic factors from the mitochondrion. The appearance of cytochrome c in the cytoplasm activates a series of cysteine proteases (caspases) which in turn activate endonucleases that begin to digest the nuclear material. The mitochondrial permeability transition can be trigged by internal calcium dysregulation. This may result from a series of causes (often interlinked) such as loss of the ATP necessary for intracellular calcium regulation, the generation of reactive oxygen species in amounts that damage mitochondrial membranes, and changes in the status of antioxidative mech-
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CHAPTER 57
Pesticides Affecting Oxidative Phosphorylation
anisms (e.g., cellular sulfuydryl groups) which protect against the mitochondrion's constant production of free radicals. The mitochondrial permeability transition is thought to be created by the interaction of at least three proteins which have other functions in mitochondrial activity (Crompton et aI., 1999): the ANT, the voltage-dependent anion channel (also termed porin) which is a large voltage-dependent pore in the outer mitochondrial membrane that upon opening allows solutes up to 5000 Daltons to enter the mitochondrion, and cyclophilin D, a soluble protein in the matrix which may be involved in protein folding through catalysis of cis-trans isomerization of prolyl peptide bonds. The manner in which these come together to form the transition pore and the possible presence of other components remain to be established. Inhibitors of the normal functions of these components (e.g., bongkrekic acid for ANT or cyclosporin A for cyclophilin D) antagonize apoptosis. The pesticides described in this chapter may be involved in the apoptotic process in several ways [e.g., rotenone may exert an anticancer action at high doses by encouraging apoptosis of tumor-initiated cells through its effects on the bioenergetics of the cell (Section 57.3.2.1)]. Organotins cause the apoptosis of thymus cells and consequent loss of cellular immunity, probably by directly triggering the mitochondrial permeability transition (Section 57.6.2.1). More speculatively, diafenthiuron reacts with and changes the characteristics of porin in some species, although it not clear whether this action has any implications for apoptosis (Section 57.6.2.2). Obviously, there is much more to learn about the relationship between oxphos disruption by pesticides and both the mechanisms and toxicological significance of cellular apoptosis and necrosis.
57.2 OXIDATIVE PHOSPHORYLATION AS A TARGET FOR PESTICIDE ACTION AND ITS RELEVANCE FOR TOXICITY 57.2.1 GENERAL CONSIDERATIONS
Somewhere in the range of 40 to 50 pesticide active ingredients, acting either as respiratory inhibitors or uncouplers, have effects on mitochondrial oxphos in vitro at nanomolar to micromolar concentrations. Taken in conjunction with signs of poisoning and other evidence, it seems reasonable to suppose that their acute toxicity and many, though perhaps not all, of their chronic toxic effects arise by these disruption of mitochondrial energetics. These compounds are the main focus of this chapter and are discussed in detail in Sections 57.3 through 57.7. However, they represent only a part of the total spectrum of the interactions of pesticides with oxphos. It is probable that other pesticides interact with oxphos in vertebrates with some specificity but with relatively low potency, and that others interact with multiple cellular targets, among which is oxphos, but they are not specific to this site in their biochemical mechanisms of cellular disruption. Examples of the latter
group are sulfuydryl reactive compounds and general membrane disruptants. Although mitochondria may be involved in their toxicity, other sites may be equally or more important and it is often difficult to determine which are the primary and critical biochemical injuries and which are secondary or noncritical. In the former group (specific for oxphos, but weakly active) are a number of compounds, particularly fungicides and herbicides, which tend to have their pesticidal actions on physiological systems that are not present in animals including photosynthesis, essential amino acid biosynthesis, and cell wall biosynthesis. These often have very low acute toxicities to vertebrates. At oral doses of 1000-5000 mg/kg, and assuming reasonably efficient uptake, internal concentrations in the cells of nontarget species might reach 1 mM or higher. Many pesticides can be shown to perturb mitochondrial functions at such concentrations in vitro. However, only rarely has research been conducted to investigate the occurrence and significance of such low potency mitochondrial actions in the toxicology of relatively safe compounds in vivo. Fairly typical results that show the frequency with which pesticides in general exert such low potency effects on mitochondria are provided in a pair of papers by Yamano and Morita (1993, 1995). They examined the effects of 48 common pesticides of varied structures on respiration in isolated rat liver mitochondria. These compounds were not chosen because of any known ability to interact with mitochondria. Of the 48 compounds, 42% (20) were found to impact oxphos at concentrations of 1 mM or lower. Half of these were uncouplers and half were electron transport inhibitors. However, it is significant to note that only 1 (the obsolete herbicide trichlamide, acting as an uncoupler) was active at a concentration as low as 1 ).lM, and only 1 more (the insecticide amitraz acting as an inhibitor) was active at 10 ).lM. There was no obvious correlation between the ability of these compounds to disrupt oxphos and their toxicities to isolated rat hepatocytes at 1 mM. Thus, although the ability to alter oxphos in vitro is a rather common property of pesticides, this often occurs at relatively high concentrations that mayor may not have toxicological significance in vivo. A similar conclusion can be drawn from a study of the in vitro effect of 47 structurally and functionally varied pesticidal chemicals and metabolites on rat liver mitochondrial energetics by Abo-Khatwa and Hollingworth (1974). Eight of the compounds chosen were already known to be active as uncouplers or inhibitors of oxidative phosphorylation and gave appropriate results. Of the other 39 compounds, only 10 (26%) had appreciable effects on oxidation or phosphorylation at concentrations below 100 ).lM. Of the others, 41 % had threshold activities at 0.1 to 1.0 mm and 33% were inactive even at 1 mM. Thus again, many varied pesticides had the capability to interact with mitochondria, but few showed high potency. Read et al. (1998) have also published data on the action of a 164 chemicals on oxphos functions in beef heart submitochondrial particles (SMPs fragments of the inner mitochondrial membrane). Of these compounds, approximately 50 have pesticidal uses, including all classes of pesticidal activity and a wide range of structural types. Twelve of these 50 are compounds
57.2 Oxidative Phosphorylation as a Target for Pesticide Action and Its Relevance for Toxicity
widely recognized to have potent effects on mitochondrial respiration that can explain most, if not all, of their toxic effects (e.g., dinitrophenols and other phenolic uncouplers, and organotins). All of these known oxphos-perturbing compounds had EC50 values of 3 ppm (roughly 10 I-lM) or less. The most active compounds were rotenone and fentin hydroxide with EC50 values of 10 and 33 nM, respectively. Among the 37 compounds which are not generally recognized to act through oxphos effects, 46% had EC50 values of 100 I-lM or higher, and 32% had EC50 values between 10 and 100 I-lM. The remaining 8 compounds with EC50 values below 10 I-lM were all highly lipophilic neurotoxicants. In general, the potency of these compounds against SMPs was higher than in the studies of Yamano and Morita (1993, 1995) with intact mitochondria, perhaps because penetration barriers are removed during the disruption of mitochondria to form the SMPs. In the study by Read et al. (1998), the correlation of potency in causing oxphos effects with whole organism toxicity (using log transformations) is surprisingly good (e.g., an r2 value of 0.763 was obtained in correlating the EC50 on SMPs with toxicity to fish for 104 varied chemicals, and a value of 0.859 was obtained for a subset of 19 neurotoxic insecticides). However, it is unlikely that these compounds have their toxic action primarily by effects on oxphos rather than the nervous system, More likely, their high lipophilicity bestows the ability to concentrate in lipid bilayer membranes which is a prerequisite for both neurological and mitochondrial actions. 57.2.2 SPECIFIC HERBICIDES AND OXPHOS IN VERTEBRATES
There are a number of examples of compounds that act as herbicides by affecting specific plant functions and that are of unknown mode of action in vertebrates, but which have been shown to interfere with oxphos in vertebrate mitochondria in vitro. These compounds include the chlorophenoxy herbicides (2,4-D, 2,4,5-T, and MCPA) that act in plants as hormone (auxin) mimics and that are weak uncouplers in rat liver mitochondria in vitro (Abo-Khatwa and Hollingworth, 1974; Zychlinski and Zolnierowicz, 1990). At 1-10 mM, 2,4-D decreased ATP, glutathione, and NADH level in rat hepatocytes (Palmeira et aI., 1994a), inhibited complexes 11 and III strongly, and uncoupled mitochondria at 0.5 mM in isolated rat liver mitochondria. However, 2,4-D is 1OOO-fold less potent as an uncoupler than the dinitrophenolic compound dinoseb (Palmeira et aI., 1994b). Nevertheless, signs of poisoning in humans include several such as hyperventilation, tachycardia, and pyrexia and sweating that are typically seen with uncouplers (Flanagan et aI., 1990). The phenylcarbamate herbicide, terbutol, and its N -demethyl metabolite were toxic to rat hepatocytes at 1 mM in vitro. Cytotoxicity was accompanied by loss of ATP and free sulfhydryl groups, including glutathione. These herbicides also impaired the respiration of rat liver mitochondria in vitro (Suzuki et aI., 1997). However, the relationship of these observation to the in vivo toxicity of terbutol is unknown.
1179
The pyridazinone herbicide chloridazon is a photosynthetic electron transport inihibitor in plants (Tomlin, 2000). It has a low acute toxicity to vertebrates (acute oral LD50 of at least 800 mg/kg in rats) and the symptoms of poisoning bear a close resemblance to those of a mitochondrial uncoupler, including apathy, dyspnea, hyperventilation, death in clonic convulsions, and very rapid rigor mortis. However, its effects on rat liver mitochondria were complex and hard to interpret involving an inhibition of succinate oxidation, an increase in respiration with glutamate as substrate, and an increase in ATPase activity, but with an increase in the efficiency of respiratory control with both substrates (Guzy et aI., 2000; Mlynarcikova et aI., 1999). Mitochondrial disruption may therefore play a central role in its acute toxicity to vertebrates, but the mechanism is unclear. In each of these cases, even though the herbicides are only weakly active on oxphos, this may be the major biochemical target in vertebrates, but clear proof is lacking. 57.2.3 SECONDARY EFFECTS OF COMPOUNDS ACTING ON OTHER TARGETS IN VERTEBRATES: NEUROTOXICANTS
There are many reports in the literature of pesticides that are known to be potent neurotoxic ants in vertebrates also affecting mitochondrial functions in vitro. This usually occurs at relatively high concentrations compared to those needed to affect ion channels enzyme and receptors in the vertebrate nervous system. These reports include several organophosphate anticholinesterases (Carlson and Ehrich, 1999; Holmuhamedov et aI., 1996; Moreno and Madeira, 1990; Sitkiewicz et aI., 1980), organochlorines such as DDT (Moreno and Madeira, 1991), and pyrethroids (Gassner et aI., 1997; Read et aI., 1998) which affect sodium channels, the DDT metabolite DDE which is not neurotoxic (Ferreira et aI., 1997), and the organochlorines chlordane, dieldrin, endosulfan, and heptachlor, which act on GABA-gated chloride channels (Kannan et aI., 2000; Meguro et aI., 1990; Mishra and Shukla, 1995). Such effects could well explain the cytotoxicity of these compounds in vitro (e.g., the apoptotic effects found by Kannan et al. with endosulfan in a human leukemic cell line at 10-200 I-lM), but their acute toxicity in vivo is much more likely to result from specific neurological effects at lower concentrations. For example, the formamidine insecticide/acaricides (chlordimeform and amitraz) have been shown to affect oxphos by acting as uncouplers at 10-100 I-lM concentrations in vitro (Abo-Khatwa and Hollingworth, 1973, 1974; Yamano and Morita, 1993). However, it is probable that their acute toxic effects in vivo in vertebrates are related to their effects as aadrenergic agonists (e.g., see Costa et aI., 1989; Hsu et aI., 1988) and local anesthetics (e.g., see Pfister et aI., 1978) rather than to their mitochondrial actions. Another example is provided by the organochlorine insecticide chlordane. This inhibits respiration (states 3 and 4) in isolated rat liver mitochondria at 50-100 I-lM (Ogata et aI., 1989)
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CHAPTER 57
Pesticides Affecting Oxidative Phosphorylation
but quite high doses in vivo (100 mg/kg daily for four days) caused no major changes in functions of liver mitochondria in rats (Ogata and lzushi, 1991). Similarly, the anticholinesterase insecticide parathion at 0.1-1.0 mM caused several effects on oxphos in rat liver mitochondria including inhibition of complex II and ATP synthase and partial inhibition of the Pi transporter (Moreno and Madeira, 1990), and parathion and methyl parathion alter the fluidity of the mitochondrial membrane in vitro at 0.05 mM (Antunes-Madeiraet aI., 1994; Lopes et aI., 1997). However, methyl parathion has been reported to have no effect on liver mitochondria in vivo in rats (Mihara et aI., 1981). Nor did chronic exposure to several structurally related organophospates in rats cause diminished mitochondrial activity and ATP production in the brain (Fukushima et aI., 1997). While it is not possible to rigorously exclude the possibility that some toxic effects of organophosphorus insecticides result from their effects on oxphos rather than from their very potent ability to inhibit acetylcholinesterase, the claim that oxphos effects may play an important role in their toxicity based on such in vitro assays (e.g., see Sitkiewicz et aI., 1980) is quite speculative. On the other hand, endrin, an organochlorine insecticide, acts as a blocker of GABA-gated chloride channels which clearly accounts for many of its acute toxic effects. Endrin induced the formation of reactive oxygen species in isolated rat peritoneal macrophages and in hepatic mitochondria and microsomes at submicromolar concentrations, and it decreased microsomal membrane fluidity in vitro (Bagchi and Stohs, 1993). Interestingly, the same effects were observed in vivo in rats when endrin was given orally at a sublethal dose (4.5 mg/kg) (Bagchi et aI., 1993). Thus it would be a mistake to routinely dismiss the mitochondrial effects of potent neurotoxic ants observed in vitro as likely to be relatively unimportant in vivo. While it certainly does not explain all the results, one underlying reason for the reports of large number of pesticides that affect mitochondrial oxphos at high concentration in vitro may be that many of these pesticides are significantly lipophilic, and, at relatively high concentrations in vitro, can affect mitochondria through the simple disordering of the membrane integrity that is critical for mitochondrial functions and survival. Such effects on membrane structure, fluidity, and integrity have been shown to lead to a changed membrane environment for the components of oxphos and consequent alterations in their activity, membrane leakage, and uncoupling, and to the increased generation of ROS and oxidative stress (e.g., see Antunes-Madeira and Madeira, 1979; Stolze and Nohl, 1994). Examples of studies showing various types of changes in membrane ordering and fluidity by pesticides, generally at quite high concentrations (10 J.LM to 1 mM) compared to their most potent known effect on target receptors or enzymes, include chlorinated hydrocarbons, e.g., Moreno and Madeira (1991) with DDT, Ferreira et al. (1997) with its nontoxic metabolite, DDE, Antunes-Madeira and Madeira (1979) with aldrin, and Bagchi et al. (1993) and Bagchi and Stohs (1993), with en-
drin. Similar studies with organophosphates include malathion (Antunes-Madeira and Madeira, 1979), parathion (AntunesMadeira et aI., 1994), and methyl parathion (Lopes et aI., 1997). Lipophilic organotins such as tributyl- and triphenyltin also are well known to have general membrane disruptive effect (e.g., causing disruption of erythrocyte membranes at concentrations as low as 5 J.LM) (Gray et aI., 1987). A final factor that makes interpretation of some in vitro results with mitochondria difficult is that many discrepancies can be found between different reports of such weakly active compounds on mitochondria. In their survey of a range of pyrethroid insecticides Yamano and Morita (1993, 1995) found no consistent pattern of effects on rat liver mitochondrial respiration and none had an effect at a concentration less than 100 J.LM. On the other hand, Gassner et al. (1997), also using rat liver mitochondria, reported that the pyrethroids permethrin and cyhalothrin inhibit mitochondrial complex I with ICso values near 10 J.LM. Similarly, in their studies with bovine heart submitochondrial particles, Read et al. (1998) found that permethrin was one of the more potent inhibitors, active at less than 1 J.LM. By contrast, permethrin is reported to have no effect on rat liver mitochondria at 1 mM by Yamano and Morita (1993). The numerous contradictions in the published reports of the effects ofDDT on oxphos are reviewed by Moreno and Madeira (1991). It has been described by different authors as either a stimulator or an inhibitor of oxphos and related mitochondrial ATPase activity. Organophosphate insecticides were generally found to have no consistent effect on mitochondrial respiration at 1 mM (Yamano and Morita, 1993, 1995) but other authors have reported varied effects at much lower concentrations. The phenoxy acetic acid herbicide MCPA is reported to act like oligomycin in increasing the mitochondrial membrane potential through inhibition of mitochondrial ATP synthase in rat hepatocytes exposed at 2.5 mM (Camatini et aI., 1996) whereas Zychlinski and Zolnierowicz (1990) concluded that MCPA and related herbicides are weak uncouplers for rat hepatic mitochondria in vitro, an action that would tend to discharge the membrane potential. Variations between the biological preparations and techniques in these studies, and the fact that some compounds have biphasic or multiple effect on oxphos as their concentration is increased, may underlie some of these discrepancies, but others are not easily explained, and considerable caution is appropriate in interpreting the results from any single study, particularly in terms of possible toxic effects in vivo. In closing this section, it is worth noting the report that a formulation of Tordon herbicide (a mixture of picloram and 2,4-D) inhibited mitochondrial complex I in vitro (Pereira et aI., 1994). Subsequent investigation revealed that all the mitochondrial effects were due to a surfactant in the formulation and not to the active pesticidal ingredients either singly or in combination (Oakes and Pollak, 1999). This nicely illustrates the dangers of using commercial formulations in such studies in vitro and the ability of detergents to damage mitochondrial membranes.
57.3 Inhibitors of Complex I
CH,~OCH'O OH
57.3 INHIBITORS OF COMPLEX I 57.3.1 INTRODUCTION Until recently, the only pesticide thought to act primarily through the inhibition of complex I was rotenone. Although this is a familiar compound to toxicologists and biochemists, it currently has only minor uses as an insecticide because of its relatively low level of activity and short duration of action. Its extremely high toxicity to fish underlies its continuing use as a piscicide, but this, too, is a negative factor for its widespread use against insects. In the last decadc, a number of new pesticides that act on complex I with powerful acaricidal activity and some insecticidal actions have been developed and several are now used worldwide. These compounds, fenazaquin, fenpyroximate pyridaben, pyrimidifen, and tebufenpyrad, have the broad structural commonality of being lipophilic nitrogen heterocycles. The properties, mechanism of action, and toxicology of these compounds have been reviewed by Hollingworth and Ahammadsahib (1995). Because these compounds have the capability to inhibit complex I in vertebrates as well as in invertebrates with high potency, they generally possess a higher degree of acute toxicity to mammals than most modem pesticides. They also resemble rotenone in having very high toxicities to aquatic species in most cases. The structure-activity relations of several groups of complex I inhibitors, including rotenone and the lipophilic heterocyclic pesticides, has been reviewed by Miyoshi (1998) with the broad conclusion that necessary structural features for the agrochemical inhibitors are a heterocyclic ring with two nitrogens and a hydrophobic tail structure. Akagi et al. (1996) mode led the three-dimensional conformations of tebufenpyrad, fenpyroximate, and pyridaben (see Fig, 57.8 for structures) and concluded that a common structure featuring a lone electron pair in the heterocyclic ring and a hydrophobic extension with a terminal tert-Bu-substituted phenyl group existed among these compounds. The active conformations were nonplanar with the heterocyclic group and the hydrophobic tail held at about a 90° angle. This is also believed to be the active configuration of rotenone (Miyoshi, 1998). The emergence of these inhibitors has provided new tools to investigate the nature of the rotenone binding domain in complex I and its relationship to the binding sites for other complex I inhibitors and to the coenzyme Q reduction site(s). These results tend to be to confusing and there remains substantial disagreement regarding the number and relationship of binding sites for inhibitors in this region. The reader is referred to reviews by Degli Esposti (1998), Liimmen (1998,1999), Ohnishi et al. (1999), and Okun et al. (1999) for additional details. However, the conclusion seems to be generally accepted that these lipophilic nitrogen heterocycles bind at, or very close to, a high affinity rotenone binding site, which, in turn is located close to the site where coenzyme Q is reduced (see Fig. 57.2). Two (or even three) Q binding sites may be present and it has been proposed that different types of inhibitors bind preferentially to
1181
B
0
C
Rotenon:
H,C
\Q{
0
~~~H
'y Fenazaqu in
Ir\\ ---",0
\'N-{)~O-C(CH,),
CH, Fenpyroxirrale
~'
H~
CI~ CH,CH,
o
CH'CH'
0
Py r i daben
Py r i mi d i fen
C'HS~H'~(CH,), -
Cl
r;' o H
r; 0
Tebufenpyrad
Figure 57.8
9
Ir\\ ~
C'HSP;_H, Cl
CH,
N H
Tol fenpyrad
Pesticides that act as inhibitors of complex L
these different Q binding sites or can otherwise be divided into classes based on the kinetics of their interactions with complex I (Degli Esposti, 1998; Degli Esposti and Ghelli, 1994; Friedrich et al., 1994; Liimmen, 1999) but the specific details vary among these models and in some cases are contradictory. As few as one binding site near the Q reduction site (Liimmen, 1999) to as many as three (Degli Esposti, 1998) have been proposed. The case for two sites corresponding to the putative two quinone binding sites is reviewed by Degli Esposti and Ghelli (1999). These results are analyzed by Liimmen (1999) and the reasons for the apparent disparity in results are discussed. However, it is clear that complex I is indeed complex and the specific binding loci of these inhibitors remain to be established definitively. The culmination of these studies is the identification, using a photoaffinity label derived from the acaricide pyridaben, of the 23 kDa PSST subunit of complex I as the high affinity binding site for this compound and probably also for rotenone and several other complex I inhibitors (Schuler et al., 1999). This subunit is believed to link electron transfer from the terminal iron-sulfur cluster, N2, to the ubiquinone reduction site (Fig. 57.2).
57.3.2 PROPERTIES OF SPECIFIC COMPOUNDS 57.3.2.1 Rotenone (2R,6aS, 12aS)-1 ,2,6,6a, 12, 12a-hexahydro-2-isoprenyl-8,9-dimethoxychromeno[3,4-b ]furo[2,3-h ]chromen-6-one (Fig. 57.8) has CAS Reg. No. 83-79-4.
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CHAPTER 57
Pesticides Affecting Oxidative Phosphorylation
Rotenone occurs in a large number of leguminous plants, but for commercial use, it has primarily been derived from the roots of Derris species (D. elliptica, D. longicarpa, and D. mallaccensis) from Southeast Asia and Lonchocarpus species (L. urucu, L. nicou, and L. utilis) from South America. The commercial material derived from Lonchocarpus is termed cube, barbasco, nekoe, or timbo, while that from Derris is called derris root or tuba root. The root and its extracts contain a variety of compounds related to rotenone which are collectively termed rotenoids. Cube root from Peru, which is currently the only commercial source of rotenone in the United States, may be used as a ground powder (5-20% total rotenoids) or further extracted with organic solvents. Additional purification and the addition of stabilizers to prevent oxidation and microbial contamination may follow. High quality commercial cube resin typically contains 80-90% total rotenoids. In addition to rotenone itself, these rotenoids include deguelin, rotenolone, sumatrol, tephrosin, and toxicarol. They have variable but lesser insecticidal activity than rotenone. In cube, rotenone itself constitutes approximately 40% of these rotenoids with deguelin being the second most common constituent at about 20%. These two components are the most active inhibitors of mitochondrial complex I in cube resin (Fang and Casida, 1998) and are responsible for virtually all of its acute toxicity to insects, fish, mammals and cells in culture (Fang and Casida, 1997; Fang et aI., 1997). A large number of minor components of cube resin have also been identified and their biological potencies assessed (Fang and Casida, 1997, 1999a, 1999b). The composition, origins, and toxicology of rotenoids have been extensively reviewed (Metcalf, 1955; Negherbon, 1959; Fukami and Nakajima, 1971; Ha1ey, 1978; Gosse1in et aI., 1984; Ray, 1991). The reader is referred to these sources for coverage of much of the earlier work on rotenone and its toxicology. The discussion below focuses primarily on more recent results. General Properties Pure natural rotenone consists of colorless dimorphic crystals, m.p. 163°C or 181°C depending on form, v.p. <1 x 10- 3 Pa (20°C), w.s. 0.142 ppm, log P4.16. As described by Miyoshi (1998), rotenone has three chiral centers (Fig. 57.8). The naturally occurring compound has the 6aS, 12aS, 5' R configuration. The three-dimensional structure of this isomer of rotenone, revealed by x-ray crystallographic analysis, shows an "L" shaped molecule, strongly bent at the joining of the Band C rings. This molecular feature is critical for high inhibitory activity against complex I. The detailed structural features of the rotenone molecule necessary to inhibit complex I have been investigated and reviewed by Ueno et al. (1996). The structure-toxicity relationships of rotenoids are reviewed further by Miyoshi (1998). Uses Rotenone has been employed in commercial agriculture since at least 1848 when it was used as an insecticide in plantations in British Malaya, but its insecticidal properties were known much earlier to native people in several parts of the world (Haley, 1978). It was introduced as a commercial insecticide in the United States at the turn of the 20th century
and was once used widely. It now has very limited significance as an insecticide and acaricide in commercial agriculture, although it remains useful in organic production systems, in home and garden products, and for ectoparasite control (e.g., fleas and ticks) on pets. Rotenone is often combined with other insecticides such as pyrethrins or with synergists such as piperonyl butoxide which inhibit microsomal oxidative reactions and thereby enhance its rather low insecticidal activity and broaden it spectrum of action. Rotenone-containing materials have long been used by indigenous people as fish poisons. Due to its high toxicity to fish, rotenone continues to be used as a piscicide to remove unwanted species from bodies of water and to conduct fish surveys. In the 10 years from 1988 to 1997, 94,739 kg of rotenone were used in fish control projects in North America (Finlayson et aI., 2000). These uses are increasingly controversial because of concerns regarding the breadth of their environmental impact, possible human exposure through drinking water or consumption of treated fish, and even animal rights issues (Finlayson et aI., 2000). Trade names include Chem Fish, Fish-tox, Noxfish, Peru Cube Powder, Prentox, Rotenone Powder, and Synpren-Fish. Toxicology Profile The primary data sources are Haley (1978), Gosselin et al. (1984), Ray (1991), and Cal EPA (1997). Most of the studies described used technical (85-90% pure) or analytical grade rotenone (95-99% pure). In a few cases the purity is unstated.
Acute Toxicity The acute toxicity of rotenone to mammals is unusually variable (Ray, 1991; Table 57.1) and is strongly dependent on the formulation and route of exposure. When formulated as an emulsifiable concentrate, the acute toxicity may be very high (EPA toxicity class I), while other formulations such as dusts are much less toxic (EPA class Ill). In rats and dogs, the dust formulation is more toxic by inhalation than by oral dosing. Another significant factor governing both the oral and the respiratory toxicity is the particle size with smaller particles being more toxic. The critical factor underlying these variations in acute toxicity is probably the effect of formulation and particle size on the rate of uptake because rotenone has a very high intrinsic toxicity to vertebrates (e.g., the LDso after intraperitoneal administration to mice is only 1.6 mg/kg) (Haley, 1978). Acute poisoning symptoms include emesis and, at higher doses, increased respiration, muscle tremors, and, ultimately, respiratory and behavioral depression, convulsions, and respiratory failure. Severe hypo glycemia is observed in dogs and rabbits. IrritationlSensitization Local effects caused by exposure to rotenone include mild skin irritation and stronger eye irritation which may be due to components in the plant extract other than rotenone itself (Gosselin et aI., 1984; Haley, 1978). Subchronic Toxicity In a 6 month oral dosing study in dogs, the lowest observed effect level (LOEL) was found at 2 mg/kg/day with endpoints including hematological changes,
1183
57.3 Inhibitors of Complex I Table 57.1 Acute Toxicity of Pesticides Acting as Inhibitors of Mitochondrial Complex I to Selected Nontarget Species
Compound
Rat (M; F)
Mouse(M)
Quail
LCSO (ppb) (24-96 h)
LCso (mg/I)
LDSO (mg/kg) Acute oral
Acute dermal
Inhalation
Duck
Rat or rabbit
Rat
Fisha
Rotenone
132-1500
350
1680b
>2000
>5000
0.02
1.9-75
Fenazaquin
134; 138
2449
1747
>2000
>5000
1.9
3.8-34
Fenpyroximate
480;245
520
>2000
>2000
>2000
0.33
>2250
>2500
>2000
0.64
445
>2000
>2000
>2000
1l00; 570
424
Pyrimidifen
148; 115
245
Tebufenpyrad
595;997
224
Pyridaben
>2000
79-290 1.1-8.3
Daphnia
26 4.1 204 0.59
93 2.66
18-95
48
a Range
of values from several species, most commonly induding the rainbow trout and bluegill. bPheasant.
emesis, diarrhea, and decreased body weight. Gastrointestinal lesions were found at 10 mg/kg/day. Chronic Toxicity In a 2-year dietary study in rats, lowered body weight was seen with rotenone at 37.5 ppm and an increased incidence of adrenal gland angiectisis and hemorrhage was noted at 75 ppm (3.8 mg/kg/day). Carcinogenicity The evidence regarding rotenone's potential carcinogenicity is conflicting and convoluted. In earlier work, a dietary study with Osborne-Mendel rats given cube powder at concentrations from 50 to 1000 ppm showed no increase in tumors (Hansen et aI., 1965). A similar result was obtained in a 28-month feeding study with dogs. The cube powder contained 5.8% rotenone and 12% other extractives with 82% inert material. In a lifetime dietary study with tubatoxin (90% rotenone) in two strains of mice, Innes et al. (1969) found no adverse effects, but only a single, low dose (3 ppm) was used. Neither of these studies would be considered adequate to assess carcinogenicity by current standards. In a series of reports that caused particular concern in the regulatory community in the United States, Gosalves and his co-workers (see Gosalvez, 1983 for a review), claimed that rotenone, given either intraperitoneally (1.7 mg/kg/day for 42 days) or orally (13.5 mg/kg/day for 45 days), greatly increased the number of fibroadenomas of the mammary gland in female albino and Wistar rats. Only one adenocarcinoma was observed, but some of the tumors were transplantable. The fibroadenomas were encapsulated and contained numerous malformed mitochondria which were deficient in respiratory control and oxphos. Gosalves subsequently suggested that these positive results, at least in part, were due to the tumor-enhancing effect of deficiencies in his diets, particularly a low level of riboflavin. These results precipitated extensive additional research. Freudenthal et al. (1981) carried out several studies. Rotenone was administered subchronically (42 days) at 1.7 or 3.0 mg/kg/ day either by gavage to Wistar rats or intraperitoneally to Sprague-Dawley rats and the animals were maintained for 14 to 18 months. The intraperitoneal study paralleled the conditions used by Gosalves et at. but in a different strain of rats. A
dietary study with Syrian golden hamsters, lasting 18 months with rotenone concentrations from 125 to 1000 ppm, was also conducted. Gross and microscopical examination of the tissues revealed no significant increases in neoplasms of any kind in the rat studies, but an increase in adrenal cortical adenomas was recorded in females in the gavage study. Three adrenal cortical carcinomas were also observed in the hamsters at the highest dosage, but it was concluded that it was questionable whether these were related to the rotenone treatment. No other increases in neoplastic lesions were seen. Evaluation of the hamster study was made more difficult by high mortality in all treatments, possibly due to E. coli infection. Subsequently, Abdo et at. (1988) reported on two National Toxicology Program lifetime dietary studies, one in F3441N rats at rotenone concentrations of 38 and 75 ppm and the other in B6C3Fl mice at 600 and 1200 ppm. In the mice of both sexes and female rats, no significant adverse effects were seen. However, in male rats a small increase in adenomas of the parathyroid gland was seen at 75 ppm. This increase in a rare benign tumor was not statistically significant but it was notable in being far above the incidence in historical controls and was judged to probably be related to the rotenone treatment. Finally, Greenman et al. (1993) attempted to repeat the experimental conditions used by Gosalves and colleagues exactly with intraperitoneal administration of rotenone for 42 days to female Wistar rats. No evidence of increased tumors, mammary or otherwise, was observed. In this paper he states that Gosalves had been unable to repeat previous mammary tumor results when the study was rerun with higher dietary levels of riboflavin. Since the difference in dietary riboflavin in Gosalves et al.'s original work and in the Greenman study is only about twofold, it seems unlikely to account for the radically different results on mammary tumorigenesis that they observed. The most reasonable conclusion to be reached from this suite of studies is that the weight of evidence suggests that rotenone is not an animal carcinogen but that the issue is still not fully resolved. The chronic feeding study with cube powder in rats of Hansen et al. (1965) gave an unexpected result. At the highest dietary doses (500 and 1000 ppm) a very clear decrease
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Pesticides Affecting Oxidative Phosphorylation
in the incidence of mammary tumor below that found in the control animals was seen. A decrease in pituitary adenomas was also observed in these high dose animals. A similar effect was observed in the carcinogenicity study in B6C3Fl mice conducted by Abdo et al. (1988) in which rotenone at 1200 ppm reduced the background incidence of hepatocellular carcinoma and neoplasms of the subcutaneous tissues well below that of the controls in males and increased their survival time. Cunningham et al. (1995) investigated rotenone's high dose anti tumorigenic effect in B6C3Fl male mice further and observed a dose-dependent decrease in hepatocellular proliferation. They attributed the antitumor action to rotenone's ability to block cell turnover which is in known to be an action that correlates with decreased tumor development. Rotenone has a well-established ability to decrease growth and division in a number of cell types in vitro (see below). In a further study to establish the biochemical mechanism for this antiproliferative effect, Wang et al. (1999) concluded that rotenone probably acted through two complementary mechanisms. The first is to cause an increase in apoptosis which removes DNA-damaged liver cells before they can progress to a carcinogenic condition, and the second is to decrease the energy status of the cells which favors a lower rate of basal hepatocyte proliferation. Both effects, in turn, result from the inihibition of electron transport. A previous study in B6C3Fl mice by Isenberg et al. (1997) using diethylnitrosamine to initiate DNA injury and a peroxisome proliferator to promote the development of these preneoplastic loci to focal lesions showed that rotenone antagonized the growth of the focal lesions. These authors arrived at the same conclusion as Wang et al. (1999) regarding a dual mechanism for this effect based on increased apoptosis and reduced proliferation. A number of other studies have also clearly demonstrated that rotenone antagonizes the proliferation of tumor cells. Figueras and Gosalvez (1973) found that rotenone injected intraperitoneally in mice at 3 mg/kg/day completely prevented the proliferation of a massive dose of Ehrlich ascites tumor cells. Mean time to death was increased by 110% compared to controls and, even so, death in the rotenone-treated animals was due to the toxic effects of the rotenone rather than tumor formation. Further confirming the anticancer activity of rotenoids, Gerhauser et al. (1995, 1997) discovered that four rotenoids from Mundulea sericae reduced the formation of chemically induced preneoplastic lesions in cultured mammary glands exposed to 7,12-dimethylbenz(a)anthracene and inhibited the development of papillomas in the two-stage mouse skin model. Subsequently, the effect of deguelin was studied in vivo in the two-stage (7, 12-dimethylbenz(a)anthracene/phorbol ester) mouse skin model and on rat mammary tumors induced by Nmethylnitrosourea. A strong anticancer effect was seen in each case (Udeani et aI., 1997). This general antiproliferative action was related to the ability of rotenoids to inhibit the expression of phorbol esterinduced ornithine decarboxylase (ODC) in mouse cells. ODC is involved in the regulation of cell proliferation and tumor promotion and is utilized as a biomarker for cancer chemopre-
ventive agents. In turn, the rotenoid effect on ODC induction was attributed to an observed rapid decrease in cellular ATP levels due to complex I inhibition. The alternative hypothesis that rotenoids could have their anticancer activity by inhibiting tubulin polymerization was rejected since deguelin is inactive in this regard, and activity correlated with complex I inhibitory activity rather than antimitotic capability among several other compounds. Rowlands and Casida (1998) confirmed this anticancer action using MCF-7 human breast cancer cells. Rotenoids such as rotenone and deguelin were also shown to decrease the mitogen-induced expression of ODC and to decrease the phorbol ester-induced production of reactive oxygen species in the cells. These authors concluded that the inhibition of complex I by rotenoids probably decreased the production of phorbol ester-induced reactive oxygen species which, in turn, regulate the expression of ODC activity. This concept was supported by the observation that two potent complex I inhibitors, the pesticides pyridaben and fenazaquin, caused an identical effect on ROS and ODC levels. This work was extended to include 29 rotenone analogs (Fang and Casida, 1998, 1999a) and 8 flavonoids and stilbenes (Fang and Casida, 1999b) isolated from cube insecticide. With these compounds, the ability to inhibit complex I from bovine heart was well correlated with their ability to decrease phorbol ester-induced ODC activity in the human cancer cells and their cytotoxicity to MCF-7 and Hepalclc7 cells, thus confirming an intimate connection between these different biological activities. MutagenicitylGenotoxicity A large number of studies on rotenone's capability to cause mutations or chromosomal effects have been conducted and are reviewed by Cal EPA (1997). Most of these results have been negative, but mutations were reported in an in vitro mouse lymphoma cell study, and increased sister chromatid exchange was observed in a study with Chinese hamster ovary cells. Recently, Guadano et al. (1998) examined the effect of rotenone on sister chromatid exchange, chromosomal aberrations, and micronuclei formation in cultured human lymphocytes. An increase in micronuclei was observed which was tentatively attributed to the well-known effects of rotenone on microtubule aggregation and spindle formation (Section 57.3.2.1). The weight of this evidence clearly suggests that rotenone is not genotoxic. Reproductive Toxicity In a two-generation dietary study in rats, the LOEL was found to be 37.5 ppm (1.88 mg/kg/day). Effects at this dose included reduced weight gain of both the pups and adults. Live litter sizes were reduced at 75 ppm. A reproduction study in hamster was conducted with 500 or 1000 ppm rotenone in the diet for 3 months. At 500 ppm, poor litter survival and maternal deaths occurred. At 1000 ppm, reduced testicular size and infertility were observed. Developmental ToxicitylTeratogenicity The status of rotenone as a teratogen is inconclusive. Several types of developmental effects have been observed in studies with rats and
57.3 Inhibitors of Complex I
mice, but generally at a dose that caused at least some minimal adverse effects in the mothers. In the rat, increased skeletal deformities including unossified sternebrae, renal pelvic cavitation, and distended ureters were seen in the offspring when 6 mg rotenone/kg/day was administered by gavage during pregnancy, but the same and lower doses also caused maternal effects such as lowered weight gain and several clinical signs (Cal EPA, 1997). Skeletal abnormalities in the form of supernumerary ribs were also seen in a teratology study in rats dosed orally with 5 mg rotenone/kg/day on days 6 to 15 of pregnancy, but again some maternal toxicity (reduced weight gain) was evident at this dose, and a dose of 10 mg/kg caused significant maternal lethality (Khera et aI., 1982). On the other hand, in an oral feeding study in rats during pregnancy, Spencer and Sing (1982) found an adverse effect on the fetuses at 100 ppm in the form of reduced fetal survival whereas the first maternal toxicity was observed only at higher doses of 400 ppm. The reason for the discrepancy between these studies is unclear. In the mouse, a LOEL dose of 24 mg/kg/day was fetotoxic and caused resorptions and decreased litter size. However, maternal toxicity, including mortality, was observed at the same dose. In an additional study in mice, no adverse developmental or toxic effects were seen at 15 mg/kg/day. Neurotoxicity Some of the earliest investigations of the toxic effects of rotenone demonstrated that it has the ability to block nerve transmission (Fukami, 1976), but there is no reason to suppose that this occurs other than by its effects as a respiratory inhibitor. As already discussed, because of its high respiratory activity, the nervous system in general is an important target tissue for compounds which affect oxphos. However, recent studies with rotenone and other complex I inhibitors have raised significant questions regarding their capability to cause specific neurodegenerative diseases, particularly parkinsonism (Betarbet et aI., 2000; Hollingworth and Ahammadsahib, 1995). A number of epidemiological studies have identified exposure to pesticides in general as a significant risk factor for the development of parkinsonism (e.g., see Gorell et aI., 1998; Menegon et aI., 1998), but specific causative agents have not been identified. At the same time, a relationship between the inhibition of complex I and the degeneration of dopaminergic tracts in the substantia nigra, which is a hallmark of parkinsonism, was brought to the fore by studies of the complex I inhibitor MPP+ in connection with the contamination of a street drug in San Francisco (reviewed by Tipton and Singer, 1993). MPP+ (which is not a pesticide) is specifically taken up and concentrated in dopaminergic neurons by the dopamine uptake system in sufficient quantities to inhibit complex I. It thus might be unusual or unique in its ability to cause destructive effects on these neurons in the substantia nigra since most complex I inhibitors are not likely to be concentrated in this way. This conclusion was supported by the work of Ferrante et al. (1997) who observed that when rotenone, a complex I inhibitor that does not interact with the dopamine transporter, was infused intravenously into rats for 7-9 days at 10-18 mg/kg/day, lesions in the striatum and globus pallida were observed but
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no specific neurodegeneration in the substantia nigra occurred. Systemic toxicity (cardiovascular effects) was also observed at these doses. However, other studies with rotenone in rats have shown that, under specific circumstances, it can cause loss of dopaminergic neurons in the substantia nigra and induce a condition that closely resembles parkinsonism [e.g., Heikkila et al. (1985) injected rotenone into the left median forebrain of rats and saw an extensive loss of dopaminergic functions in the neostriatum]. More specifically, recent work by Betarbet et al. (2000) has shown that the continuous intravenous infusion of rotenone in rats (2-3 mg/kg/day for 1-5 weeks) caused no systemic toxicity or generalized neuropathology but selectively destroyed dopaminergic tracts in the substantia nigra even though complex I was inhibited uniformally in the brain by about 75%, as judged by binding of tritiated dihydrorotenone. This treatment also induced the development of Lewy-like cellular inclusions and behavioral deficits which replicated those in parkinsonism. When administered at higher doses that cause systemic toxicity, rotenone caused more generalized neuronal lesions. The reason no comparable dopaminergic pathology was seen in the substantia nigra in the study of Ferrante et al. (1997) which was also conducted at higher rotenone doses causing systemic toxicity remains to be explained. Rotenone, as a highly lipophilic molecule, probably distributes readily throughout the nervous system and into cells. Neuronal cell death may be caused through the generation of reactive oxygen species that occurs on the partial inhibition of complex I. Free radical generation leading to apoptosis has been observed with low concentrations of rotenone in cultured neuronal dopaminergic cells (Betarbet et aI., 2000; Hartley et aI., 1994). It is significant in this regard to note that the degree of inhibition of complex I (75%) observed in their study does not inhibit overall electron transport in brain mitochondria because complex I is present in excess (Davey and Clark, 1996). Also, not all nerve cells were affected to the same degree (e.g., although dopaminergic neurons were destroyed in the striatum, GABAergic and cholinergic ones were not). The reason for the high sensitivity of the dopaminergic cells is unclear, but there is direct evidence that dopaminergic neurons have an unusually high sensitivity to disturbances of energy metabolism (MareySemper et al., 1993), and it has been suggested that this could be because they are already under oxidative stress because of their metabolism of dopamine and its precursors. These are catechols that form quinones which can take part in redox cycling reactions. It is important to keep in mind that in such studies, rotenone is being used as a convenient inhibitor of complex I to develop an animal model for parkinsonism. The results by no means implicate it as a likely cause of Parkinson's disease in human populations. No evidence for parkinsonism-like signs or neurodegenerative pathology has been reported in several subchronic and chronic dietary studies with rotenone in animals at a variety of dose levels (Cal EPA, 1997). The dietary or dermal routes of exposure obviously correspond much more closely to the typical human experience with rotenone than the intra-
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CHAPTER 57 Pesticides Affecting Oxidative Phosphorylation
venous infusion method used by Betarbet et al. (2000). This mode of administration bypasses several of the normal physical and metabolic barriers that protect the brain against poisoning. Also, many other compounds besides rotenone to which humans may be exposed, both pesticidal and naturally occurring, can inhibit complex I (Degli Esposti, 1998). They too would need to be considered if the theory that complex I inhibition underlies parkisonism is correct. Further, if the key event in the induction of Parkinson's disease is the generation of reactive oxygen species in dopaminergic neurons, many other types of compounds that generate free radicals such as inhibitors of complex III and redox cycling agents, or compounds that interfere with the normal protective radical scavenging mechanisms, might also cause this type of neurodegeneration. The available evidence and patterns of use do not suggest that rotenone is likely to be a significant primary or general cause of Parkinson's disease in human populations.
Hormonal Actions There is little indication from the available toxicology data that rotenone has endocrine disruptive effects. In a specific study to examine possible estrogenic actions in the positive rat mammary tumorigenesis studies of Gonsalves et aI., Olson and Sheehan (1979) found that rotenone did not act as an estrogen or estrogen antagonist either in vitro or in vivo. Human Toxicology Rotenone appears to be moderately toxic to humans. The oral lethal dose has been estimated to be 300500 mg/kg, but toxicity is likely to be strongly formulationdependent, and the presence of other bioactive components such as the synergist piperonyl butoxide in the commercial mixture might considerably influence its potency. In practice, human poisoning is extremely rare probably because of the low rotenone content of many commercial products, its frequently poor uptake from the gastrointrestinal tract, and its rapid emetic action (Gosselin et al., 1984; Ray, 1991). The signs, symptoms, and treatment of poisoning by rotenone are discussed by Ray (1991) and Reigart and Roberts (1999). These include numbness at sites of exposure, nausea, vomiting, and tremors (Haley, 1978). Neither fatalities nor systemic poisoning has been reported in the occupational use of rotenone. Numbness of the oral mucous membrane has been reported in workers or volunteers who were exposed to derris orally. Dermatitis and respiratory irritation, sometimes severe, have also been reported in those exposed occupationally under conditions of poor hygiene, but it is unclear whether these effects are due to rotenone or other constituents of the crude plant extracts. The one human death reported involved a child who consumed about 10 ml of an oily rotenone extract, giving an estimated exposure of 40 mg/kg. After vomiting, she gradually lost consciousness and died of respiratory arrest with postmortem evidence of anoxic damage to heart, lungs, and brain (DeWilde et aI., 1986). The ingestion of rotenone-containing roots was apparently a common means of committing suicide among inhabitants of New Ireland (Ray, 1991).
No antidotes for rotenone poisoning are known and decontamination followed by supportive therapy, if necessary, is indicated. Since menadione (vitamin K3) has been shown to bypass the rotenone inhibition site in complex I and to antagonize rotenone-induced respiratory depression and decreased blood pressure in rabbits (Santi and Toth, 1965), its use in human therapy has been proposed (Gosselin et aI., 1984). Biochemical Mechanism of Action The ability of rotenone to inhibit respiration and glutamate oxidation was first described by Fukami (1956, reviewed in Fukami 1976) using insect tissues. The extension of this work to show potent inhibition of rat liver mitochondrial respiration and the localization of rotenone's site of action in or near complex I was described in a series of papers by Lindahl and Oberg (1961). A large number of subsequent papers, reviewed by Haley (1978) and (Singer and Ramsay, 1992, 1994), have confirmed the view that the primary site of action for rotenone is the inhibition of respiration by interaction with complex I. Two binding sites have been identified for rotenone within complex I, a high affinity site and a lower affinity one, both of which must be occupied for full inhibition of electron flow through the complex (Singer and Ramsay, 1992, 1994). The recent studies with a photoaffinity label for the putative rotenone binding site already described also reveal a high affinity and a lower affinity binding site (Schuler et aI., 1999). These authors concluded that the 23 kDa PSST subunit of complex I plays a key role in high affinity binding. Rotenone inhibits cellular growth and division in many biological systems with inhibitory effects on cell division observed at 10-100 nM concentrations in a number of types of isolated cells [Ray, 1991; see Haley (1978) for earlier work]. At least two possible mechanisms for this antimitotic action have been described: a slowing of progression through the cell cycle due to rotenone's effects on respiration and ATP availability, and a direct effect on microtubule assembly and spindle formation which arrests cell division in mitosis and resembles the effects of colchicine in disrupting spindle formation (Brinkley et aI., 1974; Barham and Brinkley, 1976a, 1976b). Rotenone inhibited the rate of polymerization of rat brain tubulin in vitro with an ICso of 0.46 I-lM, a potency about fivefold higher than that of colchicine (Hoebeke and van Nijen, 1975), and, at a fairly high concentration (12 I-lM), it inhibited the polymerization of tubulin from bovine brain in vitro. Rotenone antagonized the binding of 3H-colchicine to bovine brain tubulin over the concentration range of 0.1 to 10 I-lM (Brinkley et aI., 1974), indicating an ability to bind directly to a site on the tubulin molecule. Although tubulin polymerization is an ATP-dependent process and in whole cells, failure to assemble tubulin could arise by an indirect action on ATP levels, Figueras and Gosalvez (1973) showed that the effect of rotenone on microtubule assembly does not arise through ATP deficiency but through direct binding to tubulin. Meisner and Sorfensen (1966) also observed that rotenone specifically arrests the cell cycle in metaphase by inhibition of spindle formation and concluded that this was probably due to binding to tubulin rather than
57.3 Inhibitors of Complex I
a direct effect on cell bioenergetics. Recently, Barrientos and Moraes (1999) found that the rotenone concentration needed to impact microtubule assembly in a human osteosarcoma-derived cell lines was fivefold higher than that which inhibited complex I by 100% (10- 7 M) and concluded that the two effects occur at nonoverlapping concentrations. The mechanism of cytotoxicity of rotenone has been studied in some detail. The depletion of ATP and inhibition other cellular functions driven by the mitochondrial membrane potential is potentially deleterious, although cells in culture treated with rotenone can frequently maintain adequate ATP for basic survival and cell division through glycolysis if glucose is provided. A second biochemical effect underlying rotenone's cytotoxicity may be an increase in superoxide and other ROS in the cell due to its inhibitory actions on complex I, as described in Section 57.1.7.1. This situation is quite complex since under some circumstances rotenone may also reduce the production of ROS (e.g., see the discussion of the role of rotenone as an anticancer agent and the induction of ODC above). This varied effect on ROS is discussed in detail by McLennan and Degli Esposti (2000). Absorption, Metabolism, and Elimination In mammals, rotenone is rapidly metabolized by oxidation, involving both ring hydroxylation and O-demethylation, and then eliminated (Ray, 1991). After an oral dose of 14C-rotenone in rats, most of the label was excreted in the feces and elimination was complete in 48-72 hr. No parent compound was excreted. Evidence was obtained for enterohepatic recycling of compounds excreted in the bile (Cal EPA, 1997). The metabolism of rotenone has been reviewed by Yamamoto (1969) and further studies were conducted by Yamamoto et al. (1971). In vitro, rotenone is subject to a variety of cytochrome P-450-catalyzed oxidations in vertebrate and insect tissues oxidation occurs (e.g., at both the methyl group and the double bond of the isopropenyl group, and at the carbon atom adjacent to the keto group between the Band C rings to yield a pair of enantiomers, rotenolones I and 11) (Fig. 57.8). Rotenoids are also subject to oxidative 0demethylation (Unai et al., 1973). In each case the metabolic product is less active as a complex I inhibitor. However, some oxidations do not eliminate toxicity. The 8' -hydroxy metabolite produced by oxidation of the methyl group of the isopropenyl side chain has an intraperitoneal toxicity to mice equal to that of rotenone, and rotenolone I also possesses considerable toxicity (Yamamoto, 1969). In fish, oxidative metabolism by liver microsomes also produces rotenolones and the products from the oxidation of the isopropenyl double bond (Erickson et aI., 1992). The comparative metabolism of rotenone in mammals, fish, and insects was investigated by Fukami et al. (1969) in relationship to its differential toxicity between these groups. Environmental Fate and Toxicity Rotenone is only slightly toxic to birds (Table 57.1). However it is highly toxic to many aquatic species, including fish (Table 57.1), probably because of the rapid uptake of this lipophilic compound and the presence of a sensitive and critical target site. Extensive additional
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data on the toxicity of rotenone to fish and other aquatic species are provided by Haley (1978). The generally high toxicity of rotenone to aquatic organisms includes many other species such as frogs and leeches (lethal effects at 100 ppb) and oysters (reduced shell growth at 10 ppb). Gingerich studied the uptake, metabolism, and elimination of rotenone in rainbow trout (Gingerich, 1986) and bluegill sunfish (Gingerich and Rach, 1985). In the bluegill, metabolism was quite rapid with only 20% of the radioactivity in the body being in the form of rotenone. The biological concentration factor (BCF) for total radioactivity in the whole body was 125, but because of the extensive metabolism, the BCF for rotenone was only 26. Six metabolites were found, with one being major. The half-life for the first and major phase of elimination was 25.8 hr and 80% of the radioactivity was lost in 3 days during the depuration period. Rainbow trout cleared rotenone more slowly with an elimination half-life of 68.5 hr and 85% of the radioactivity in the tissues was unchanged rotenone. An exception was the liver where rotenone represented less than 40% of the radioactivity present. The major route of excretion was through hepatobiliary excretion of polar metabolites. The high sensitivity of rotenone to photodecomposition (Cheng et aI., 1972) and oxidation limits its residual activity in the field and residues rapidly dissipate with a half life of 1-3 days under typical conditions. The speed of degradation depends on such predictable factors as the intensity of sunlight and the water temperature, e.g., the half life for rotenone in water varies from 10.3 days at 5°C to 13.9 hr at 25°C (Gilderhus et aI., 1986; Finlayson et aI., 2000, pp. 191-192). Rotenone binds strongly to soils with an estimated Koc of 10,000 mUg (Augustijn-Beckers et aI., 1994). Because of its rapid degradation after use in aquatic systems, the immediate impacts of rotenone's use in fish management programs on nontarget species are likely to be short-lived (Finlayson et aI., 2000), but even so, due to the initial impact, reductions in some benthic aquatic invertebrate populations may still be appreciable as long as five years after use (Mangum and Madrigal, 1999). It's use in a small pond study led to the immediate loss of the zooplankton population which did not return to normal until 8 months later (Beal and Anderson, 1993). 57.3.2.2 Lipophilic Nitrogen Heterocycles A group of novel acaricides, some also having insecticidal activity, that have their primary toxic action by inhibition of respiration in complex I was discovered in the 1980s and is now widely used. All are lipophilic compounds (log P values range from 4.6 to 6.4) with nitrogen-containing heterocyclic (specifically pyrazole, pyrimidine, or pyridazine) ring systems. Their properties and mechanism of action have been reviewed by Hollingworth and Ahammadsahib (1995), Degli Esposti (1998), and Liimmen (1998). Fenazaquin: General Properties and Uses 4-Tert-butylphenethylquinazolin-4-yl ether (Fig. 57.8), with CAS Reg. No. 120928-09-8, was discovered by DowElanco Co. and is described by Longhurst et al. (1992).
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It exists as colorless crystals, m.p. 78-80°C, v.p. 3.4 x 10-6 Pa (25°C), w.s. 0.22 ppm (25°C), log P5.5I. Fenazaquin (EL-436) is an acaricide on a wide range of crops under the trade names of Boramae, Demitan, Magister, Magus, Matador, Pride, and Totem.
Toxicology Profile (1993a).
The primary source of data is Anonymous
Acute Toxicity Fenazaquin has a relatively low acute oral toxicity to mice but it is much more toxic to rats (Table 57.1). The toxicity to guinea pigs lies between these two extremes with an LDso at 812 mg/kg. The dermal and inhalation toxicity are also moderate to low. IrritationlSensitization Fenazaquin causes slight eye irritation in rabbits but it is not a skin irritant or sensitizer. Subchronic Toxicity In 90-day dietary or gavage studies in rats, no notable toxicity was seen other than deceased food intake and weight gain at 30 mg/kg/day. The same was true in 90-day dietary studies at 50 mg/kg/day in hamsters and 15 mg/kg/day in dogs and in a one year dietary study in dogs at 12 mg/kg/day. A 21-day dermal study in rabbits was equally devoid of toxicological finding at 1000 mg/kg/day. Chronic Toxicity A two-year oral study in rats at doses up to 18 (M) or 25 (F) mg/kg/day revealed a moderate decrease in serum cholesterol and triglycerides. Treatment-related focal hepatocellular abnormalities were seen but no neoplasms were found. The no observed effect level (NOEL) was about 0.5 mg/kg/day. A second 18-month gavage study in Syrian golden hamsters revealed no specific toxicity other than a reduction in body weight at the highest dose, 35 mg/kg/day. A significant decrease in amyloidosis (a common disease in aging hamsters) was compound-related. The mechanism of this effect is unknown. The LOEL was 15 mg/kg/day. Carcinogenicity hamster studies.
Results were negative in the above rat and
MutagenicitylGenotoxicity No genotoxicity was observed in standard battery of seven assays with microbial and mammalian systems. Reproductive Toxicity No reproductive effects were observed in a multigenerational study in rats at doses as high as 25 mg/kg/day. Developmental ToxicitylTeratotogenicity No developmental toxicity or teratogenicity was observed in rats at the highest dose tested (40 mg/kg/day). In rabbits a marginal increase in early resorptions was found at the highest dose (60 mg/kg/day).
Biochemical Mechanism of Action Fenazaquin was shown to be a strong inhibitor of respiration in complex I of rat liver mitochondria and purified bovine heart complex I with ICso values in the range 20-50 nM by Hollingworth et al. (1994). Okun et al. (1999) confirmed this result using bovine submitochondrial particles. In a study using 3H-fenazaquin, Wood et al. (1996) showed high affinity binding to electron transport particles from bovine heart mitochondria. A variety of inhibitors of complex I competed with fenazaquin in this assay, including rotenoids and pyridaben, and their potency as inhibitors of complex I correlated well with their ability to displace fenazaquin from its binding site. However, subsequent studies with a photo affinity analog of fenazaquin labeled a protein (probably subunit d) in the stalk region of ATP synthase. Since fenazaquin does not inhibit the oligomycin-sensitive mitochondrial Mg2+ ATPase activity in mitochondria (Gadelhak and Hollingworth, unpublished), the relevance of this binding site for the toxic action of fenazaquin is dubious. Probably this is a low affinity site and the labeling arises from the high concentration of the photo affinity label used in the study. Fenazaquin is a peroxisome proliferator in mice and rats in vivo (Stott, 1996) but, as is typical of these compounds (Roberts, 1999), it is much less potent in hamsters. Studies with primary cultured hepatocytes from several species showed that fenazaquin caused a sevenfold increase in the activity of the peroxisome indicator enzyme, acyl coA oxidase, in mice, and a threefold increase in rats, but no change in hamster and human hepatocytes. In mouse hepatocytes an inhibition of gap junctional intercellular communication and significant oxidative stress were observed, but this was not seen with hepatocytes from the other species. As already noted, chronic feeding studies did not show an increased incidence of tumors (either of the liver or any other organ) in rats and hamsters. However, it is worth noting that the decrease in cholesterol and triglycerides and increase in hepatocellular focal abnormalities seen in the chronic study in rats are typical of the actions of a peroxisome proliferator (Roberts, 1999). Because of these species differences, Stott concluded that the hamster and rat, rather than the mouse, were the most appropriate models for human cancer risk assessment with fenazaquin. Fenazaquin inhibits complex I and blocks the induction of ornithine decarboxylase by several agents in human breast cancer cells in vitro. In this regard, fenazaquin's action parallels that of rotenone, and the block of ornithine decarboxylase correlates with anticancer activity. It is postulated that both the block of the decarboxylase and the anticancer action arise from a common cause (i.e., a reduction in the production of reactive oxygen species by the mitochondrial respiratory chain) (Rowlands and Casida, 1998). This hypothesis is discussed further in relationship to rotenone (Section 57.3.2.1). Absorption, Metabolism, and Elimination After an oral dose in rats, fenazaquin was rapidly absorbed with a peak plasma level appearing in 8 hr, almost quantitatively and rapidly metabolized, and eliminated primarily in the feces (80%) with significant amounts also in the urine (20%). Similar results were
57.3 Inhibitors of Complex I
obtained with goats. Initial metabolites in rats and goats were formed by three reactions; cleavage of the ether linkage, oxidation of the tert-butyl group to an alcohol and then a carboxylic acid, and oxidation of the quinazoline ring between the two nitrogen atoms.
Environmental Fate and Toxicity Fenazaquin has a low acute toxicity to birds (Table 57.1). The LCso value in 5day feeding studies in mallard ducks and bobwhite quail were >5000 ppm and the dietary NOEL in a one-generation reproduction study was 287 ppm. Fenazaquin does have a very high toxicity to fish and other aquatic species (Table 57.1). The no-effect concentration in a 63-day early life stage study in rainbow trout was extremely low, (0.96 ppb). Aquatic invertebrates are equally sensitive, with a no-effect concentration of 1.4 ppm for Daphnia growth and reproduction in a 3-week study. Shell growth is inhibited in the Eastern oyster at 5.4 ppb, and the acute LCso for brown shrimp is 15 ppb (Anonymous, 1993a). A mean BCF of 500-520 was found in trout but on depuration, 80% of the radioactivity was eliminated in 24 hr. Based on this result, Perkins et al. (1992) concluded that fenazaquin is unlikely to represent a bioaccumulation threat in fish. An aquatic microcosm study was used to simulate drift and runoff after application at 5 times the highest recommended use rate. No adverse effects were seen on aquatic life. Fenazaquin is rapidly and strongly adsorbed onto soil particles with Kd values ranging from 54 to 687 and Koc values from 18,700 to 41,200. The half-life for photolysis on the soil surface is 15 days under the summer sun at 40° latitude and 25°C, and the aerobic half-life in soils varies from 33 to 114 days with degradation mainly depending on microbial activity. It therefore has a low potential for leaching and accumulation in soils (Anonymous, 1993a). The half-life for photolysis in water is about 15 days under the same conditions as the soil photolysis. Fenazaquin is extremely stable to hydrolysis at neutral and alkaline pH (half-life is over a year) but it is hydrolyzed much more rapidly under acidic conditions. It is resistant to degradation by microbial action. In an aqueous microcosm study fenazaquin had a half-life in the water of 1 to 2 days as it partitioned onto solids where its half-life was about 5 months (Perkins et aI., 1992). Fenpyroximate: General Properties and Uses Tert-butyl (E)-a-(1,3-dimethyl-5-phenoxypyrazol-4-ylmethyleneaminooxy)-p-toluate (Fig. 57.8), CAS Reg. No. 134098-61-6, was discovered by Nihon Nohyaku Co and is described by Konno et al. (1990). The development of fenpyroximate has been reviewed by Hamaguchi et al. (1995). It exists as white crystals, m.p. 101-102°C, v.p. 7.4 x 10-6 Pa (25°C), W.s. 14.6 ppb (20°C), log P5.0I. Fenpyroximate can exist as E (trans) and Z (cis) geometrical isomers. The commercial insecticide is the E-form. In solution, fenpyroximate is readily degraded by photolysis to the Z-isomer with a half-life of 1.5 hr under conditions replicating those of sunlight. The Z-isomer then degrades with a half-life of 10.5 hr (Swanson et aI., 1995).
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Fenpyroximate (NNI-850, HOE 555-02A) is a widely used acaricide. Trade names include Acaben, Akari, Danitron, Dintron, Dynamite, Meteor, Naja, Ortus, Pamanrin, and Sequel.
Toxicology Profile The toxicological properties of fenpyroximate have been reviewed (FAOIWHO, 1996; U.S. EPA, 1999a) based almost entirely on unpublished studies submitted for registration. These are the source of much of the information below. An earlier summary of toxicity studies on fenpyroximate has also been published (Anonymous, 1992a). Acute Toxicity Fenpyroximate is moderately toxic after oral dosing in rodents but is somewhat more toxic by inhalation (Table 57.1). Signs of acute toxicity include hypoactivity and hypopnea. At necropsy, irritation of the gastrointestinal tract after oral dosing, and of the respiratory system after inhalation, was observed. IrritationlSensitization Fenpyroximate is not a skin irritant. It is a mild to moderate eye irritant and a moderate dermal sensitizer. Occular and dermal irritation have been noted among workers manufacturing fenpyroximate. Subchronic Toxicity In mice fed fenpyroximate at levels as high as 2000 ppm (175 mg/kg/day) for four weeks the only effects observed were several changes in hematological parameters and reduced food consumption and body weight gain. At 100 ppm (7.4 mg/kg/day) results in rats were similar with minimal liver hypertrophy and decreased white blood cell counts and plasma protein levels. At 500 ppm females had lowered acetyl- and butyrylcholinesterase levels in the blood. A subchronic oral feeding study with dogs caused some mortality at 50 mg/kg/day after appetite and body weight loss. Organ weight changes and signs of histopathological changes in the liver and kidney were also seen at this dose. Slight bradycardia and increased diarrhea were recorded at lower doses down to 10 mg/kg/day, the LOEL. Chronic Toxicity Non-neoplastic effects seen in chronic toxicity studies in rats at the highest dose tested (150 ppm; 6.9 mg/kg/day) included depressed growth, gastric ulceration, and pancreatic lobular degeneration in males and interstitial proliferation of the ovary and distention of the uterus in females. Pituitary neoplasia which compressed the brain was observed in males. There was a high incidence of pituitary adenoma in all treatment groups which nevertheless fell within the historical control range. In a lifetime feeding study in mice with doses as high as 72 mg/kg/day, no adverse effects were observed except reduced weight gain which was first noted at 10 mg/kg/day. In dogs, the results were quite similar to those in the subchronic study. Carcinogenicity Studies were negative in lifetime dietary studies in mice and rats and it was concluded that there was no evidence of compound-related carcinogenicity.
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MutagenicitylGenotoxicity Results were uniformly negative in a battery of in vitro and in vivo tests including point mutations in S. typhimurium and E. coli, mutation in Chinese hamster lung cells, chromosomal aberrations in human lymphocytes in vitro, micronucleus formation in vivo in CD-l mice, DNA repair in Bacillus subtilis, and unscheduled DNA synthesis in rat hepatocytes. Reproductive Toxicity No negative effects were observed in reproductive performance in rats in a two-generation study. Reduced weight gain in both the parents and offspring occurred at the highest dose tested (100 ppm; 8.6 mg/kg/day). Developmental ToxicitylTeratogenicity No evidence of teratogenocity or fetotoxicity was observed in rats and rabbits at doses that were not also maternally toxic. An increase in the number of thoracic ribs was observed at the highest dose, 25 mg/kg/day, in rats. An increase in incidence of retinal folds in rabbits at 5 mg/kg/day fell within the historical control range. Neurotoxicity No evidence of delayed neurotoxicity was observed in hens at two oral doses of 5000 mg/kg given 21 days apart. Biochemical Mechanism of Action A variety of studies show that fenpyroximate is a powerful inhibitor of mitochondrial respiration, acting specifically at complex I. This was first described by Motoba et al. (1992) who obtained an ICso for complex I from the mite (Tetranychus) in vitro of 80 nM compared to 400 nM for that from rat liver. Fenpyroximate in vivo depleted ATP in mites and caused malformation of the mitochondria, particularly in peripheral nerves. Friedrich et al. (1994) confirmed that fenpyroximate is a high potency inhibitor of vertebrate complex I and assigned it to the same inhibitor class as piericidin A. Okun et al. (1999), using a bovine mitochondrial preparation, characterized the relative potencies of several complex I-inhibiting acaricides as pyrimidifen > fenazaquin > fenpyroximate > rotenone. The Iso values in this study ranged from 2 to 20 nM. Degli Esposti (1998) obtained a similar result with an ICso of 4.6 nM for beef heart submitochondrial particles, making fenpyroximate slightly more active than rotenone. Jewess (1994) reported that fenpyroximate strongly inhibits NADH oxidation in blowfly flight muscle mitochondria and displaces 3H-dihydrorotenone from its specific binding site on complex I in blowfly muscle submitochondrial particles. The high specificity of fenpyroximate as a miticide is not based primarily on differences in target site sensitivity since it inhibits the mitochondrial complex I from rat liver and from spider mites with less than a 10-fold difference in potency (Motoba et aI., 1992). The main mechanism of selectivity has been shown to depend on differential rates of metabolic detoxification, particularly through removal of the t-Bu group yielding the free carboxylic acid analog. This metabolite is inactive as a complex I inhibitor. This apparent hydrolysis is largely catalyzed by cytochrome P-450 through hydroxylation of the t-Bu
group followed by intramolecular ester cleavage. Oxidative ester cleavage was rapid in the several mammals, fish, and insects tested but it did not occur in mites (Motoba et aI., 2000). Absorption, Metabolism, and Elimination Fenpyroximate is well absorbed, extensively metabolized, and rapidly excreted in rats after an oral dose, primarily (70-90%) in the feces (FAOIWHO, 1996; Nishizawa et aI., 1993). The elimination half-life was 6-9 hr at 2 mg/kg and 35-49 hr at 400 mg/kg. Considerable biliary excretion occurred (approximately 50% in 48 hr). Multiple sites of metabolism were identified including oxidation of the tert-butyl group and pyrazole methyl group, ester and oxime ether hydrolysis, aryl hydroxylation, N -demethylation, and EIZ isomerization. The uptake of fenpyroximate after dermal application in rats was very low. Environmental Fate and Toxicity Fenpyroximate has a very low toxicity to birds, and it is less toxic to fish and Daphnia than most of the other members of this group of pesticides, but it is still potent enough to be classified as highly toxic by the D.S. Environmental Protection Agency (Table 57.1). It is moderately persistent in soils with a half-life of 26-50 days where it is degraded primarily by microbial action (Izawa et aI., 1993). Pyridaben: General Properties and Uses 2- Tert-butyl-5-( 4tert-butylbenzylthio )-4-chloropyridazin-3(2H)-one (Fig. 57.8), CAS Reg. No. 96489-71-3, was discovered by Nissan Chemical Industries and is described by Hirata et al. (1988). The development of pyridaben is discussed by Hirata et al. (1995). It exists as white crystals, m.p. Ill-112°C, v.p. 2.5 x 10-4 Pa (20°C), W.s. 12 ppb (24°C), log P6.37. It is relatively stable to heat and hydrolysis but sensitive to photolytic decomposition with a half-life of about 30 min at pH 7. Pyridaben (NC-129, BAS 3001) is widely used as an acaricide with a long residual action and as an insecticide mainly against sucking insects. Trade names include Nexter, Oracle, Poseidon, Pyramite, Sanmite, and Starling. Toxicology Profile The general sources of data are D.S. EPA (1997a, 1998a, 2000a). Another useful source is the detailed summary of the regulatory toxicological studies of pyridaben provided by Igarashi and Sakamoto (1994).
Acute Toxicity Pyridaben shows moderate to low acute toxicity to mammals (Table 57.1). The intraperitoneal LDso was 68 mg/kg in male rats (Igarashi and Sakamoto, 1994). The dermal toxicity is low but toxicity by the inhalation route is quite high. With sublethal doses in mice and rats, clinical signs included decreased food consumption, diarrhea, hypothermia, bradycardia, bradypnea, decreased spontaneous motor activity, abnormal gait, prostration, eye closing, amd piloerection. At near lethal or lethal doses (300 mg/kg or more) depression of the central nervous and cardiovascular systems were stronger, but no change occurred in motor functions, including coordination, muscle strength, and neuromuscular transmission, or in
57.3 Inhibitors of Complex I sensory functions. Early gastric lavage was effective in presenting poisoning in rats and loperamide was efficacious in reducing the diarrhea that occurred at low doses. IrritationlSensitization Pyridaben caused slight and readily reversible eye irritation in rabbits, but it was not a skin irritant or sensitizer in guinea pigs in either the maximization test or the modified Buelher test. However, moderate to severe skin reactions were seen in a dermal exposure study when pyridaben was applied to pregnant Himalayan rabbits at 70 mg/kg/day over 2 weeks. SubchroniclChronic Toxicity In subchronic and chronic feeding studies with mice, rats, and dogs, toxicological observations were unremarkable with endpoints being decreased food intake and weight gain, and sporadic changes in clinical chemistry or organ weights. The most sensitive endpoint was in the dog with an LOEL in a one-year dietary study of 0.5 mg/kg/day based on increased clinical signs (emesis, salivation) and decreased body weight. Carcinogenicity Pyridaben was not oncogenic in typical lifetime feeding studies in the rat and mouse. It is classified by the U.S. Environmental Protection Agency as a Group E compound (no evidence for carcinogenicity to humans). MutagenicitylGenotoxicity Pyridaben was negative in a battery of microbial and mammalian tests (Ames Salmonella assay, Chinese hamster V79 cell point mutation assay, Chinese hamster lung cell cytogenetic damage in vitro, micronucleus assay in mice in vivo, the rec-assay in Bacillus subtilis, and DNA damage and repair test in E. coli). Reproductive Toxicity No adverse reproductive effects were observed in a multigenerational study in rats at dietary doses up to 80ppm. Developmental ToxicitylTeratogenicity Delayed ossification was seen in rats and rabbits after oral administration to pregnant animals during organogenesis, but this was believed to be a secondary result of maternal stress. Fetal and placental weights were decreased in rats at 30 mg/kg/day and abortions occurred in rabbits at the highest dose (15 mg/kg/day). In both species these events occurred only at doses that were clearly maternally toxic. No evidence for teratogenicity was seen. Neurotoxicity Pyridaben caused only a low degree of acute neurotoxicity in a standard battery of neurobehavioral tests when given at a single oral dose of 200 mg/kg in males. Effects included piloerection, hypoactiviy, tremors, and lowered body temperature, but these were sporadic and transient. In a longer term (90 day) study in rats, no neurotoxicity or neuropathology was seen at oral doses up to 27 mg/kg/day, but plasma cholinesterase activity was reduced in females.
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Biochemical Mechanism of Action Pyridaben inhibited respiration in insects in vivo and blocked respiration at complex I in several mitochondrial preparations in vitro (Hollingworth et aI., 1994). The inhibitory potency was extremely high with IC50 values of 0.8 nM for isolated beef heart complex I and 4.0 nM for rat liver mitochondria. The very high activity of pyridaben as an inhibitor of complex I has been confirmed by Degli Esposti (1998). As already described in the Introduction (Section 57.3.1), using photoaffinity label methodology, the binding site for pyridaben in bovine heart submitochondrial particles has been shown to be the PSST subunit of complex I (Schuler et aI., 1999). Like fenazaquin (Section 57.3.2.2) and rotenone (Section 57.2.2), pyridaben also blocks complex I and the induction of ornithine decarboxylase in human breast cancer cells in vitro, an action that correlates with antiproliferative and anticancer activity (Rowlands and Casida, 1998). Oxidation of the sulfur group in pyridaben yields the corresponding sulfoxide and sulfone. Compared to pyridaben, these compounds have a reduced ability to inhibit complex I and reduced toxicity to vertebrates. However, they show an increased toxicity to mammalian cells in vitro. This may indicate a second mechanism of toxicity for pyridaben in which sulfur oxidation activates the molecule to become reactive with nucleophiles such as tissue thiols (Schuler and Casida, 1998). The occurrence and possible toxicological significance of such an action in vivo are unknown. Absorption, Metabolism, and Elimination An oral dose of pyridaben in rats was absorbed fairly well (38-46% of the dose), rapidly and completely metabolized, and eliminated mainly in the feces (80-97%) within 96 hr. Nearly 20% of the excreted residue in the feces was the parent compound. A large number of metabolites (20-30) were detected in the urine and feces with none predominant. They arose primarily from oxidation of the two tert-Bu groups and glutathione conjugation with the pyridazinone ring, probably following oxidation of the sulfur atom (Hirata et aI., 1995). Considerable biliary excretion occurred (22-30% in 24 hr) and there was evidence of enterohepatic circulation of these metabolites. Environmental Fate and Toxicity Pyridaben has a low acute toxicity to birds, but it is extremely toxic to aquatic species (Fig. 57.8). Its persistence in soil is relatively brief due to rapid microbial degradation (e.g., the half-life under aerobic conditions is reported to be less than 3 weeks). In natural water in the dark, the half-life is about 10 days, due mainly to microbial action since pyridaben is stable to hydrolysis over the pH range 5-9. The half-life including aqueous photolysis is about 30 min at pH 7 (Tomlin, 2000). Pyrimidifen: General Properties and Uses 5-Chloro- N{2-[4-(2-ethoxyethy1)-2,3-dimethylphenoxy]ethyl }-6-ethylpyrimidin-4-amine (Fig. 57.8), CAS Reg. No. 105779-78-0, was discovered by Ube Industries with joint development by the Sankyo Company.
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It exists as a white crystal, m.p. 69-71 QC, v.p. 1.6 x 10-7 Pa (25°C), w.s. 2.17 ppm (25°C), log P4.59.1t is stable to hydrol-
ysis over a broad pH range. Pyrimidifen (E-787, SU-8801, SU-9118) is used as an acaricide and insecticide under the trade name Miteclean. Toxicology Profile (1999b).
The general data source is Anonymous
Acute Toxicity Pyrimidifen has a higher acute oral toxicity to mammals and birds than most other members of this group (Table 57.1). Signs of poisoning in rats include apathy, decreased respiration, abnormal gait, and decreased urinary output. IrritationlSensitization Pyrimidifen causes slight eye irritation but no dermal irritation or sensitization. Subchronic Toxicity In a 90-day dietary study in rats, toxicological effects were rather nonspecific including decreased weight gain, some changes in blood chemistry, and changes in organ weights. An increase in liver weight was seen in females at 0.69 mg/kg/day, the lowest dose tested. Similar results were seen in mice and the LOEL was established at 17.7 mg/kg/day. In dogs the LOEL was 0.5 mg/kg/day based on observations of diarrhea. Increased salivation and reduced body weight gain were seen at the highest dose, 4.5 mg/kg/day. Chronic Toxicity In a 2-year study in rats, dietary levels at 100 ppm (3.9 mg/kg/day) increased the incidence of benign adrenal pheochromocytoma. Increased weight and discoloration of the kidney and lipofuscin deposition in the tubules were also seen. The signs of toxicity in parallel long-term feeding studies in dogs and mice were unremarkable. Carcinogenicity Long-term feeding studies in rats, mice, and dogs proved negative for carcinogenicity. MutagenicitylGenotoxicity Pyrimidifen gave negative results in the Ames assay and a bacterial DNA repair test (Rec-assay). It did not cause chromosome aberrations in the Chinese hamster lung fibroblast assay. Reproductive Toxicity None was seen in a two-generation study in rats at doses of 7.6-9.5 mg/kg/day. Developmental ToxicitylTeratogenicity Some minor skeletal abnormalities were found when pyrimidifen was fed to pregnant rats and rabbits at 20-25 mg/kg/day, but these doses also caused maternal toxicity and it was concluded that pyrimidifen is not teratogenic.
Biochemical Mechanism of Action Pyrimidifen is an extremely potent inhibitor of complex I in bovine submitochondrial particles (150 of 2 nM) and competes with high affinity for a binding domain that is in common with fenazaquin, fenpyroximate, and rotenone (Okun et aI., 1999).
Tebufenpyrad: General Properties and Uses N-(4-tertbutylbenzyl )-4-chloro-3-ethyl-1-methylpyrazole-5-carboxamide (Fig. 57.8), CAS Reg. No. 119168-77-3, was discovered by Mitsubishi Kasei Corporation and was developed in partnership with the American Cyanamid Co. It is described by Kyomura et al. (1990) and Inoue and Fukuchi (1994). The structural conformation of tebufenpyrad has been determined by x-ray crystallography (Osano et aI., 1991). The synthesis and structure-acaricidal activity relations of tebufenpyrad and related compounds are described by Okada et al. (1991). A closely related compound, tolfenpyrad (OMI-88; CAS Reg. No. 129558-76-5; Fig. 57.8), with stronger insecticidal activity than tebufenpyrad and reasonable mammalian safety is currently under development in Japan (Okada et aI., 1999). General Properties Tekufenpyrael exists as white crystals, m.p. 61-62°C, v.p. 1 x 10-5 Pa (25°C), w.s. 2.8 ppm (25°C), log P4.61-5.04. It is stable to aqueous hydrolysis with a halflife over 28 days at pH 5-9. Uses Tebufenpyrad(MK-239, SAN-831, AC 801,757) is used primarily as an acaricide but it also has activity against sucking insects. Trade names include Comanch6, Masa'i, Oscar, and Pyranica. Toxicology Profile The general sources of the data are Anonymous (l993b) and Mitsubishi Chemical Industries (1995). Acute Toxicity Tebufenpyrad shows moderate acute toxicity to mammals (Table 57.1). The toxicity to rabbits is higher with an acute oral LDso between 40 and 100 mg/kg. By the dermal and inhalation routes its toxicity is low. Clinical signs of poisoning include slowed respiration, decreased locomotor activity, and prostration. Increased salivation is also seen after respiratory exposure. In an attempt to test therapeutic strategies, rabbits dosed orally at 100 mg/kg (LDlOO) were given dimorpholamine intravenously at 3 or 6 mg/kg intravenously when respiration had decreased to 50% of normal. This treatment improved respiratory function significantly but it is not stated whether mortality was decreased. IrritationlSensitization Tebufenpyrad caused slight, reversible eye irritation in rabbits, but it was not a skin irritant or sensitizer. Subchronic Toxicity A uniform picture emerges for 90-day dietary or gavage studies in rats, mice, and dogs. The only effects consistently observed were decreased weight gain and increased liver weight at the highest doses [about 28 mg/kg/day in the rat, 160 (M) to 220 (F) mg/kg/day in mice, and 10 mg/kg/day in dogs]. Some liver hypertrophy was seen in rats, and vomiting and diarrhea occurred in dogs. Chronic Toxicity Longer term exposures in these animals caused essentially the same responses as in the 90-day studies, but at somewhat lower doses. The NOELs in both rats
57.4 Inhibitors of Complex II (24-month dietary exposure) and dogs (12-month oral dosing) were found at 1 mg/kg/day. Chronic gastritis was observed in the dogs at the highest dose (20 mg/kg/day). The NOEL in an 18-month dietary study in mice was 4 mg/kg/day. Carcinogenicity No evidence of carcinogenicity was seen in the lifetime dietary studies in mice and rats. MutagenicitylGenotoxicity Tebufenpyrad is not mutagenic or clastogenic in a typical battery of bacterial and mammalian tests including the Ames Salmonella, Chinese hamster ovary cell, chromosomal aberrations in lymphocytes, mouse micronucleus, and unscheduled DNA synthesis assays. Reproductive Toxicity No serious adverse reproductive effects were observed in a two-generation dietary study in rats. A decrease in pup weight gain was recorded at the highest dose, 16 mg/kg/day, which was the LOEL for the study. Developmental ToxicitylTeratogenicity No developmental or teratogenic effects were observed when tebufenpyrad was given orally during organogenesis to rats at 150 mg/kg/day or to rabbits at 40 mg/kg/day, the highest doses tested. The NOEL for maternal effects was about 15 mg/kg/day in each study.
Biochemical Mechanism of Action Like the other compounds in this group, tebufenpyrad is a powerful and specific inhibitor of complex I with an Iso value of 6 nM for bovine heart submitochondrial particles (Degli Esposti, 1998) and 2 nM for mitochondria from housefly flight muscle (Liimmen, 1998). Absorption, Metabolism, and Elimination Tebufenpyrad is rapidly metabolized and cleared after an oral dose in rats. The metabolism of tebufenpyrad in rats in vitro and in vivo was investigated by Ogawa and Ihashi (1993). Hydroxylations of the ethyl and t-Bu groups were the predominant reactions both in vitro, using an S-9liver fraction, and in vivo. Subsequent oxidation of these initial alcohols to carboxylic acids and conjugation of the alcohols with sulfate occurred in vivo. Little cleavage of the amide bond was observed. Environmental Fate and Toxicity Tebufenpyrad is relatively safe to birds (Table 57.1). In 8-day feeding studies in mallard ducks and quail, the LCso was > 5000 ppm in both species. Like the other members of this group it is highly toxic to fish and to Daphnia (Table 57.1). Other aquatic invertebrates are also highly sensitive to tebufenpyrad (e.g., the LCso for mysid shrimps is 22 ppb and the ECso for the inhibition of shell growth in the Eastern oyster is 62 ppb). The 22-day no-effect concentration for reproduction in Daphnia is very low at 2.4 ppb. The uptake, metabolism, and excretion of tebufenpyrad by carp have been studied by Saito et al. (1994). The BCF at steady state, which was reached within about 4 days of exposure, was 864. However, less than 4% of the radioactivity in the body was unchanged tebufenpyrad, so the BCF for the parent is only 29. The major metabolites were formed by sequential oxidation of the
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tert-butyl group to the alcohol and then to the carboxylic acid and their subsequent conjugation with sulfate and glucuronic acid residues. The half-life for elimination during the depuration phase was about 12 hrwith 98% of the radioactivity cleared within 7 days. Thus, although tebufenpyrad is lipophilic, is readily taken up from water, and might be expected to show a high level of bioaccumulation, it is also rapidly metabolized and cleared which greatly decreases the degree of accumulation. The acylation of the amide nitrogen between the two rings in analogs of tebufenpyrad produces compounds that are improved as acaricides and also show much lower toxicity to fish, perhaps due to differential rates of metabolic deacylation to release the active compound in these organisms (Obata et aI., 1999). Tebufenpyrad binds firmly to soil organic matter with Koc values from 1380 to 4930. Together with its low water solubility, this indicates a very low potential for leaching. It has a moderate persistence in aerobic soils with a half-life of about 1 to 2 months.
57.4 INHIBITORS OF COMPLEX 11 57.4.1 INTRODUCTION Relatively few pesticides are thought to have their primary toxic action through effects on complex 11. Important pesticides that do act in this way are fungicides in the carboxamide (carboxanilide) group. These have been reviewed by Kulka and von Schmeling (1995). In some cases the information regarding these compounds that is available in the open literature or toxicological databases is quite minimal. Based on this rather limited information, they seem to have virtually no notable adverse effects and they have attracted minimal toxicological interest beyond the studies required by regulatory authorities to obtain approval for use. 57.4.2 PROPERTIES OF SPECIFIC COMPOUNDS 57.4.2.1 Carboxamides The forerunner of the group is the fungicide carboxin which was discovered in the mid-1960s (Kulka and von Schmeling, 1995). A relatively large number of carboxamides have subsequently been used as agricultural fungicides, but several of these are not now produced commercially [e.g., benodanil, mebenil, methfuroxam, metsulfovax, and pyracarbolid; Tomlin (2000)]. On the other hand, new members are still being added to the group (e.g., thifluzamide and furametpyr were both first registered for use as pesticides in the 1990s). The carboxamides are systemic fungicides with both protectant and curative actions which are used in seed treatments and in foliar and soil treatments, primarily to control Basidiomycetes. They are often used in mixtures with other fungicides or insecticides. They generally have low or very low toxicities to
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terrestrial vertebrates and aquatic species, earthworms, and insects. Their chronic toxicity is equally unremarkable. The lack of recorded incidents of human poisoning appears to substantiate their safety under practical conditions of use. However, the toxicological information available in the open literature is quite limited for most of these compounds. They are rapidly metabolized and eliminated by mammals. It is interesting to note that fungicides within the carboxin group have proved effective as potential therapeutic agents for the human immunodeficiency virus (HIV-l) by inhibiting the viral reverse transcriptase (Bader et aI., 1991). Considerable work has followed from this discovery but this lies outside the scope of the chapter. Succinate-ubiquinone oxidoreductase was shown to be the target for carboxin and oxycarboxin in fungi 30 years ago (Mathre, 1971; White, 1971). The extensive work on this topic is reviewed in detail by White and Georgopoulos (1992) and Schewe and Lyr (1995). Despite the statement by Schewe and Lyr (1995) that mammalian mitochondria "show a sensitivity to carboxins comparable to those of sensitive fungi," a view that is also supported by the work of Mowery et al. (1976, 1977), other results indicate that complex 11 from vertebrate mitochondria is often significantly less sensitive to inhibition by carboxamides than that of target fungi (Mathre, 1971; Motoba et aI., 1988; Shimizu et aI., 1992). It appears that results vary with the type of preparation, the electron acceptor used in the assay, and the structure of the carboxamide tested. The complex 11 from nontarget fungal and plant mitochondria is generally found to be markedly less sensitive than that of target fungi. This allows for the use of these compounds in plant protection without phytotoxicity (Schewe and Lyr, 1995; Shimizu et aI., 1992). The precise binding site for the carboxamides is not known. Solubilized succinic dehydrogenase from bovine heart or fungal mitochondria, consisting of the FP and IP subunits (Fig. 57.3), is not inhibited by carboxin and other carboxanilides (Mowery et aI., 1976; Schewe and Lyr, 1995; Shimizu et aI., 1992), so the binding site probably lies within the membrane-associated anchor portion of the complex. Inhibition arises from a disruption of the transfer of electrons from the iron-sulfur center to ubiquinone, a situation analogous to the acaricide-insecticides that inhibit complex I (Section 57.3.1). This is in accord with photoaffinity labeling studies that indicate that carboxamides bind to a site associated with the ubiquinone-binding proteins of complex 11 (CII-3 and CII-4, Fig. 57.3) and not to the flavoprotein-containing or iron-sulfur proteins (White and Georgopoulos, 1992). A recent study by Matsson et al. (1998) using carboxamide-resistant mutants of Paracoccus denitrificans concluded that the key mutation is in one of these two membrane-located anchor polypeptides at a location adjacent to the Fe-S cluster. It was suggested that this may form part of the carboxamide binding site. On other hand, in the fungus Ustilago maydis, a study of resistance to carboxin showed that it is associated with a point mutation in the gene encoding the third iron-sulfur cluster in the IP subunit (Keon et al., 1994). These authors concluded that carboxamides are probably interposed between this high poten-
fY
S
0;3
~O
C~N--o o H
H
Carboxin
Oxycarboxin
Fenfurarn
Flutolanil
~?
H3C~Cl H;~ 3
H3
0
0;3
Mepron i I
Furametpyr CF 3
~oBr
H,~ :~F" Br
Thifluzarnide Figure 57.9
Pesticides that act as inhibitors of complex IL
tial iron-sulfur center and the ubiquinone binding proteins and thereby hinder electron transfer between them. This resistance mutation may then cause a conformational change in the IP subunit that allows it to continue to feed electrons to the ubiquinone site even in the presence of the carboxin. An identical result was found by Skinner et al. (1998) in a different organism. They concluded that this mutation affecting carboxin sensitivity may not be occurring within the carboxin binding domain but at an allosteric site that controls the access of carboxin to its binding site. Inhibitor access is restricted in the resistant form leading to a decrease in target site sensitivity. Clearly, multiple mutations may lead to carboxin resistance (Georgopoulos, 1995) and these appear to operate by different mechanisms. The exact nature of the inhibition of complex 11 by carboxamides therefore still awaits clarification and depends on the development of a more detailed knowledge of the structure and mechanism of the complex.
Carboxin: General Properties and Uses 5,6-Dihydro-2methyl-l,4-oxathiin-3-carboxanilide (Fig. 57.9), CAS Reg. No. 5234-68-4, was discovered by Uniroyal Chemical Co. and is described by von Schmeling and Kulka (1966). It exists as white crystals (dimorphic), m.p. 92-93°C or 98100°C (depending on form), v.p. 2.5 x 10-5 Pa (25°C), w.S. 199 ppm (25°C), log P2.2, HLC 3 x 10- 5 Pam3 mol-I. It is stable to hydrolysis from pH 5 to 9. Carboxin is a systemic fungicide used as a seed treatment and to prevent seedling diseases of cereals and many other crops. It is often formulated in combination with other fungi-
57.4 Inhibitors of Complex 11 cides or insecticides. Common trade names include Cerevax, Enhance, Hiltavax, Kemikar, Kisvax, Oxatin, and Vitavax. Toxicology Profile The primary sources of data are V.S. EPA (1989a) and Cal EPA (1994), which provide an extensive overview of the toxicology and fate of carboxin based on unpublished studies submitted for pesticide registration, and the HSDB (2000). Acute Toxicity Carboxin has a very low acute toxicity to terrestrial vertebrates by all routes of exposure (Table 57.2). IrritationlSensitization Carboxin is not a skin irritant but it can cause serious eye irritation. Subchronic Toxicity A 90-day feeding study in rats at doses up to 1000 mg/kg body weight/day (20,000 ppm in the diet) revealed no gross pathological changes. Increased blood urea nitrogen and decreased hemoglobin were seen, but only in females at the highest dose. The LOEL was 30 mg/kg/day and involved microscopic inflammatory degenerative renal changes including cortical tubular degeneration. Fibrosis of the kidney medulla was observed at 100 mg/kg/day. In a parallel study with mice, female mortality was seen at the highest dose (912 mg/kg/day). No gross pathology was seen, but microscopic liver centrilobular hypertrophy was observed at higher doses (about 400 mg/kg/day and higher). Chronic Toxicity In a 2-year dietary studies in rats, a result similar to the 90-day study was obtained with chronic nephritis and degeneration of tubular cells observed at 200 ppm in males and 300 ppm in females (Cal EPA, 1994). No serious adverse effects were detected in a I-year feeding study in dogs at the highest dose tested, 7500 ppm. Carcinogenicity In mice, no carcinogenic responses were recorded in dietary studies with doses as high as 751 (M) and 912 (F) mg/kg/day although decreased survival and liver hypertrophy were seen at the highest doses. An increase in pulmonary and alveolar-bronchiolar adenomas in males at the highest dose was judged probably not to be compound-related (V.S. EPA, 1989a). Results were also negative in a 2-year studies in rats with dietary levels up to 600 ppm. Because of the equivocal positive effect in male mice, carboxin was placed by the V.S. Environmental Protection Agency in carcinogen group D (not classifiable as to human carcinogenicity). MutagenicitylGenotoxicity Carboxin is nonmutagenic or very weakly mutagenic in several bacterial and fungal cell systems (Adhikari and Grover, 1989; de Bertoldi et aI., 1980; Grover et aI., 1990; Moriya et aI., 1983; U.S. EPA, 1989a) and in the SOS microplate assay using E. coli (Venkat et aI., 1995). It was reported to cause dose-dependent clastogenic activity with a variety of chromosomal aberrations in rat bone marrow cells after in vivo administration at doses of 191 and 382 mg/kg, and in studies with plants (Adhikari and Grover, 1988, 1989).
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However, in other similar tests with rats, the clastogenic response was negative at doses as high as 2000 mg/kg (Cal EPA, 1994). In studies with Chinese hamster ovary cells in vitro an increase in chromosomal aberrations was seen with metabolic activation, although false positives in this test are common (Cal EPA, 1994). Carboxin also caused unscheduled DNA repair in primary rat hepatocytes over the concentration range of 5.1 to 103 f.!g/ml (V.S. EPA, 1989a). The evidence regarding the genotoxicity of carboxin is therefore mixed, but a significant possibility does exist that carboxin has genotoxic effects. Reproductive Toxicity No treatment-related effects on reproductive performance were observed at 5-30 mg/kg/day in a three-generation study in rats, but moderate growth suppression was seen in nursing pups at 30 mg/kg/day (U.S. EPA, 1989a). A subsequent two-generation study (Cal EPA, 1994) confirmed this result but at 200 ppm (M) and 300 ppm (F) it also revealed the nephrotoxicity typically seen in rats with this compound. Developmental ToxicitylTeratogenicity In a developmental study in pregnant rats at 175 (but not 90) mg/kg/day, carboxin caused decreased fetal weight but this dose also caused decreased weight gain in the dams. No specific teratogenicity was observed. Rabbits dosed at 375 and 750 mg/kg/day during gestation had increased abortions but maternal toxicity was also observed. No developmental effects or fetal malformations were detected. Carboxin therefore does not appear to be a developmental toxic ant or teratogen. Human Toxicology The only published example of human poisoning by carboxin involves a 7-year-old boy who ate several handfuls of carboxin-treated wheat seed. Symptoms included vomiting and headache which developed within 1 hr but were rapidly resolved after administration of an emetic. The amount of carboxin consumed is unclear (U.S. EPA, 1989a). Biochemical Mechanism of Action Carboxin is a strong inhibitor of succinic deydrogenase from a target fungus, Ustilago maydis (Mathre, 1971). With Rhizoctonia solani mitochondrial succinate dehydrogenase, carboxin has an 150 of 0.32 J.lM. White and Thorn (1975) obtained an 150 value of 0.5 J.lM in the same system. Reports of its activity against vertebrate mitochondrial complex 11 are varied. Mathre (1971) and Shimizu et al. (1992) report that carboxin is at least 10- to 20-fold more active against the sensitive fungal enzyme than complex 11 in mitochondria from rodent liver for which IC50 values were > 30 J.lM. On the other hand Mowery et al. (1976, 1977) found that using various complex 11 preparation from bovine heart, carboxin was a potent inhibitor with 150 values below 1 J.lM. It is possible that differences both between the species and the types of complex 11 preparations studied can explain these discrepancies. Carboxin is also reported to inhibit oxidative metabolism and, specifically, succinate dehydrogenase in mitochondria from rat liver and bone but no estimate of its potency is given (Gosselin et aI., 1984).
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CHAPTER 57
Pesticides Affecting Oxidative Phosphorylation
Table 57.2 Acute Toxicity of Pesticides Acting as Inhibitors of Mitochondrial Complex II to Selected Nontarget Species LDso (mg/kg)
LCso (mg/l)
Acute oral Compound
Rat (M; F)
Carboxin Fenfuram
3820 12,900
Mouse (M) 41S0 24S0c
Quail 24,000b
Flutolanil Furametpyr
>10,000 640;S90
>10,000 660
>2000
Mepronil
> 10,000; > 10,000 S816; 1632 >SOOO
>10,000
>8000b
Oxycarboxin Thifluzamide
Duck
>2000
(24--96 hr)
Inhalation
Rat or rabbit
Rat
Fisha
>20 >10.3 >S.98
1200-2300 1l,000 2300-S400
>4000 4S00 >SOOO >2000
12S0
LCSO (ppb)
Acute dermal
>10,000 >SOOO >SOOO
>1.32 >2.0
16S0 8600-10,000 19,900-28,100 1200-2900
Daphnia
84,400 SO,OOOd
>10,000 69,100 1600
aRange of values from several species, most commonly including the rainbow trout and carp. bData for hen. CData for cat. d Six hr exposure.
Absorption, Metabolism, and Elimination When carboxin was fed to rats and rabbits by gavage at 235 mg/kg, approximately 40% of dose was excreted in the feces by rats mostly as unchanged carboxin, but only 10% was excreted through the fecal route in rabbits (Waring, 1973). In the urine, excretion was almost complete in 24 hr. Some parent compound was excreted in the urine unchanged, particularly by rats. Major metabolites were the products of para and ortho hydroxylation of the phenyl moiety followed by glucuronidation (Waring, 1973). Excretion of carboxin sulfoxide was not seen although some other sulfoxidized metabolites were identified. In dogs, oxidation to the sulfoxide occurred and this was excreted in the urine along with the parent compound (Gosselin et aI., 1984). Environmental Fate and Toxicity Carboxin's tOXICIty to birds is very low (Table 57.2). In 8-day feeding studies with birds, the LCso was greater than >4640 ppm in the feed with mallard ducks and> 10,000 ppm with bobwhite quail. Carboxin has a low acute toxicity to aquatic species including fish and invertebrates such as Daphnia (Table 57.2) and juvenile crayfish (LCso of 217 ppm). A bioconcentration factor of 34 has been calculated for carboxin which suggests little risk of accumulation by aquatic species. Carboxin is converted to its sulfoxide in vivo and in water and soils, but further oxidation to the sulfone (oxycarboxin, see below) is extremely slow. Carboxin's half-life in soils is 1-3 days (Balasubramanya and Patil, 1980; Chin et aI., 1970, Wauchope et aI., 1992). The range of Koc values (80-259) indicates a medium to high mobility in soil and significant leaching is observed in laboratory tests (U.S. EPA, 1989a). Fenfuram: General Properties and Uses 2-Methylfuran3-carboxanilide (Fig. 57.9), CAS Reg. No. 24691-80-3, was discovered by Shell Research Ltd. and developed by Aventis CropScience. It exists as colorless to light brown crystals, m.p. 109110°C, v.p. approximately 2 x lO- s Pa (20°C), W.s. 100 ppm
(20°C). It is very stable to heat and to hydrolysis except under strongly acid or alkaline conditions. Fenfuram is a systemic fungicide used in seed treatments for cereals under the trade name Pano-ram. Toxicology Profile Published data regarding the toxicology of fenfuram are very limited. The primary source of data is Tomlin (2000). Acute Toxicity Fenfuram has a low acute tOXICIty to vertebrates (Table 57.2). The intraperitoneal LDso in rats is 1490 m/kg. SubchroniclChronic Toxicity In a 90-day dietary study in dogs a NOEL of 300 mg/kg/day was observed. In a 2-year feeding study in rats, the NOEL was 10 mg/kg/day. Mutagenicity
Fenfuram was negative in the Ames assay.
Absorption, Metabolism, and Elimination After an oral dose, rats eliminated 83% in 16 hr, primarily in the urine. Biochemical Mechanism of Action Fenfuram inhibits succinate oxidation in mitochondria from the target species, Ustilago maydis, with an Iso value of 4.2 J..lM (White and Thorn, 1975). Environmental Fate and Toxicity Fenfuram's toxicity to fish appears to be very low (Table 57.2). The half-life for fenfuram in soil is about 42 days and the Koc value is about 300, indicating reasonably firm binding to soil organic matter (Augustijn-Beckers et aI., 1994). Flutolanil: General Properties and Uses a, a, a-Trifiuoro3' -isopropoxy-o-toluanilide (Fig. 57.9), CAS Reg. No. 6633296-5, was discovered by Nihon Nohyaku Co. Ltd. and is described by Araki and Yabutani (1981) and Araki (1985). Its development is reviewed by Kurono (1985) and Araki and Yabutani (1993).
57.4 Inhibitors of Complex II It is exists as white crystals, m.p. 104-105°C, v.p. 6.5 x 10-6 Pa (25°C) (Tomlin, 2000), also widely cited as 1.8 x 10- 3 Pa (20°C) (e.g., see Araki, 1985), W.s. 9.6 ppm (20°C), log P3.7. It is stable to heat (5 hr at 100°C) and in solution at pH 3-9. Flutolanil is relatively resistant to photodegradation (Tsao and Eto, 1991). Flutolanil (NNF-136) is a systemic fungicide with both protective and curative actions on a wide range crops turf and ornamentals through foliar, seed, or soil application. Trade names include Folistar, Iota, Monarch, Moncut, Prostar, Protar, and Symphonie.
Toxicology Profile The primary sources of data are Araki (1985) and the U.S. Environmental Protection Agency's IRIS (Integrated Risk Information System) data base (http://www. epa.gov/ngispgm3/iris). The toxicology of flutolanil has been reviewed in Japanese (Anonymous, 1988). Acute Toxicity Flutolanil has a very low acute toxicity to terrestrial vertebrates (Table 57.2). IrritationlSensitization and skin.
Flutolanil is nonirritating to the eyes
Subchronic Toxicity Ninety-day feeding studies in rats and dogs gave LOELs of 200 and 400 mg/kg/day, respectively. The major finding was liver enlargement. Chronic Toxicity A 2-year feeding study in dogs had an LOEL of 250 mg/kg/day with emesis, salivation, soft stools, and lowered body weight gain as the major observations. Carcinogenicity Flutolanil was not oncogenic in a 2-year feeding study. The U.S. Environmental Protection Agency's category is Class E (evidence of noncarcinogenicity for humans). MutagencitylGenotoxicity Flutolanil is not mutagenic in several standard tests (Ames assay, Rec DNA repair assay). Reproductive Toxicity In a three-generation reproduction and teratology study in rats, even at 10,000 ppm (662-1002 mg/kg/ day) in the diet, no compound-related clinical signs of toxicity, mortality, or differences in food consumption were observed. No effect was seen on pregnancy or litter size but increased fetal mortality did occur, though only at a dietary level of 1000 ppm. Increased liver weight was found in all three generations. At both the 1000 and 10,000 ppm doses, reproductive toxicity was observed in the form of reduction in pup and, subsequently, in adult body weights in both sexes. Possible enlargement of the renal pelvis was seen in the high dose group. Developmental ToxicitylTeratogenesis A teratology study in rabbits had a LOEL of 200 mg/kg/day with an endpoint of increased resorptions. However, maternal toxicity was also seen at this dose. No teratogenic effects were seen.
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Biochemical Mechanism of Action Flutolanil was shown to inhibit growth of the fungal pathogen Rhizoctonia solani and to inhibit its succinate dehydrogenase complex 100% at 2 ppm in vivo, a concentration that also decreased ATP production by 50% (Hirooka et aI., 1990). At a concentration of 10 Il-M in vitro, flutolanil inhibited succinate dehydrogenase activity in mitochondria from target fungi by 80-90%, but only 15% inhibition was seen with the comparable enzyme from rat liver mitochondria (Motoba et aI., 1988). Yamano and Morita (1995) tested flutolanil against rat liver hepatocytes, mitochondria, and microsomes in vitro at 1 mM. No cytotoxicity or lipid peroxidation were observed and tissue sulfhydryl levels were not significantly affected. However, mitochondrial respiration was inhibited with a threshold (20% inhibition) at 0.1 mM. Absorption, Metabolism, and Elimination Flutolanil is rapidly metabolized and excreted in rats, mainly through oxidative removal of the isopropyl group and subsequent conjugation of the phenol (Murakami et al., 1983). Environmental Fate and Toxicity Flutolanil has a low acute toxicity to birds (Table 57.2). It also has a very low acute toxicity to most aquatic species (Table 57.2). It has a low potential for bioconcentration from water by fish. A bioconcentration factor of 20 in carp, with a plateau reached after 1 day of exposure, and a moderately high clearance rate (half-life of 5.8 days) have been reported by Tsuda et al. (1992). Flutolanil undergoes firm binding to soil colloids and has a low leaching potential. Degradation in soils is primarily by microbial action but is relatively slow with half-lives in both flooded and upland soils in the range of 160 to 320 days. Furametpyr: General Properties and Uses (RS)-5-chloroN -(1,3-dihydro-l, 1,3-trimethylisobenzofuran-4-yl)-1,3-dimethylpyrazole-4-carboxamide(Fig. 57.9), CAS Reg. No. 12357288-3, was discovered by Sumitomo Chemical Co. and is described by Oguri (1997). It exists as colorless to light brown crystals, m.p 150°C, v.p. 4.7 x 10- 6 Pa (25°), W.s. 225 ppm (25°C), log P2.36, HLC 7 x 10-6 Pam3 mol-I. Furametpyr consists of a mixture of two optical isomers. Furametpyr (S-658) is a systemic fungicide for the control of leaf sheath blight of rice under the trade name Limber. Toxicological Profile Available data are very limited. The primary source is Oguri (1997). Acute Toxicity The acute oral toxicity of furametpyr to mammals is higher than that of other members of the group, but it is still relatively modest (Table 57.2). IrritationlSensitization Furmetapyr causes slight eye irritation and skin sensitization but no skin irritation. SubchroniclChronic Toxicity No data seem to have been published regarding furametpyr's chronic or other toxic actions.
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CHAPTER 57 Pesticides Affecting Oxidative Phosphorylation
Biochemical Mechanism of Action Furametpyr inhibits succinate but not NADH oxidation in mitochondria from the fungus Rhizoctonia solani (Oguri, 1997).
significantly affected. However, mitochondrial respiration was inhibited with a threshold (20% inhibition) at 0.1 mM.
Environmental Fate and Toxicity No data were found regarding furametpyr's toxicity to birds, but its toxicity to fish is low.
Environmental Fate and Toxicity Mepronil has a very low acute toxicity to birds and aquatic vertebrates and invertebrates (Table 57.2). The reported half-life for mepronil in flooded soils ranges from 46 to 51 days.
MeproniI: General Properties and Uses 3'-Isopropoxyo-toluanilide (Fig. 57.9), CAS Reg. No. 55814-41-0, was discovered by Kumiai Chemical Industry Co. Ltd. and is described by Kawada et al. (1985). Mepronil's development is discussed by Doi (1981). It exists as white crystals, m.p. 92-93°C, v.p. 5.6 x lO- s Pa (20°C), w.s. 12.7 ppm (20°C), log P3.66. It is stable to hydrolysis over the range of pH 5 to 9 and relatively stable to photolysis with 34-62% destruction by sunlight after 80 days exposure (Yumita and Yamamoto, 1982). Mepronil (KCO-l) is a systemic fungicide with protective and curative actions in a broad range of crops against basidomycete fungi through seed treatment, soil, and foliar applications. Trade names include Basitac.
Oxycarboxin: General Properties and Uses 5,6-Dihydro2-methyl-l,4-oxathiin-3-carboxanilide 4,4-dioxide (Fig. 57.9), CAS Reg. No. 5259-88-1, was discovered by Uniroyal Chemical Co. and is described by von Schmeling and Kulka (1966). It exists as white crystals, m.p. 128-130°C, v.p. <5.6 x 10-6 Pa (25°C), w.s. 1400 ppm (25°C), log PO.772, HLC < 1 x 10-6 Pa m3 mol-I. It is stable to heat but slowly hydrolyzed (half life of several weeks) at neutral pH and room temperature. Oxycarboxin is a systemic fungicide with curative action, mainly used in foliar applications on turf and ornamentals but also used on some food crops. Trade names inelude Carbexsin, Carboject, Oxykisvax, and Plantvax.
Toxicology Profile The primary source of data is Doi (1981). Acute Toxicity The acute toxicity of mepronil to vertebrates is very low (Table 57.2). Even by intraperitoneal injection, the acute LDso in rats and mice is over 5000 mg/kg. IrritationlSensitization skin.
Mepronil is nonirritating to eyes and
Subchronic!Chronic Toxicity NOELs in 2-year feeding studies ranged from 5.9 mg/kg/day for male rats to 72.9 mg/kg/day in females. Results with mice gave NOELs of 13.7 and 17.8 mg/kg/day in males and females, respectively. MutagenicitylGenotoxicity standard tests. Teratogenicity
Mepronil is not mutagenic in
Mepronil is not teratogenic in the rat or rabbit.
Biochemical Mechanism of Action A series of substituted benzanilides, including mepronil, was shown to inhibit succinate dehydrogenase activity with considerable potency in isolated fungal mitochondria (Motoba et aI., 1988; White, 1987). Shimizu et al. (1992) also demonstrated that mepronil inhibited succinate oxidation in mitochondria isolated from the mycelia of the fungus Rhizoctonia solani with an Iso value of 0.25 f.!M. The oxidation of NADH was not inhibited. Mitochondria from other fungi, plants, and from rat and mouse liver were much less sensitive with Iso values for the vertebrate enzyme in the range of 50-100 f.!M. Yamano and Morita (1995) tested mepronil against rat liver hepatocytes, mitochondria, and microsomes in vitro at 1 mM. No cytotoxicity or lipid peroxidation were observed and tissue sulfhydryl levels were not
Toxicology Profile The primary sources of data are HSDB (2000) and Cal DFA (1987a). Acute Toxicity Oxycarboxin has a low acute toxicity to vertebrates (Table 57.2). IrritationlSensitization not a skin irritant.
Oycarboxin is a mild eye irritant but
Chronic Toxicity In 2-year dietary studies a NOEL for rats was observed at 15 mg/kg/day based on minor thyroid changes and for dogs the NOEL was 75 mg/kg/day. MutagenicitylGenotoxicity Oxycarboxin was not mutagenic in yeast and fungal assays (de Bertoldi et aI., 1980) or in the Ames assay (Moriya et aI., 1983). Reproductive Toxicity In a three-generation study in rats, no adverse reproductive effects were recorded at doses that were not toxic to the parents.
Biochemical Mechanism of Action Oxycarboxin inhibits oxidative metabolism and specifically succinate dehydrogenase activity in both fungal (Mathre, 1971; White and Thorn, 1975) and vertebrate (Gosselin et aI., 1984) mitochondria, but the potency was much higher against the enzyme from the target species (Rhizoctonia solani, Iso = 6.8 f.!M; Ustilago maydis, Iso = 8.0 f.!M) than vertebrates (rat and mouse liver, Iso = 50100 f.!M) (Shimizu et aI., 1992). It is less potent than carboxin as an inhibitor of complex 11 in both fungi and vertebrates (Mathre, 1971; Shimizu et aI., 1992; White and Thorn, 1975).
57.5 Inhibitors of Complex III Environmental Fate and Toxicity Oxycarboxin's acute toxicity to birds is relatively low (Table 57.2). In 8-day dietary studies, it also showed low toxicity to birds; bobwhite quail had an LCso of > 10,000 ppm and mallard ducks had an LCso of >4640 ppm. Oxycarboxin has a notably low acute toxicity for fish and other aquatic species (Table 57.2). The half-life of oxycarboxin in soil is 2-8 weeks with the most prominent metabolites resulting from the opening of the oxathiin ring. The corresponding aniline is also produced by amide hydrolysis. Its low log P value and high water solubility (Kd values ranging from 74 to 98) suggest that leaching in soil could occur quite readily and that bioconcentration in fish is unlikely to occur. Thiflnzamide: General Properties and Uses 2',6'-Dibromo-2-methy1-4'-triftuoromethoxy-4-trift uoromethy1-1,3-thiazole-5-carboxanilide (Fig. 57.9), CAS Reg. No. 130000-40-7, was discovered by Monsanto and developed by Rohm & Haas. It is described by (O'Reilly et aI., 1992). It exists as a white to light brown powder, m.p. 178-179°C, w.s. 1.6 ppm (20°C), log P4.1 and is stable to hydrolysis over the range pH 5-9. Thiftuzamide (MON 24000) is used as a seed and foliar fungicide on a wide range of crops and turfgrass under the trade name of Greatam. Toxicology Profile al. (1992).
The primary source of data is O'Reilly et
Acute Toxicity Thiftuzamide shows a low acute toxicity to vertebrates, (Table 57.2). IrritationlSensitization Thiftuzamide is moderately irritating to the eyes and slightly irritating to skin.
1199
changed with the introduction of a new class of broad-spectrum fungicides colloquially termed the "strobilurins" after the naturally occurring strobilurins from mushrooms that acted as their model. These are rapidly becoming a major component in fungal pathogen management programs worldwide. Remarkably, several other new fungicides, which are not in the strobilurin family, also have their toxic effect in target species through inhibition of complex Ill. Two compounds with activity as insecticides or acaricides also fall within this class. At least three groups of inhibitors of complex III have been identified (Jordan et aI., 1999a; von Jagow and Link, 1986). The first group displaces ubiquinone from the Qo binding on the Iow potential heme of cytochrome b located on the outer face of the inner mitochondrial membrane (Fig. 57.4; b, c). Several natural products including myxothiazol and the strobilurins bind at this site. A second group of inhibitors, the hydroxyquinones, acts at the Rieske iron sulfur center rather than cytochrome b (Fig. 60.45 bh). The third group, typified by antimycin A, displaces coenzyme Q from the Qi binding site located toward the inner face of the inner mitochondrial membrane by binding to the high potential heme of cytochrome b (Fig. 60.45-66). The mechanism of operation of complex III and the molecular nature of the interactions of these inhibitors with the quinone binding sites have been reviewed by Berry et al. (1999). Pesticides are now in use that affect each of these three sites. Like most recently developed pesticides, they appear to have toxicological properties that indicate a high degree of safety to terrestrial species. A greater degree of risk may exist for aquatic organisms in some cases. 57.5.2 PROPERTIES OF SPECIFIC COMPOUNDS 57.5.2.1 StrobiIurin Analogs
MutagenicitylGenotoxicity Thiftizamide gave negative results in the Ames and mouse micronucleus assays. Biochemical Mechanism of Action Thiftuzamide is a very potent inhibitor of succinate dehydrogenase from mitochondria isolated from Rhizoctonia solani with an Iso value of about 20 nM compared to 180 nM for carboxin (Phillips and RejdaHeath, 1993). Environmental Fate and Toxicity Data are unavailable for the acute toxicity to birds, but in short term studies with bobwhite quail and mallard ducks, thiftuzamide was practically nontoxic with LCso values > 5620 ppm in the diet. Its toxicity to aquatic organisms is also low (Table 57.2).
57.5 INHIBITORS OF COMPLEX III 57.5.1 INTRODUCTION Until recently, very few pesticides had complex III as their primary site of action and none had broad usage. This has now
The strobilurins are an extremely important, new, and expanding group of fungicides which were developed from the naturally occurring strobilurins and oudemansins (Fig. 57.10). These natural compounds, which contain the key methyl (E)-fJmethoxyacrylate grouping, are produced by fungi in the genera Strobilurus and Oudemansiella. They are fungicidal but too photounstable to be useful in agriculture. The discovery, development, and general properties of the commercial strobilurin fungicides have been reviewed by Beautement et al. (1991), Leroux (1996), Sauter et al. (1996, 1999), and Clough and Godfrey (1998). Important synthetic strobilurin fungicides include azoystrobin (Zeneca), pyraclostrobin (BASF), kresoxim-methyl (BASF), picoxystrobin (Zeneca), triftoxystrobin (Novartis), and metominostrobin (Shionogi). In the near future more strobilurins are likely to be developed including compounds with acaricidal (e.g., ftuacrypyrim; Nippon Soda) as well as fungicidal activity. As a group, their toxicological properties appear to be near ideal with a high potency against pathogenic fungi but presenting little risk to most other organisms. Both natural and synthetic strobilurins have their primary toxic effect by inhibition of mitochondrial respiration in complex III (Becker
1200
CHAPTER 57 Pesticides Affecting Oxidative Phosphorylation
Strobi Iurin A
fi_fi~iOJ
IN v
Kresoxim~methyI
Trifloxystrobin
Pyraclostrobin
:3:~~3
Azoxystrobin
Picoxystrobin
Metominostrobin
Figure 57.10 Pesticides that act as inhibitors of complex III-Strobilurin and the fungicides derived from it.
et a!., 1981). The specific binding site has been shown to be within the Qo domain (Fig. 57.4). Available toxicology data indicate a high to very high degree of safety to mammals and birds and several strobilurins have been classified as reduced risk pesticides by the U.S. Environmental Protection Agency. Their acute toxicity is shown in Table 57.3. This high degree of safety probably does not arise primarily from an insensitivity of the site of action in higher organisms since the mitochondrial complex III of fungi, yeast, maize, insects, and rats is strongly inhibited by a range of strobilurin analogs in vitro. Only minor species selectivity is evident except that the plant enzyme is significantly less sensitive than the others (Rohl and Sauter, 1994; Sauter et a!., 1996). The major factor in their safety appears to be differential uptake and biotransformation, particularly the rapid hydrolysis of the ester group in higher organisms which deactivates the compound as a respiratory inhibitor (Kohle et a!., 1994). This view is supported by their generally high toxicity to aquatic species, both vertebrate and invertebrate, in standard bioassays which probably results from the rapid rate of uptake of these lipophilic compounds from water and their high intrinsic activity as respiratory inhibitors. The recent development of acaricidal analogs of strobilurins again indicates that this group of compounds is capable of inhibiting complex III across a broad range of organisms. Because of their recent and continuing development, most of the toxicological information regarding these compounds must be drawn from the results of standard toxicology tests submitted
to regulatory authorities prior to registration. In these studies, the strobilurins generally show no, or only weak, evidence of mutagenicity, carcinogenicity, neurotoxicity, reproductive toxicity, or teratogenicity. An exception is kresoxim-methyl which has shown hepatic carcinogenic effects in rats at high dietary levels. Irritancy to skin and eye is either absent or mild in most cases. Trifioxystrobin is a strong dermal sensitizer but the other strobilurins do not share this characteristic. Azoxystrobin: General Properties and Uses Methyl (£)-2{2- [6-(2-cyanophenoxy )pyrimidin-4-y loxy ]phenyl} -3 -methoxyacrylate (Fig. 57.lO), CAS Reg. No. 131860-33-8, was discovered by Zeneca Agrochemicals and is described by Godwin et a!. (1992). It exists as a white powder, m.p. 116°C, v.p. 1.1 x IQ-IO Pa (25°C), w.S. 6 ppm (20°C), log P2.5, HLC 7.3 x lO-9 Pam 3 mol-I. It is table to hydrolysis. A half-life of 11-17 days was obtained in an aqueous photolytic study. Azoxystrobin (ICIA5504) is a very broad spectrum fungicide with systemic activity and both protectant and curative actions used on a broad range of crops including fruits, vegetables, small grains, and turf grass. Trade names include Abound, Amistar, Bankit, Heritage, Ortiva, Priori, and Quadris. In 1999 axozystrobin was the leading proprietary fungicide in the world with sales of $415 million (Godwin et a!., 2000). Toxicology Profile (l997b,1997c).
The primary data sources is the U.S. EPA
Acute Toxicity Azoxystrobin has a very low acute and chronic toxicity to mammals (Table 57.3). IrritationlSensitization Azoxystrobin is a slight to moderate eye irritant and a slight skin irritant. It is not a dermal sensitizer. Subchronic Toxicity Sustained dietary exposure to azoxystrobin in vertebrates causes minimal toxicity. In 90-day dietary studies the LOEL was 211 mg/kg/day for rats and 250 mg/kg/day for dogs. Chronic Toxicity In a 2-year dietary study in mice the LOEL was found at 2000 ppm (272 (M) or 363 (F) mg/kg/day) marked by decreased weight gain and food utilization. In a 2-year dietary study in rats at a dose of 82 mg/kg/day, biliary toxicity was observed in males (but not females). No significant adverse effects were observed in dogs fed at 200 mg/kg/day in the diet for 1 year. At doses of 25 mg/kg/day and above, liver changes (increased weight and serum liver enzymes) were seen, but no histopathology was evident and the response was judged to be adaptive rather than toxic. Carcinogenicity Standard lifetime dietary studies with rats (1500 ppm) and mice (2000 ppm) revealed no compoundrelated increase in tumors. The U.S. Environmental Protection Agency considered azoxystrobin "not likely to be a human carcinogen."
57.5 Inhibitors of Complex III
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Table 57.3 Acute Toxicity of Pesticides Acting as Inhibitors of Mitochondrial Complex III to Selected Nontarget Species LDso (mg/kg)
LCso (mglI)
Acute oral Compound
Rat (M; F)
Mouse (M)
Acute dermal
Inhalation Rat
Quail
Duck
Rat or rabbit
>2000
>2000
LCso (ppb)
(24-96 hr) Fisha
Daphnia
Strobilurin analogs Azoxystrobin
>5000, >5000 >5000
>2000
Kresoxim-methyl
>2150
Metominostrobin
>5000 776;708
Picoxystrobin
>5000
Pyraclostrobin
>5000
>2000
Trifloxystrobin
>5000
>2000
>2250
>2000
>2000
>2000 >2000
1778
0.96 >5.6
470-1600
>1.88
190-499 17,500-22,500 65-75
>2000
>2.12
>2000
>0.31 <1.07
>2000
>4.65
>2000
>0.84
0.2-77 14-78
259 186 14,000-22,300 18 15.7 25
Other pesticides Acequinocyl
>5000
Cyazofamid
>5000
Famoxadone
>5000
Fenamidone
>5000;2028 817; 1502
Hydramethylnon
>5000
>2000 >5000
>2250 >2000 1828
>2510
>33,000
16
>70,500
>2000
>5.3
>2000
>2.1
>5000
2.9
11-49
12
740 90-1700
1140
QRange of values from several species, most commonly including the rainbow trout and bluegiII.
MutagenicitylGenotoxicity Evidence regarding the genotoxicity of azoxystrobin is mixed. Negative results were obtained in the Ames, mouse micronucleus, and unscheduled DNA synthesis assays in rats, but the compound was weakly positive for forward mutations in the mouse lymphoma cell assay and for chromosomal aberrations in human lymphocytes in vitro. Since no clastogenic effects were seen in vivo, azoxystrobin was determined not to be genotoxic. Reproductive Toxicity No adverse reproductive effects were observed in a multigenerational study in rats. The NOEL for toxicity was found at 300 ppm (32 mg/kg/day) based on liver changes and body weight reductions in the parents. Biliary toxicity, again confined to adult males, was observed at 200 ppm in the form of hyperplasia and inflammation of the lining of the bile duct and hepatic proliferative cholangitis. Developmental ToxicitylTeratogenicity No developmental effects were seen in rabbits at doses as high as 500 mg/kg/day. In rats, no developmental effects were seen at doses that were not also maternally toxic. No evidence of teratogenicity was noted in either study. Neurotoxicity None was seen with a single dose of 200 mg/kg in rats or with subchronic dosage at 38.5 mg/kg/day.
Biochemical Mechanism of Action The inhibitory effect on complex III is typical of other strobilurins and is specifically described by Wiggins and Jager (1994). Absorption, Metabolism, and Elimination After an oral dose, azoxystrobin is well absorbed in the rat. It is extensively metabolized and rapidly excreted, particularly in the
feces, with the primary metabolic route being ester hydrolysis. Glutathione-S-transferase-catalyzed conjugations with glutathione also occur at several sites in the molecule in vitro with the major site of attack being the pyrimidine ring, but these conjugates were not observed in vivo (Turner and Joseph, 1998). Percutaneous absorption is minimal. Environmental Fate and Toxicity This aspect of azoxystrobin's properties has been reviewed by Pilling et al. (1996) and D.S. EPA (l997b). It has a very low toxicity to birds, but it is highly toxic to both freshwater fish and invertebrates (Table 57.3). Feeding bobwhite quail and mallard ducks for 8 days with azoxystrobin in the diet gave LC50 values over 5200 mg/kg/day in each case, and a chronic reproduction study in mallards had a LOEC at 3000 ppm. Its toxicity to estuarine and marine invertebrates is variable from moderate to very high based on studies with mysid shrimps (acute LC50 of 56 ppb) and pacific oyster larvae (acute LC50 of 1300 ppb). The chronic effects on aquatic invertebrates are also potent. An early lifestage chronic toxicity test in the fathead minnow gave a LOEC of 193 ppb, and a life-cycle toxicity study in Daphnia had a LOEC of 84 ppb. In aerobic and anaerobic soil metabolism studies in the laboratory, a half-life of 72 to164 days was obtained, indicating a moderate degree of persistence. However, in field studies, azoxystrobin is degraded quite rapidly in soils with a half-life of 1 to 5 weeks. Both photodegradation and, to a lesser degree, microbial metabolism are involved in its dissipation since the compound is relatively stable to hydrolysis. The half-life for photolysis in soils is in the range of 11 to 15 days. With a typical Koc of about 500 (range 300 to 1690) and Kd values in the range of 1.5 to 23 ml/g, depending on nature of the soil, azoxy-
1202
CHAPTER 57 Pesticides Affecting Oxidative Phosphorylation
strobin has a low to moderate potential for mobility in soil, but little movement down the soil profile was seen in field studies. Kresoxim-Methyl: General Properties and Uses Methyl (E)- 2-methoxyimino-[2-(o-tolyloxymethyl)phenyl] acetate (Fig. 57.10), CAS Reg. No. 143390-89-0, was discovered by BASF AG and is described by Ammermann et al. (1992). Its development and structure-activity relations are reviewed by Sauter et al. (1996). It exists as white crystals, m.p. 97-102°C, v.p. 2.3 x 10-6 Pa (20°C), W.s. 2 ppm (20°C), log P3.4, HLC 3.6 x 10-4 Pa m3 mol-i. It is slowly hydrolyzed in water at pH 7 (half-life 34 days) but much faster at pH 9 (half-life 7 hr). Kresoxim-methyl (BAS 490F) is a long-lasting broad spectrum fungicide with protective and curative capabilities used on fruit, cereals, and greenhouse ornamentals. Trade names include Alliage, Candit, Cygnus, Discus, Mantra, Mentor, Sovran, and Stroby. Toxicology Profile The primary sources of data are U.S. EPA (1998b), FAO/WHO (1998), and Cal EPA (1999). Acute Toxicity Kresoxim-methyl has a very low acute toxicity to mammals (Table 57.3). No clinical signs of poisoning were observed in rats at doses of 5000 mg/kg. IrritationlSensitization Kresoxim-methyl is not a dermal irritant or sensitizer, but it causes slight eye irritation. Subchronic Toxicity In 90-day dietary studies, azoxystrobin caused minimal toxicity at high doses. In mice the only effect seen at dietary levels up to 8000 ppm [1900 (M) and 2600 (F) mg/kg/day] was an adaptive increase in liver weight. In rats, the LOEL was 580 mg/kg/day based on decreased weight gain and changes in clinical chemistry parameters. In dogs, the LOEL was found at 25,000 ppm (770 mg/kg/day) based on vomiting and reduced body weight. Dermal application in rats at 1000 mg/kg/day for 21 days caused no observable effects. Chronic Toxicity In chronic studies, LOEL values in 1 to 2 year feeding studies in mice, rats, and dogs were about 350, 440, and 735 mg/kg/day, respectively. Effects noted include decreased food intake and body weight gain and the occurrence of microscopic liver and biliary alterations in rats. Histopathological examination revealed alterations in liver (amyloidosis) and kidney (renal papilliary necrosis) in mice. No notable adverse effects were seen in dogs. Carcinogenicity Liver carcinoma was observed in the lifetime rat feeding study in both sexes at 8000 ppm [370 (M) and 500 (F) mg/kg/day]. The findings included increased hepatocellular carcinoma and neoplastic changes in the biliary system. An extensive series of mechanistic studies in rats clearly indicate that kresoxim-methyl at this dose does not cause an increase in liver cell foci but does cause reversible hepatocellular proliferation. It therefore appears that it acts as a tumor
promoter rather than an initiator, and it may therefore have a threshold for its tumorigenic action. It certainly appears that kresoxim-methyl is not genotoxic (below). Considerable discussion of this issue is provided in Cal EPA (1999). No carcinogenic response was seen in parallel lifetime feeding studies in mice. Kresoxim-methyl has been given an interim carcinogen classification of category C (possible human carcinogen) by the U.S. Environmental Protection Agency. MutagenicitylGenotoxicity Kresoxim-methyl was uniformly negative in an extensive battery of standard tests for mutagenic and clastogenic activity using microbial and mammalian cells in vitro and in vivo studies with mammals. Reproductive Toxicity No effects on reproduction were observed in rats in a two-generation study with doses up to 1000 ppm in the diet. At 4000 ppm, reduced weight gain in pups was observed but this dose was also maternotoxic. Developmental ToxicitylTeratogenicity In the two-generation reproductive toxicity study in rats, decreased pinna unfolding occurred at 4000 ppm. An increase in incomplete vertebral ossification was found in rats at a dose of 1000 mg/kg/day during gestation which was not toxic to the mothers. No developmental or teratogenic effects were observed in rabbits given doses up to 1000 mg/kg/day. Neurotoxicity A single dose of 2000 mg/kg failed to induce any behavioral changes or neuropathology in rats. Subchronic (90-day) feeding studies at doses up to 16,000 ppm also proved negative for neurotoxicity.
Biochemical Mechanism of Action Using cytochrome redshift analysis, Rohl (1994) showed that kresoxim-methyl binds to the hCJ complex of yeast mitochondria with high affinity (Kd = 70 nM) and a single binding site. Rohl and Sauter (1994) showed that differences in the sensitivities of complex III from a broad range of species [yeast, Botrytis cinerea, corn (maize), house fly, and rat] to kresoxim-methyl and related strobilurins were relatively small, though the rat enzyme tended to be less sensitive than that from Botrytis. In the case ofkresoxim-methyl this interspecific difference in Iso values was about IO-fold. The corn enzyme was the least sensitive overall. Although, as these authors stress, the quantitative comparison of Iso values obtained with different enzyme preparations must be treated with caution, this suggests that much of the very high selective toxicity of kresoxim-methyl must reside in pharmacokinetic factors. This was confirmed by Sauter et al. (1995) who conclude that rapid degradation by hydrolysis is a critical factor governing the low toxicity of this compound to vertebrates. Absorption, Metabolism, and Elimination Kresoxim-methyl is poorly absorbed after an oral dose in rats with 50% uptake at 50 mg/kg and 25% at 500 mg/kg. The absorbed material is quickly eliminated in urine (20 to 30%) and feces (70 to 80%) with extensive metabolism. Peak plasma levels were
57.5 Inhibitors of Complex III
1203
reached in 0.5 to 1 hr and 8 hr at the low and high dose, respectively. Plasma half-lives range from 16.9 to 30.5 hr, depending on the dose. Ester hydrolysis is the first major metabolic step, but there are multiple sites of biotransformation on the molecule and 34 metabolites were identified from rats. Ester hydrolysis by plasma from rats and mice was very rapid (half-life of 2 to 5 min). The metabolism of kresoxim-methyl by hepatocytes from a range of vertebrate species is described and compared to its metabolism by these species in vivo by Salmon and Kohl (1996).
through an independent lead from screening synthetic chemicals (Kataoka et aI., 1998). It thus represents an example of convergent evolution in chemical discovery. It is also differs from other stobilurins in being an amide rather than an ester. Despite these differences, its mode of action is the same as that of the other strobilurins involving inhibition of respiration by binding to the cytochrome bCj complex (Furuta, 1999; Mizutani et aI., 1996). It is proposed that, as part of its action, the inhibition of complex III induces the production of cytotoxic superoxide radical anions in target fungi (Furuta, 1999).
Environmental Fate and Toxicity Kresoxim-methyl has a low toxicity to birds. It is highly toxic to aquatic vertebrates and invertebrates in standard laboratory tests (Table 57.3), but studies in more complex ecosystems indicate that it is not hazardous to aquatic organisms, when used as recommended, due to its rapid breakdown. Kresoxim-methyl is rapidly degraded in soils with a half-life of 1-2 days. The primary metabolite is the free acid. Its moderate soil absorption (Koc 219-372) suggests the possibility of significant mobility in soil.
Environmental Fate and Toxicity Data on the acute toxicity to birds were not found, but in an 8-day dietary study in mallard ducks the LCso value was >5200 ppm, indicating a low degree of senstivity to this compound. The low toxicity to aquatic species in standard tests is very notable compared to that of the other strobilurins so far disclosed (Table 57.3). The half-life in aerobic soil is 98 days, indicating a moderate level of environmental persistence.
Metominostrobin:
General
Properties
and
Uses
(E)- 2-methoxyimino- N -methyl-2-(2-phenoxyphenyl)acetami-
Picoxystrobin: General Properties and Uses Methyl (E)-2{2- [6-(trifiuoromethyl)pyridin-2-yloxymethy l]-phenyl} -3-methoxyacrylate (Fig. 57.10), CAS Reg. No. 117428-22-5, was discovered by Zeneca Agrochemicals and is described by Godwin et al. (2000). It exists as a solid, m.p. 75°C, v.p. 5.5 x 10-6 Pa (20°C), log P3.6, w.s. 3.1 ppm (20°C). Picoxystrobin is degraded fairly rapidly in water with a half-life of 7-15 days. Picoxystrobin (ZEN 90160) is being developed as a broad spectrum fungicide with preventative and curative actions specifically intended for use on cereals.
de (Fig. 57.10), CAS Reg. No. 133408-50-1, was discovered by Shionogi Co. Ltd. and is described by Furuta (1999). Structurefungicidal activity relationships in this series are described by Kataoka et al. (1998). It exists as a white crystalline powder, m.p. 87-89°C, v.p. 1.8 x 10- 5 Pa (25°C), w.s. 128 ppm (20°C), log P2.32. Metominostrobin exists as geometrical isomers at the oxime moiety. The E-isomer is 5- to 20-fold more fungicidal than the Z-isomer and the commercial product consists only of the E-form (Kataoka et aI., 1998). Metominostrobin (SSF-126) is registered in several Asian countries as a systemic long-lasting preventative and curative agent against rice blast under the trade name of Oribright.
Acute Toxicity Picoxystrobin has a very low acute toxicity to the rat by the oral, dermal, and inhalation routes (Table 57.3).
Toxicology Profile The primary source of data is Furuta (1999).
IrritationlSensitization Picoxystrobin is not a skin or eye irritant or a dermal sensitizer.
Acute Toxicity The acute toxicity to rats and mice is somewhat higher than with other strobilurin fungicides but it is still quite low (Table 57.3).
Carcinogenicity No evidence of carcinogencity was seen in lifetime feeding studies in rats and mice.
IrritationlSensitization The formulated material causes slight eye irritation, but it is not a skin irritant or sensitizer.
Toxicology Profile The primary source of data is Godwin et al. (2000). Only an outline of the toxicology data for picoxystrobin is currently available.
Genotoxicity tery of tests.
No genotoxicity was observed in a standard bat-
Reproductive Toxicity No reproductive toxicity was observed in a multigenerational study in rats.
MutagenicitylGenotoxicity Metominostrobin was negative in a battery of four genotoxicity tests involving mutagenic and clastogenic endpoints in bacteria and mammals.
Developmental ToxicitylTeratogenicity No adverse effects were observed in developmental studies in the rabbit and rat.
Biochemical Mechanism of Action Metominostrobin differs from the other members of this class in two ways. First it was not developed using the natural strobilurins as a model but
Absorption, Metabolism, and Elimination Studies in rats showed that picoxystrobin is well absorbed, extensively metabolized, and rapidly eliminated. Dermal absorption is low.
1204
CHAPTER 57
Pesticides Affecting Oxidative Phosphorylation
Environmental Fate and Toxicity
No acute toxicity data for birds were found, but short-term and subchronic feeding studies with birds revealed low toxicity (e.g., it had a 21-week NOEL of 1350 mg/kg/day in mallard ducks). Picoxystrobin is highly toxic to aquatic species in laboratory bioassays (Table 57.3). However, studies under normal use conditions in the field indicated a low level of risk to aquatic species in practice. The BCF in fish is 290, indicating a low to moderate bioaccumulation capability. Picoxystrobin degrades rapidly in the soil with a half-life of 3-35 days under field conditions. It is quite firmly bound to soil components (Koc 790-1200), indicating a low potential for leaching.
Pyraciostrobin: General Properties and Uses Methyl N(2- {[1-(4-chlorophenyl)-1 H -pyrazol-3-yl]oxymethyl }phenyl)N-methoxycarbamate (Fig. 57.10), CAS Reg. No. 175013-18-0, was discovered by BASF Corp. and is described by Ammermann et al. (2000). It exists as white to light beige crystals, m.p. 64-65 DC, v.p. 2.6 x 10-8 Pa (20DC), log P3.99, w.S. 1.99 ppm (20DC), HLC 5.3 x 10-6 Pam3 mol- 1 . It is stable to hydrolysis at pH 5-7 (half-life >30 days) but is quite rapidly photolyzed with a halflife less than 2 hr. Pyraclostrobin (BAS 500 F) is a very broad spectrum fungicide with preventative and curative properties and a long residual action. It affects all stages of fungal growth and has uses on fruits, vegetables, small grains, and grasses. Trade names include Cabrio EG and Headline.
Toxicology Profile Primary sources of data are Ammermann et al. (2000) and Cal EPA (2001). Acute Toxicity Pyraclostrobin has a very low oral toxicity to mammals though it is somewhat more toxic by inhalation (Ta-
Chronic Toxicity No major adverse effects were seen in either rats or dogs in typical chronic dietary studies. In the 2-year study in rats, changes in liver-derived serum enzymes and a slight decrease in body weight were observed at 200 ppm. In a second study to determine carcinogenicity, hepatocelluar necrosis was thought to be treatment-related in rats at 200 ppm. In dogs the LOEL was obtained at 400 ppm (11 mg/kg/day) based on increased diarrhea and lowered weight gains. Carcinogenicity No evidence for carcinogenicity was obtained in long-term rat and mouse dietary studies. MutagenicitylGenotoxicity Pyraclostrobin was not mutagenic or clastogenic in a battery of five typical tests in microbial and mammalian systems. Reproductive Toxicity No results of toxicological concern were seen in rats in a two-generation dietary study with doses up to 300 ppm. Developmental ToxicitylTeratogenicity Developmental toxicity studies were negative in the rat and rabbit. Early resorptions were seen in rabbits at higher doses (5 mg/kg/day or greater) but these doses also elicited maternal toxicity. No teratogenic effects were observed. Neurotoxicity No evidence for neurotoxicity was discovered in a battery of neurobehavioral and neuropathological assays in rats at acute oral doses up to 1000 mg/kg, or in a 90-day test at 50 mg/kg/day (males) or 112 mg/kg/day (females).
Biochemical Mechanism of Action Pyraclostrobin inhibits complex III from yeast and fungus with ICso values of 2029 nM (Ammermann et aI., 2000).
ble 57.3).
Absorption, Metabolism, and Elimination Only about 50% IrritationlSensitization Pyraclostrobin causes slight eye and moderate skin irritation. It is not a dermal sensitizer. Subchronic Toxicity Hyperplasia of the duodenal mucosa and increases in liver and spleen weights were observed in rats fed pyraclostrobin at 500 ppm and above in a 90-day dietary study. Histological findings at doses as low as 150 ppm (11.7 mg/kg/day) included hepatocellular hypertrophy, extramedullary hematopoiesis, distension of the sinusoids, and histocytosis of the spleen. These effects were the basis for setting the LOEL. In a 90-day dietary study in dogs, hypertrophy of the duodenal mucosa was seen at 450 ppm (13.3 mg/kg/day). Reduced boy weights and changes in blood chemistry also were seen at this dose. Mice, in a 90-day feeding study, also showed thickening of the duodenal mucosa together with erosion or ulcers in the glandular stomach and a decrease in lipid vacuolization in the adrenal cortex. Females were more sensitive than males with adrenal effects occurring at 50 ppm (12.9 mg/kg/day).
of an oral dose of either 5 or 50 mg/kg pyraclostrobin was absorbed in rats. The absorbed dose was rapidly metabolized and eliminated in the bile (33% of the dose) and urine (10-15% of the dose). Major routes of metabolism involved demethoxylation and hydroxylation of the pyrazole and other ring systems followed by glucuronidation. No sign oftissue bioaccumulation was observed.
Environmental Fate and Toxicity Pyraclostrobin has a low toxicity to birds, but it shows an extremely high level of toxicity to fish and other aquatic species in laboratory tests (Table 57.3). The toxicity to estuarine and marine invertebrates is equally high. Mysid shrimps have a 96-hr LCso of 4.2 ppb and the ECso for the inhibition of shell deposition in oysters is 12.5 ppb. However, it is claimed that in practice, risk to aquatic species is likely to be minimal due to the low levels of exposure. Pyraclostrobin is rapidly degraded in aerobic soils (the half-lives in the field range from 2 to 37 days). It is immobile in soil (Koc 6000-16,000 ml/g) and therefore unlikely to leach.
57.5 Inhibitors of Complex III Triftoxystrobin: General Properties and Vses
Methyl
(E)-methoxyimino-{ (E)-a-[I-(a, a, a-triftuoro-m-tolyl)ethyl-
idene-aminooxy]-o-tolyl} acetate (Fig. 57.10), CAS Reg. No. 141517-21-7, was discovered by Novartis Crop Protection, Inc. and is described by Margot et al. (1998). It exists as a white powder, m.p. 73°c, v.p. 3.4 x 10-6 Pa (25°C), W.s. 0.61 ppm (25°C), log P4.5, HLC 2.3 x 10- 3 Pam3 mol-I. Triftoxystrobin has two oxime linkages, each showing geometrical isomerism. The commercial compound is the E, E-isomer. The hydrolysis half-life is 11.4 weeks at pH 7 and 27 hr at pH 9. It is stable at pH 5. Triftoxystrobin (CGA-279202) is a very broad spectrum foliar fungicide registered for use on fruits, vegetables, turf, and ornamentals. Trade names include Compass, Flint, and Stratego. Toxicology Profile The primary sources of data are Margot et al. (1998), V.S. EPA (1999b), Anonymous (2000a), and Cal EPA (2000a).
Acute Toxicity Triftoxystrobin shows a very low acute toxicity to mammals and birds (Table 57.3). IrritationlSensitization Triftoxystrobin is a strong dermal sensitizer in the maximization test in guinea pigs but not in the Beuhler test. It is a mild eye and skin irritant. Subchronic Toxicity In 90-day feeding studies, no remarkable findings were made in mice at 500 ppm but microscopic abnormalities in liver and spleen were observed at 7000 ppm (1275-1650 mg/kg/day). Blood chemistry and organ weights were affected in dogs at 500 mg/kg /day. In rats, pancreatic hypertrophy was the most sensitive endpoint, seen at 2000 ppm (130 mg/kg/day). Chronic Toxicity In chronic dietary studies, mice fed 1000 ppm in diet for 18 months developed hepatocellular hypertrophy and focal liver necrosis. Rats showed reduced weight gain at 750 ppm and this was associated with a decrease in tumor incidence which was attributed to the decreased body weight. In a I-year study in dogs, clinical signs, increased liver weights, and hepatocellular hypertrophy were noted at 50 mg/kg/day. Carcinogenicity Triftoxystrobin is regarded by the V.S. Environmental Protection Agency as "unlikely to be a human carcinogen" based on negative studies in rats and mice. MutagenicitylGenotoxicity No adverse effects were observed in a standard battery of tests for clastogenicity and DNA repair. An increased mutation rate was observed in Chinese Hamster V79 cells, but triftoxystrobin was negative for mutagenicity in the Ames Salmonella test. Reproductive Toxicity In a two-generation study in rats, no reproductive effects were observed at the highest dose, 1500ppm.
1205
Developmental ToxicitylTeratogenicity In a 2-year dietary study in rats, the most sensitive effect was a reduction in pup weight gain at 750 ppm. Teratology studies in with dietary exposure of pregnant rats and rabbits gave no indication of higher sensitivity to triftoxystrobin in utero compared to maternal effects. The only effect noted was enlargement of the thymus gland in rats at 1000 mg/kg. Neurotoxicity No neurotoxicity was observed in rats at single oral dose of 2000 mg/kg in a functional observation test battery. Biochemical Mechanism of Action Like other strobilurins, triftoxystrobin inhibits respiration at complex III (Margot et al., 1998). Absorption, Metabolism, and Elimination In rats, goats, and poultry, triftoxystrobin was poorly absorbed from the gastrointestinal tract and much was excreted in the feces unchanged. In rats, the absorbed compound was rapidly cleared (tissue half-life of 13 to 42 hr) with extensive metabolism, particularly through hydrolysis of the ester group. Other significant metabolic routes were O-demethylation of the methoxyimino group and oxidation ofthe methyl side chain to the corresponding alcohol and carboxylic acid. Percutaneous absorption is very slight. Environmental Fate and Toxicity Triftoxystrobin has a low acute toxicity to birds (Table 57.3). Prolonged dietary exposure at levels of 320 ppm (bobwhite quail) or 500 ppm had no effects on health or reproduction. It is extremely toxic to fish and several kinds of aquatic invertebrates. The LCso to mysid shrimp is 9 ppb (Anonymous, 2000a). Bioconcentration in fish is limited, probably due to rapid metabolism and depuration. Triftoxystrobin is rapidly degraded in surface water by microbial action with a half-life of a few hours. The photolysis half-life in water is 31.5 hr with the initial reaction being E to Z isomerization. Triftoxystrobin binds firmly to soil particles (log Koc 16423745 in five soils) and its half-life in soils in field studies ranges from a 1.9 to 16 (mean 5.4) days, indicating a low environmental persistence and leaching potential (Anonymous, 2000a). 57.5.2.2 Other Pesticidal Inhibitors of Complex III Cyazofamid, famoxadone, and fenamidone are three other recent fungicides which inhibit mitochondrial respiration at complex III but that fall outside the strobilurin chemical class. Famoxadone and fenamidone, having obvious structural similarities and a similar spectrum of fungicidal activity, form a subclass of their own. They act at the same general biochemical site as the strobilurins. Cyazofamid is in a separate structural class and inhibits complex III at a different site than the compounds so far discussed. Two pesticides acting on invertebrates, acequinoyl (an acaricide) and hydramethylnon (an insecticide), are also thought to have their primary pesticidal actions as complex III inhibitors.
1206
CHAPTER 57
Pesticides Affecting Oxidative Phosphorylation
differs from that of other recent fungicidal inhibitors of complex III which act at the Qo site (Fig. 57.4). It is claimed to be specific in its inhibitory action to mitochondria from oomycete fungi (Mitani et aI., 1998).
Fenami done
Famoxadone
Acequinocyl
Cyazofami d
CF
6~_
d
N_t]
-)=N - N H
CF~ Figure 57.11 hibitors.
H
Hydramethylnon
Pesticides that act as inhibitors of complex Ill-Other in-
Cyazofamid: General Properties and Uses 4-Chloro2 - cyano - N, N - dimethy 1- 5 - p - toly limidazole - 1- sulfonamide (Fig. 57.11), CAS Reg. No. 120116-88-3, was discovered by Ishihara Sangyo Kaisha, Ltd. and is described by Mitani et al. (1998). It exists as an ivory powder, m.p. 153°C, v.p. < 1.3 x 10-5 Pa, w.s. 0.12 ppm (20°C, pH 5), log P3.2 (25°C). Cyazofamid (IKF-916) is a new systemic foliar and soil fungicide. Trade names include Docious, Mildust, and Ranman. Toxicology Profile The primary source of data, which are so far very limited, is Mitani et al. (1998). Acute Toxicity Cyazofamid appears to have a low acute toxicity to vertebrates (Table 57.3).
Environmental Fate and Toxicity Too few data are yet available on cyazofamid's acute effects on birds and fish to judge its level of toxicity to these organisms, but limited data for fish (carp) indicate a very low level of toxicity (Table 57.3). With a Koc value of 490-6300 and a half-life in soil of 4-5 days, cyazofamid binds firmly to soils and undergoes rapid environmental degradation (Mitani et aI., 1998). Famoxadone: General Properties and Uses 3-Anilino5 -methyl-5 -(4-phenoxyphenyl)-I, 3 -oxazolidine- 2 ,4-dione (Fig. 57.11), CAS Reg. No. 131807-57-3, was discovered by DuPont Agricultural Products and is described by Joshi and Sternberg (1996). Its discovery, development, and structureactivity relations are discussed by Sternberg et al. (2001). It exists as a pale cream powder, m.p. 140-142°C, v.p. 6.4 x 10-7 Pa (20°C), W.s. 52 ppb (20°C), log P4.65 (pH 7), HLC 4.6 x 10- 3 Pam3 mol- i (20°C). The hydrolytic half-life is 2 days at pH 7. Hydrolysis is more rapid under alkaline conditions and slower at acidic pH (Jernberg and Lee, 1999). Famoxadone exists as a pair of optical isomers. The S-isomer is significantly more fungitoxic than the R-isomer. Famoxadone (DPX-JE874) is a new broad spectrum fungicide for vegetables, fruits, and cereals. It has both preventative and curative capability. Trade names include Charisma and Famoxate. Toxicology Profile The primary source of data is Anonymous (2000b). Acute Toxicity Famoxadone is practically nontoxic to mammals by all routes of exposure (Table 57.3). IrritationlSensitization Famoxadone is a minimal skin and eye irritant. It does not cause dermal sensitization. Subchronic Toxicity NOELs of 40, 200, and 350 ppm were obtained in 90-day dietary studies in dogs, rats, and mice, respectively.
MutagenicitylGenotoxicity Cyazofamid is negative in the Ames bacterial mutagenesis test.
Chronic Toxicity The long-term (1-2 year) dietary NOELs were also 200 ppm in rats and 40 ppm in dogs. A NOEL of 700 ppm was established in mice. A NOEL of 100 mg/kg/day was observed when famoxadone was given to monkeys by gavage for 1 year. Effects seen in these studies at higher doses (LOELs) included hepatotoxicity in rats and mice and hemolysis in rats and monkeys. Lens opacities were seen in dogs fed 300 ppm for 1 year.
Biochemical Mechanism of Action Cyazofamid is reported to inhibit complex III but at the Qi site (antimycin site) which
Carcinogenicity No carcinogenic response was observed in long-term feeding studies in rats and mice.
IrritationlSensitization
Data are not available.
SubchroniclChronic Toxicity Cyazofamid caused no toxic effects in dogs fed at 1000 mg/kg/day in a 90-day study.
57.5 Inhibitors of Complex III
MutagenicitylGenotoxicity Famoxadone was positive for chromosomal aberrations in an in vitro test, but negative in bacterial and mammalian mutation and unscheduled DNA synthesis tests in vitro, in a mouse micronucleus test, and in an unscheduled DNA synthesis tests in vivo. The weight of evidence therefore suggest that famoxadone is not genotoxic. Reproductive Toxicity No reproductive effects were seen in a two-generation study in rats at doses below those causing maternal toxicity. Pup weight was decreased at 800 ppm. Developmental ToxicitylTeratogenicity No teratogenic or developmental effects were seen in rats or rabbits at the highest doses tested (1000 mg/kg/day). Neurotoxicity No neurotoxic effects were observed in rats with dietary exposure at 800 ppm for 90 days.
Biochemical Mechanism of Action Studies establishing that famoxadone is an extremely potent Inhibitor of mitochondrial electron transport acting at complex III in mitochondria from fungi, plants, and mammals are reviewed by Jordan et al. (1999a, b). The site of inhibition is cytochrome b within the Qo domain which prevents the transfer of electrons from cytochrome b to cytochrome er (Fig. 57.4). The site appears to be close to, but not fully identical with, the binding sites on cytochrome b of myxothiazol on the one hand and the strobilurins on the other. However, field isolates of the fungus Mycosphaerella fijiensis having a point mutation in the target site that caused resistance to strobilurin fungicides also showed cross-resistance to famoxadone (Sierotzki et aI., 2000). Famoxodone shows little selectivity at the target site level since Iso values for electron transport in rat and beef heart submitochondrial particles are in the range of 10-20 nM. These are somewhat lower than the Iso values for several comparable preparations from fungi. The S( -) isomer is approximately lOO-fold more active as an inhibitor than the R ( +) isomer and 25- to 30-fold more active in disease control (Jordan et aI., 1999a). Famoxadone is notably more potent as an inhibitor of complex III from rat heart and fungal mitochondria than azoxystrobin and kresoxim-methyl (Anonymous, 2000b). Absorption, Metabolism, and Elimination After an oral dose famoxadone is poorly absorbed and rapidly excreted in mammals, mainly in the feces. Hydroxylation of the two phenyl rings at the para position is the major routes of metabolism (Lee et aI., 1999). Environmental Fate and Toxicity Famoxadone has a low acute toxicity to birds (Table 57.3). Dietary studies in mallard ducks and quail (probably over 8 days) also indicated a low toxicity with LCso values > 5620 ppm. In standard laboratory assays, the acute toxicity to aquatic species is very high, including both vertebrates and invertebrates (Table 57.3). Daphnia, oysters, and mysid shrimp all have 48-96 hr ECsos of 1.4 to 12 ppb. It is believed that these potentially negative effects are
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likely to be minimized in practice by the rapid breakdown of famoxadone in water and other environmental media and its strong binding to particulates (Jernberg and Lee, 1999). With an average Koc of 3740 mUg and a half-life in soils ranging from 6 to 12 days (Jernberg and Lee, 1999), famoxadone degrades rapidly and is unlikely to leach into groundwater. The half-life in the water phase of natural systems is less than 0.5 hr. Fenamidone: General Properties and Uses S-5-methyl-2methylthio-5-phenyl-3-phenylamino-3,5-dihydro-4H-imidazol4-one (Fig. 57.11), CAS Reg. No. 161326-34-7, was discovered by Rhone-Poulenc Agro. and is described by Mercer et al. (1998). It exists as a white woolly powder, m.p. 137°C, v.p. 3.4 x 10-7 Pa (25°C), w.s. 7.8 ppm (20°C), log P2.8. Only the S-isomer is active and the commercial product is stereospecific for this isomer. Fenamidone (RPA 407213) is a potent fungicide from a new chemical class primarily active against downy mildews. It has both curative and protective actions. Trade names include Fenomen and Reason. Toxicology Profile The primary sources of data are Mercer et al. (1998) and Anonymous (1999b). Acute Toxicity Fenamidone has a low to very low acute toxicity to mammals (Table 57.3). IrritationlSensitization Fenamidone is not a skin or eye irritant nor is it a dermal sensitizer. MutagenicitylGenotoxicity Fenamidone is negative in the Ames and mouse micronucleus assays. Developmental ToxicitylTeratogenicity teratogenic in rats and rabbits.
Fenamidone is non-
Biochemical Mechanism of Action Fenamidone inhibits electron transport in complex Ill. The exact site of action within the complex is unknown. With the S-enantiomer, the Iso value for the inhibition of respiration in mushroom mitochondria is 0.56 ).!M. The R-enantiomer is much less active (Mercer et aI., 1998). Environmental Fate and Toxicity The acute toxicity of fenamidone to birds is low (Table 57.3) and 8-day dietary studies with bobwhite quail and mallard ducks also reveal a very low toxicity with an LCso > 5200 ppm (Mercer et aI., 1998). The toxicity to fish is high (Table 57.3) but it is claimed that fenamidone's impact on aquatic species is likely to be low when it is used according to the label (Anonymous, 1999b). The soil half-life is short « 10 days). Fenamidone is bound to soil fairly strongly (Koc > 400) and leaching is unlikely.
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Acequinocyl: General Properties and Uses 3-Dodecyl-2hydroxy-l A-naphthoquinone acetate (Fig. 57.11), CAS Reg. No. 57960-19-7, was discovered by DuPont and is under development by Agro-Kanesho Co. Ltd. and Tomen Agro. It is described by Kinoshita et al. (1999) and Wakasa and Watanabe (1999). It exists as a fine yellow powder, m.p. 60°C, v.p. 5.18 x 10-5 Pa (40°C), W.s. 6.7 ppb (25°C), log P > 6.2. Hydrolysis half-life (pH 7, 25°C) is 53 hr. Acequinocyl is less stable at higher pH (half-life is 76 min at pH 9) and more stable under acidic conditions (half-life is 86 days at pH 4). Acequinocyl is a highly lipophilic naphthoquinone derivative. Similar compounds (2-alkyl-3-hydroxy-lA-naphthoquinones) have been known for a considerable time to have a variety of pesticidal activities including antibacterial, acaricidal, antimalarial, fungicidal, and insecticidal actions (e.g., see Jacobsen and Pedersen, 1986). Acequinocyl was patented as a pesticide by E. I du Pont de Nemours & Co. (Bellina and Fost, 1977), but it was not immediately developed. Closely related compounds, showing good activity against sucking insects and mites, have recently been described based on 2-alkyl3-hydroxy-l A-naphthoquinone derivatives with unsaturated, branched alkyl chains that were isolated from a Chilean plant, Caleeolaria andina (Khambay et aI., 1997). These, too, are under evaluation as commercial pesticides. Acequinocyl (DPX-3792, AKD-2023, TM-413) is a specific acaricide currently under development in many countries (Kinosh ita et aI., 1999) under the trade name Kanemite. Toxicology Profile Watanabe (1999).
The primary source of data is Wakasa and
Acute Toxicity Acequinocyl shows an extremely low level of acute toxicity to vertebrates (Table 57.3). IrritationlSensitization It has very slight activity as a dermal and eye irritant and is not a dermal sensitizer. Carcinogenicity It showed no evidence of carcinogenicity in rats and mice in chronic dietary studies. MutagenicitylGenotoxicity Acequinocyl was negative in a range of standard tests (Ames Salmonella assay, DNA repair, chromosomal aberrations). Reproductive Toxicity No notable reproductive toxicity was observed in rats in a multigenerational study. Developmental ToxicitylTeratogenicity Acequinocyl caused no adverse effects in rats and rabbits in studies of developmental toxicity. Biochemical Mechanism of Action The de-acetylated metabolite of acequinocyl, 2-hydroxy-3-dodecyl-lA-naphthoquinone, strongly inhibits respiration in complex III of insect mitochondria (Koura et aI., 1998). Compounds in this class
have long been known to be complex III inhibitors in vertebrate systems (von Jagow and Link, 1986). Structurally they resemble ubiquinone and they bind at the Qo center, probably competing with ubiquinone for this site. They differ from the strobilurin like inhibitors by acting to prevent electron transfer to cytochrome Cl and the consequent reoxidation of the Rieske iron-sulfur center (Fig. 57.4). Their inhibitory potency depends on their lipophilicity and the optimal alkyl chain length is reached with the undecyl group. Acequinocyl itself is initially inactive as a respiratory inhibitor using mitochondria in vitro, but inhibition develops slowly with time (Koura et aI., 1998). The hydrolytic conversion of acequinocyl to the deacetylated metabolite has been shown to occur in isolated mitochondria (Koura et aI., 1998; Rich, 1996) and it seems very likely that this metabolite is the active toxicant and that acequinocyl is a propesticide. It has no significant effects on complex I. As described in Section 57.1.6.4, some naphthoquinones can act as electron acceptors from the mitochondrial electron transport chain leading to redox cycling and serious oxidative stress. It is not known whether this occurs with acequinocyl or its active metabolite, but this effect was not seen with the closely related naphthoquinones developed by Khambay (Khambay, 1998). Environmental Fate and Toxicity Acequinocyl has a low acute toxicity to birds and its toxicity to fish is extremely low (Table 57.3). In a striking contrast, its toxicity to Daphnia is extremely high (Wakasa and Watanabe, 1999).The bioconcentration factor in fish (carp) is moderate (170-387) despite acequinocyl's lipophilicity. Acequinocyl is rapidly degraded by soil microorganism with a soil half-life of less than 3 days. Because of its very high lipophilicity and the correspondingly high Koc value that ranges from 33,900 to 123,000, it is firmly bound to soil organic matter and has a very low leaching potential. Hydramethylnon: General Properties and Uses 5,5-Dimethy lperhydropyrimidin-2-one 4-trifluoromethy I-a -( 4-trifluoromethylstyryl) cinnamylidenehydrazone (Fig. 57.11), CAS Reg. No. 67485-29-4, was discovered by American Cyanamid Co. and is described by Lovell (1979). It exists as yellow to tan crystals, m.p. 189-191°C, v.p. 3.7 x 10-6 Pa (25°C), W.s. 7 to 9 ppb (25°C), log P4.45 (U.S. EPA, 1998c) [also reported as 2.31 (Tomlin, 2000)], HLC 0.78 Pa m3 mol- 1 (25°C). It is relatively stable to hydrolysis over the pH range 5 to 9. Hydramethylnon (AC 217,300) is the sole member of the class of amidinohydrazine insecticides. It is used as a slow acting bait for the agricultural and household control of ants, cockroaches, and termites. Trade names include Amdro, Combat, Maxforce, and Siege. Toxicology Profile The primary data sources are U.S. EPA (l998c) and Cal EPA (2000b). Acute Toxicity As shown in Table 57.3, hydramethylnon has a low or moderate toxicity for terrestrial vertebrates. The acute
57.6 Inhibitors of ATP Synthase
clinical signs of poisoning in rats include hyopactivity, diuresis, ataxia, anorexia, epistaxis, chromodacryorrhea, and salivation. IrritationlSensitization Hydramethylnon is a moderate eye irritant but it is not a skin irritant or sensitizer. SubchronicJChronic Toxicity Decreased food consumption and wasting leading to death were seen in dogs fed at 6 mg/kg/day for 3 months. Similar but less severe effects were also seen in rats in a 90-day feeding study. In chronic dietary studies, rats developed glomerulonephrosis at 25-50 ppm. Testicular atrophy was observed at 50 ppm and higher doses which was not accompanied by any general decrease in body weight. Carcinogenicity Hydramethylnon is classified by the U.S. Environmental Protection Agency as a group C carcinogen (possible human carcinogen) due to an increase in lung adenomas and carcinomas in female mice dosed at 4.45 mg/kg/day. Much discussion has ensued regarding the possible confounding effects of Sendai virus infections on this diagnosis (Cal EPA, 2000b). No carcinogenic responses was observed in male mice or rats of either sex. MutagenicitylGenotoxicity Hydramethylnon was negative in a typical battery of five tests involving mutation in microorganisms and chromosomal aberrations and dominant lethality in mammals. Reproductive Toxicity Testicular atrophy was seen in a number of subchronic and chronic feeding studies with rats, mice, and dogs. Histopathologic findings included aspermia hypospermia, interstitial cell hyperplasia of Leydig cells, and germinal cell degeneration. This led to decreased male reproductive success in a two-generation study in rats (LOEL of 3.32 mglkglday). Hydramethylnon is therefore a general reproductive toxicant that specifically targets germinal cells of the testis. Developmental ToxicitylTeratogenicity Hydramethylnon caused no developmental effects in rats and rabbits at doses that did not also cause maternal toxicity. Immunotoxicity When Amdro was fed to calves at 113.5 g/calf/day, leukopenia was evident after 2 weeks with decreased numbers of lymphocytes and eosinophils. However, no evidence was seen for a depression of cellular immunity (Evans eta/., 1984).
Biochemical Mechanism of Action Hydramethylnon inhibits respiration in insects in vivo, resulting in behavioral depression and eventual paralysis similar to that caused by rotenone and antimycin A. It also depresses respiration in Chinese hamster ovary cells in vitro with a 24-hr LCso of about 1 !lM. It inhibits respiration at micromolar concentrations in isolated rat liver mitochondria acting specifically at complex III (Hollingshaus, 1987). Inhibition is slow in developing and the
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specific binding location within the complex is not known. Hydramethylnon is reported to have a disordering effect on model lipid bilayer membranes (Hollingshaus, 1987) but the role of such effects in its inhibitory actions on complex III are unclear. In its slowly developing inhibitory action on mitochondrial respiration in vitro and its ability to disrupt membrane structure, as well as in aspects of its structure, hydramethy Inon resembles some alkylguanidines and biguanidines used in the treatment of diabetes mellitus and as antimicrobial agents (Schiifer, 1981). Absorption, Metabolism, and Elimination After oral administration in rats, hydramethylnon is rapidly eliminated, almost entirely as the unchanged compound in the feces. Absorbed material was metabolized relatively slowly. Dermal absorption is very limited. Environmental Fate and Toxicity Hydramethylnon's acute toxicity to birds is moderate to low (Table 57.3). Its acute toxicity to aquatic species is quite variable. It is of relatively low toxicity to some fish and Daphnia (Table 57.3) but it is highly toxic to other fish species (e.g., channel catfish have a 96-hr LCso value of 90 ppb). Because of its limited use patterns, strong binding to soil components, and rapid photodegradation it is unlikely that hydramethylnon presents a significant threat to any aquatic species in practice. It shows some tendency to bioconcentrate in fish (BCF is 1300 for the whole body) and depuration is relatively slow with a half-life of about 14 days. Hydramethylnon is rapidly degraded in the environment by photolysis (the half-life in water is less than or about 1 hr). The half-life in soil is 3-55 days in field studies but degradation is biphasic with a much slower second phase. Hydramethylnon binds tightly to soil (Kd values of 1039-1782 mUg) and it is unlikely to leach into ground water.
57.6 INHIBITORS OF ATP SYNTHASE 57.6.1 INTRODUCTION In view of the large and exquisitely complex machinery that is responsible for the synthesis of ATP from ADP and inorganic phosphate, and its critical role in cellular bioenergetics, it is perhaps surprising that no large number of poisons are known that act at this site. Only two groups of pesticides are strong inhibitors of this complex with evidence that this inhibition plays an important role in their toxic actions. These are the organotins and diafenthiuron, an insecticide that generates a reactive carbodiimide inhibitor. The inclusion of the organotin pesticides in this section is a matter of convenience rather than an indication that complex V is known to be the major target for their toxic effects. In fact, they have several effects on mitochondria and oxphos in addition to inhibiting complex V, and they have a number of potent effects on nonmitochondrial biochemistry also. It is often unclear which of these effects, if any, predominates under given conditions in vivo and thus could be
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regarded as the most significant biochemical lesion underlying an adverse response. Several other pesticides show considerable activity against this site in vitro, but the evidence that it is their primary site action in vivo is lacking. These compounds include two chlorinated diphenyl sulfone acaricides, chlorfenson and tetradifon. These are reported to be potent (submicromolar or low micromolar) inhibitors of ATP synthase from mites, fish, and mammals in vitro (Bustamente and Pedersen, 1973; Cutcomp et aI., 1972) although they were less active in other reports (Abo-Khatwa and Hollingworth, 1974; Kadir and Knowles, 1991). Kadir and Knowles (1991), in a limited study, found that a number of other specific halogenated acaricides including chloropropylate, bromopropylate, flubenzimine, chlorfenson, and propargite inhibited oligomycin-sensitive Mg2+ -ATPase from the bulb mite in vitro with a potency greater than that of the organotin cyhexatin. Of these compounds, only tetradifon, bromopropylate, and propargite are still use (Tomlin, 2000). There appears to be no evidence regarding the possible occurrence and significance of this inhibition of complex V in vivo either in mites or vertebrates, but it may be considered as a plausible target site for these compounds. 57.6.2 ORGANOTIN PESTICIDES 57.6.2.1 History and Uses Since the 1950s, organotins have been widely and increasingly used as pesticides with multiple applications as fungicides, acaricides, molluscicides, wood preservatives, disinfectants, rodent repellants, and antifouling agents. All pesticidal organtins are trisubstituted with aromatic or aliphatic groups such as phenyl or butyl moieties. The less toxic dialkyl and monoalkyltins have important industrial uses as stabilizers and catalysts in plastic production (Fent, 1996a) but are not employed as pesticides. Trimethyl and triethyltins are too toxic to both plants and vertebrates to be utilized as pesticides. However, organotins with longer alkyl chains (tributyltin, fenbutatin-oxide), aromatic rings [triphenyltin (fentin)], or alicyclic groups (azocyclotin, cyhexatin) have proved to be effective for a variety of pesticidal purposes. Each of these triorganotins may occur in several forms which vary in the fourth group coordinated with the tin atom (e.g., they can exist and be utilized as hydroxides, chlorides, acetates, or as other forms). As reviewed by Fent (1996a) many of these compounds have been plagued by problems with chronic toxicity which have led to the cancellation and limitation of uses. Most notably, the use of tributyltins (and some triphenyltins) as antifouling agents in paints for ships' hulls became widespread in the 1970s. This has led to significant environmental effects on aquatic species in and around harbors and coastal areas due to the leaching of the tin compounds from the paint. Because they have long half-lives in sediments, are lipophilic, which leads to ready bioaccumulation by biota, and many aquatic species are highly sensitivity to their toxicological effects, population reductions have been widely observed. This is particularly notable with
mollusks, which may be severely affected at concentrations below 10 ppb. Because of these impacts, restrictions were placed on this use of tributyltin by many nations in the 1980s (Fent, 1996a). However, these measures have not been universally effective in reducing organotin levels or their biological adverse effects, and a complete international ban on this use of organotin biocides is now being debated and would take effect in 2003. Much less is known regarding the occurrence and toxicological impact of organotins in wastewater and sewage sludge, although tributyl- and triphenyltin can be detected in sludge in the concentration range of 0.1 to 1.0 ppm (Fent, 1996b), and this merits additional study (Fent, 1996a, 1996b). 57.6.2.2 Chemical Properties The chemistry of bioactive organotins has been reviewed in detail by Fent (1996a). Organotins have ionic character and tend to dissociate in water, releasing the corresponding substituted tin cation. The pKa of tributyltin hydroxide is approximately 6.4, meaning that at physiological pH it is predominantly but not wholly in the form of the undissociated hydroxide. Other pesticidal organotins hydroxides have similar pKa values in the 5.0 to 6.5 range. Forms other than the hydroxide, such as tin oxides, acetates, or chlorides, tend to hydrolyze to form the corresponding hydroxide in aqueous solutions, but the rate varies depending on the solution conditions. The aqueous speciation of tributyl- and triphenyltins and its relationship to n-octanol partition coefficients have been studied by Amold et al. (1997) who developed a simple model of the effects of pH and salt concentration on their partitioning behavior. The speciation of these compounds therefore is complex depending on the starting compound, pH, temperature, and the presence of counterions in solution (e.g., in seawater with its high chloride ion level, the formation of the organotin chlorides is favored). In turn speciation governs water solubility, lipophilicity and thus partitioning into lipoprotein membranes, rates of uptake into living organisms, and strength of binding to soil colloids. Tributyl- and triphenyltins have a water solubility minimum in the pH range 6 to 8, and solubility decreases with increasing salinity, so that the solubility in seawater is only 40% of that in distilled water at pH 7 to 8 (Inaba et aI., 1995). The undissociated organotin molecule, which is generally more lipophilic than the corresponding cation, will be more effective in each of the processes above (e.g., the octanol: water partition coefficient of tributyltin chloride increases from 3.25 at pH 6.0 to 3.9 at pH 8.0 as ionization is suppressed, and its bioaccumulation by, and toxicity to, water fleas and carp increases correspondingly). Similar results showing an increase in bioaccumulation by several aquatic organisms with increasing water pH have been reported with tripheny ltin (Fent, 1996a; Looser et aI., 1998). It is worth noting that the organotin cations themselves are relatively lipophilic, so that uptake into living organisms does not cease at lower pHs (Looser et al., 1998), and their lipophilicity will tend to increase with the increasing carbon atom content of the attached groups. Because of their more or less ready conversion in solution to a mixture of the hydroxide and cationic forms, any differences
57.6 Inhibitors of ATP Synthase in toxicological effects between organotins sharing the same organic substituents but with a different fourth coordinated group (e.g., tributyltin hydroxide and tributytin chloride) will tend to be small. The major difference will arise primarily from the effect of the fourth subsitituent on their lipophilicity and initial bioavailability. Other than undergoing hydrolytic conversion to the hydroxide form, pesticidal organotins are relatively unreactive. The conclusion that triorganic tins are only weakly interactive with sulfhydryl groups (Aldridge and Cremer, 1955), has been disputed by Byington and his colleagues (e.g., see Wulf and Byington, 1975). Under some circumstances organotin cations have the capability to coordinate with proteins, particularly those with sulfhydryl groups [e.g., triethyltin forms a pentacoordinate complex with hemoglobin through its cysteine 13a and histidine 20a (Taketa et aI., 1980)]. A significant observation is that, like arsenite, the organotins bind much more avidly to vicinal dithiol groups than monothiols (Stridh et aI., 1999b). This also true of dialktltins (Aldridge, 1976). In at least two cases, dithiothreitol, a dithiol, has been found to be active in antagonizing organotin inhibition of enzymes while monothiols such as glutathione were inactive (Nebbia et aI., 1999; Rao et aI., 1987). The inhibition of several key enzymes by organotins has been attributed to sulfhydryl binding (see below). 57.6.2.3 General Toxicology
The literature addressing the toxicology and environmental impact of tributyltins and, to a lesser degree, triphenyltins, is substantial. Much less attention has been paid to the organotins that are primarily of interest in agriculture, such as cyhexatin, and, particularly, azocyclotin and fenbutatin-oxide. Since the tributyltins have already been reviewed in depth on a number of recent occasions (Benya, 1997; Boyer, 1989; Meador, 2000; Tanabe, 1999) they are not included here specifically, but reference is made to those aspects of their toxicology that can throw additional light on the other members of the group. Reviews of the general toxicology of tin compounds include Hall and Pinkney (1985) focusing on their aquatic toxicology and Fent (l996a) focusing on their general and ecotoxicology. The toxicity of organotins has a strong tendency to follow the order trisubstituted > disubstituted > monosubstituted and this sequence is also the general pathway of metabolism for pesticidal organotins in both vertebrates and the environment. However, some disubstituted tin metabolites do show high toxicity which must be considered in evaluation of the overall safety of the parent trisubstituted compound. In experimental animals, organotins have a general ability to cause cytotoxicity in many tissues and organs including skin, intestine, kidney, liver, lung, and brain. The impacted organ systems tend to be those most directly exposed by the route of administration. These results are similar to observations with human exposures in which dermal, hepatotoxic, nephrotoxic, and neurological effects have been recorded. These broadly toxic actions in vivo are paralleled in vitro by the high toxicity of trisubstituted organotins to isolated cells from many sources.
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Evidence of cytotoxicity may be observed at aqueous concentrations in the general range of 10 nM to 1 ).LM (Fent, 1996a). The toxicology of individual compounds is presented below, but it is clear that many organotin pesticides as a group present several specific toxicological hazards at relatively low doses.
1. They have endocrine modulatory characteristics as indicated by their effects on male reproduction in mammals (reduced testicular size, reduced spermatazoon production) and their ability to cause sex alterations in mollusks. In this case, female mollusks develop male characteristics, a process termed imposex. Their potency in the latter effect is very high with masculinization of females occurring at a water concentration of just a few ppt. The biochemical mechanism(s) for these effects is not known with certainty, but there are a large number of studies relating to the imposex phenomenon, which are briefly reviewed here. Changes in androgenJestrogen levels in mollusks undergoing organotin-induced imposex have been observed in several studies (e.g., see Bettin et aI., 1996; Morcillo et aI., 1998; Spooner et aI., 1991). The role of aromatases, which are cytochrome P-450-catalyzed enzymes responsible for the conversion of testosterone to estradiol, in controlling the androgenJestrogen ratio has been studied. Evidence has been presented both in favor (Bettin et aI., 1996; Morcillo et aI., 1998) and against (Morcillo and Porte, 1999) the hypothesis that estradiollevels are decreased and testosterone levels increased specifically because organotins inhibit aromatase action in vivo. Other possibilities to explain imposex include an effect on the enzymes that degrade sex hormones and excretory transport systems (Ronis and Mason, 1996), and a direct stimulatory effect on androgen receptors. In this regard, Yamabe et al. (2000) examined the effect of tributyl- and triphenyltin on androgen-dependent transcription and cell proliferation in human prostate cancer cells. Stimulation of androgen-dependent gene transcription and cell proliferation occurred at concentrations oftriphenyltin as low as 1 nM. They concluded that trisubstituted organotins activate androgen-receptor-mediated transcription but act at a target site other than the hormone-binding site of the receptor. Whether such direct androgen receptor effects occur in mollusks is unknown. 2. Many organotins are clearly immunotoxic, both in vitro and in vivo, in a range of vertebrate species including mammals and fish (Boyer, 1989; Fent, 1996a). In rats, atrophy of the thymus is observed, antibody responses are reduced, and resistance to disease is diminished (Snoeij et al., 1989). In fish, and several marine invertebrates, decreased phagocytic activity and resistance to microbial pathogens were observed at environmentally relevant concentrations of tributyltin (Cooper et aI., 1995; Wishkovsky et aI., 1989). Many studies have shown that tributyltins cause thymus atrophy and T-cell immunodeficiency in vivo (reviewed in Benya, 1997). This, at least in part, is related to the ability of organotins to initiate apoptosis in thymocytes which has been reported in many studies both in vitro, at low micromolar concentrations are
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effective (e.g., see Aw et aI., 1990; Bollo et al., 1996; Oyama et aI., 1991; Raffray and Cohen, 1991), and in vivo (see Boyer, 1989; Nishida et aI., 1990). Tributyltin is significantly more potent than triphenyltin in this thymolytic action (Aw et aI., 1990). Both the T-cell-dependent humoral and cellular immune responses are inhibited. In addition to effects on T-cells, organotins also have deleterious effects on human B lymphocytes derived from tonsil tissue, including the induction of apoptosis at in vitro concentrations of 100 nM (De Santiago and Aguilar-Santelises, 1999). Tributyltin has been shown to inhibit the tumor killing activity of human natural killer lymphocytes at sub micro molar concentrations in vitro (Whalen et aI., 1999), and triphenyltin is only slightly less potent than tributyltin in this action (Whalen et aI., 2000). If this effect also occurs in vivo it would be expected to increase sensitivity to viruses and tumor formation. In fact, triphenyltin is a tumorigen (Section 57.6.2.6), and dietary tributyltin is suspected of increasing susceptibility to infections and causing deaths in several mammalian wildlife species (Kannan et aI., 1997,1998). In a comparative study, triphenyltin and cyhexatin were most active in decreasing spleen weight in mice dosed orally, while tributylin oxide and fenburatin-oxide were least active (Pieper and Casida, 1976). 3. Pesticidal organotins as a group are generally not genotoxic, although some have been shown to be clastogenic (or co-clastogenic) in vitro and in vivo (Sasaki et aI., 1993; D.S. EPA, 1999c; Yamada and Sasaki, 1993), nor have they been found to be carcinogenic. Fentin is the exception to this rule. 4. They tend to be strong skin and eye irritants. Skin contact may cause local bums and dermatitis with pruritus and pustular outbreaks. Contact with concentrated materials on the eye can cause acute conjunctivitis and corneal opacity that can be irreversible. Their respiratory toxicity tends to be high and, after inhalation, severe irritation of the airways and the pulmonary tissues may occur. By the oral route, inflammation of the gastric mucose and ulceration can occur. 5. The aquatic toxicology of most organotins gives particular cause for concern (Fent, 1996a; Tanabe, 1999). Not only are they highly toxic by acute exposure with LC50 values around 1-10 ppb, but they have cumulative properties which means that, on more prolonged exposure, toxic endpoints can be seen at ppt concentrations. Additionally, their endocrine disruptive and immunotoxic effects occur in aquatic species also. As already described, these have given rise to well-documented, significant, and widespread impacts on aquatic species and communities which are very slow to reverse because of the persistence of these compounds in sediments and because of their continued input into the aquatic environment. Much less seems to have been done to study their potential impact on terrestrial nontarget organisms. 57.6.2.4 Human Toxicology
The signs, symptoms, and treatment of organotin poisoning are discussed by Reigart and Roberts (1999). Absorption of
organotins, either through the skin or gastrointestinal tract, is generally limited and relatively few poisonings and no deaths have been reported as a result of occupational exposure to organotin pesticides. The primary effects are on the central nervous system and include headache, nausea, vomiting, dizziness, mental disturbances, photophobia, and, sometimes, convulsions. Epigastric pain and elevation of blood sugar have been observed. There are no specific antidotes and treatment involves decontamination and supportive therapies. A number of studies of human illnesses, some of them serious, after occupational exposure or suicide attempts have been reported. Since they all involve triphenyltin acetate, they will be discussed in the section on fentin (Section 57.6.2.6). The irritant properties of organotins also cause illnesses. Inhalation exposure to latex paint containing bis(tributyltin) oxide used on internal walls caused mucous membrane irritation (Wax and Dockstader, 1995), and contact dermatitis among painters has been attributed to this compound in paints (Goh, 1985). Additional examples of human pathology arising from organotin exposure are provided by Boyer (1989). The general public is exposed to organotin pesticides through the diet. Organotin residues in fish for human consumption occur at typical levels of 0.01-1.0 ppm, primarily as tributyl- and triphenyltin. Levels in shellfish may be higher. Generally consumption is below current acceptable daily intakes (ADls) set with immunotoxicity as the endpoint (Cardwell et aI., 1999; Fent, 1996a). However the margins of safety below the AOI are not large and ADIs are subject to revision in the light of new knowledge. Consequently, high levels of consumption of fish and mollusks from contaminated areas may be of some toxicological concern. A recent study of blood levels of butyltins in residents of Michigan gave a peak value of 100 ng/ml. It was concluded that these levels were below those needed to affect human natural killer lymphocytes, a sensitive end point, but that sporadic incidences of higher exposure were possible that could have toxicological impacts (Kannan et aI., 1999). 57.6.2.5 Biochemical Mechanisms of Toxicity
Pesticidal organotins have several potential biochemical mechanisms of toxicity which may vary in relative importance between different compounds or be evident at different concentrations in vitro. These have been reviewed by Boyer (1989) and Fent (1996a). Since there may be overlaps in the potencies with which individual organotin compounds cause these different effects, it is generally difficult to determine which is the critical biochemical lesion for a given response and, in some cases, it seems reasonable to suppose that several relevant biochemical effects may occur concurrently.
1. Organotins inhibit mitochondrial ATP synthase (complex V), often with considerable potency. Tricyclohexyltin has been reported to inhibit mitochondrial oligomycin-sensitive Mg2+ -ATPase from several nonmammalian sources with 150 values of 0.1-100 nM (Desaiah et aI., 1973; Mehrotra et aI., 1985; Pieper and Casida,
57.6 Inhibitors of ATP Synthase 1965). This exceeds the sensitivity shown by typical mammalian enzymes (1-10 J.l.M; Aldridge, 1976) and would help explain the selective toxicity of miticidal organotins such as cyhexatin and azocyclotin for invertebrates. In general agreement with these results, triphenyl-, tricyclohexyl-, and tribenzyltin derivatives were found to inhibit ATP synthase activity at low micromolar concentrations but they did not inhibit electron transport or cause uncoupling in rat and plant mitochondria (Chandra et al., 1989). Orangotins inhibit ATPase activity in mitochondria only when the entire FoF! complex is present. The dissociated F! component also has ATPase activity, but this is not susceptible to inhibition by oligomycin or organotins, and inhibition of proton conductivity persists even after its removal (Gould, 1976; Papa et aI., 1982). This indicates that the triorganotins inhibit ATP synthase by interaction with the membrane-bound Fo component. Studies by Matsuno-Yagi and Hatefi (1993a, 1993b) confirm that inhibition occurs in the Fo segment of the complex and suggest that it may be due to freezing the ATPase structure in such a way as to inhibit the rotary motion required for rapid proton flux and ATP synthesis. However, the site of action of triorganotins is different from that of other Fo inhibitors such as oligomycin and dicyclohexylcarbodiimide (Papa et aI., 1982; Matsuno-Yagi and Hatefi, 1993a). The inhibitory potency toward ATP synthase tends to increase with the increasing size (and lipophilicity) of the organic substituents (Aldridge et aI., 1977). 2. Since both the hydroxide and chloride forms of triorganic tins are lipophilic and dissociable, they can locate within the lipoprotein inner mitochondrial membrane and, in the presence of chloride ions, shuttle OH- ions outward in exchange for inward-moving CI- ions. This tends to destroy the pH gradient created by the electron transport chain and represents a form of uncoupling, which increases mitochondrial respiratory rates and causes mitochondrial swelling (Selwyn, 1976; Wulf and Byington, 1975). However, this effect does not approach that of 2,4-DNP in the degree of stimulation of respiration (Aldridge et al., 1977). The CI- IOH- exchange diminishes the chemical portion of the protonmotive force (/1p) that is created by the differential pH, but it is electroneutral and does not alter the more important electrical portion of the mitochondrial membrane potential (/1 \{fm). The potency of organotins in this effect peaks when the organic substituents are propyl groups and decreases as the size of the substituents increases further (Aldridge et al., 1977). A variant of this mechanism has been proposed by Bragadin and Marton (1997, Fig. 60.6). They provide evidence that trialkyltins enter the mitochondrial inner membrane as the tin cation, depending on their lipophilicities. The cations are driven to the inner membrane face by the membrane potential where they pick up a hydroxyl ion to form the hydroxide. This is followed by the outward diffusion of the hydroxide to the outer membrane face where, at the lower pH there, it dissociates again. This tin cation/tin hydroxide cycling is no longer an electoneutral process and results in the discharge of
1213
the membrane potential as well as the proton gradient. In this context, it is notable that tributyltin acts as a mitochondrial uncoupler even in a chloride-free medium which proves that a mechanism other than 0- IOH- exchange is operating (Bragadin et aI., 2000). 3. Mitochondrial swelling is observed even in the absence of chloride ion in the medium which indicates an additional and independent mechanism from that related to CI- IOHexchange. Potency here is somewhat lower than for the first two effects but increases rapidly with the size of the organic substituents (Aldridge et aI., 1977). Bragadin and Marton (1997) and Bragadin et al. (2000) conclude that, in this case, organotins are causing the opening of the mitochondrial permeability transition pore. In turn this leads to rapid and severe swelling. In the case of fentin, the opening of the pore has been attributed to a special mechanism involving the interaction of the organotin with sulfhydryl groups since it is antagonized by sulfhydryl agents but not by cyclosporin A, a compound which reverses the effects of many compounds which induce the permeability transition (Zazueta et aI., 1994). These three actions on mitochondria have been reviewed by Aldridge (1976) and Selwyn (1976). Their relative contributions vary with structure [e.g., for short chain trialkyltins tins the most potent action is the facilitation of CI- IOH- exchange, whereas for the compounds with longer chains or aryl substituents which are used as pesticides, the potency is approximately the same for all three effects (Aldridge et aI., 1977)]. Because of this, with pesticidal tributyl and triphenyltins, no respiratory stimulation attributable to the CI- IOH- exchange is seen because it is overridden by the inhibitory effects on ATP synthase. Little further elaboration of these conclusions has been made in the last 25 years, and it is not clear which (if any) of these mitochondrial actions predominates as the ultimate cause of toxic effects in vivo. 4. Organotins inhibit other ATPases such as Na+ /K+ -ATPase and Ca 2+ -ATPase (Mehrotra et al., 1985; Sahib and Desaiah, 1986). Inhibition of the latter enzyme, which occurs with appoteeny comparable to the inhibition of the mitochondrial Mg2+ -ATPase, could contiribute to the ability of organotins to disrupt calcium regulation within cells. The potency in the inhibition of the Na+ IK+ -ATPase is generally lower. S. Organotins are lipophilic and amphipathic and can act as detergents with general membrane-disruptant properties at concentrations a low as 1-10 J.l.M. Tributyltin disrupts the erythrocyte membrane structure causing hemolysis at about 5 J.l.M (Gray et aI., 1987) and studies with planar bilayer lipid membranes (Radecka et aI., 1999) reveal that the membrane depolarization was greatest with trialkyltins and was directly related to their lipophilicity. Langner et al. (1998) compared tetra-, tri-, and diphenyltin chlorides for their actions on lipid bilayer membranes made from phosphatidylcholine and for their hemolytic potencies. Tetraphenyltin had little interaction with the model membrane, diphenyltin located toward its center, and triphenyltin oriented more to the surface in the
1214
CHAPTER 57
Pesticides Affecting Oxidative Phosphorylation
headgroup region. The order of hemolytic potencies was tri > di > tetra, and these differences in potency were explained on the basis of the different locations of these compounds within the bilayer membrane. 6. Membrane permeability and calcium regulation are affected in cells in vitro leading to an influx of calcium, a decrease in transmembrane potential, activation of NF-kappa B, and efflux of cellular proteins including cytochrome c and tumor necrosis factor-a (Aw et aI., 1990; Marinovich et aI., 1996b; Oyama et al., 1991; Stridh et aI., 1999a; Zazueta et aI., 1994). These events underlie the well-established ability of organotins to cause apoptosis in some cell types (e.g., thymocytes) at concentrations lower than those which cause necrosis (Aw et aI., 1990; Raffray and Cohen, 1991). An apoptotic response was observed in cultured mouse thymocytes at a concentrations of triphenyltin acetate as low as 12 nM (Bollo et aI., 1996). Interestingly, the increase in membrane permeability does not seem to involve the typical activation of the mitochondrial membrane permeability transition (Stridh et aI., 1999b; Zazueta et aI., 1994). Current knowledge regarding the mechanism by which organotins induce apoptosis has been reviewed by Robertson and Orrenius (2000) but much remains to be learned about this complex web of events. 7. Tributyl and triphenyltins cause microtubule disassembly and disruption of the mitotic spindle and other elements of the cytoskeleton (Chow and Orrenius, 1994; Jensen et aI., 1991; Marinovich et aI., 1996b). This effect increases with increasing lipophilicity of the tin compound (Jensen et aI., 1991) and may involve thiol group modification by the organotins (Chow and Orrenius, 1994). The mechanism of this effect is unclear since triphenyltin did not have a direct effect on actin polymerization in a cell-free system and it may occur indirectly through effects on the cellular environment (Marinovich et aI., 1996a). 8. Triorganotins inhibit several important sulfhydryl-dependant enzymes which has been attributed to their coordination with protein sufhydryls and other functional groups. These enzymes include Na+ ,K+ -ATPase (Rao et aI., 1987), caspases involved in apoptosis (Stridh et aI., 1999b), glutathione S-aryltransferase (Henry and Byington, 1976), hemoglobin (Santroni et aI., 1997; Taketa et al., 1980), and cytochrome P450-dependent monooxygenases (Fent, 1996a; Nebbia et aI., 1999). 9. Organotins appear to be able to decrease the activity of cytochrome P-450-dependent monoxygenase systems in several ways, including both direct interactions with the enzyme proteins and indirect effects through modulation of their synthesis and degradation. Rosenberg et al. (1981) in a study of the effects of tricyclohexyltin and tributyltin on heme biosynthesis in rats discovered that they increase heme degradation and decrease its synthesis. The major effect is an enhancement in degradation of heme through the induction of the activity of he me oxygenase (Fent and Stegeman, 1993; Rosenberg and Kappas, 1989). This oxygenase is the rate-limiting enzyme in heme degradation and its activity was
increased at a subcutaneous dose of cyhexatin as low as 1.9 mg/kg. This is reflected in a decrease in level and activity of cytochrome P-450 isoforms. The level of inhibition of monooxygenase activity was more than 50% at a dose of 15 mg/kg. The maximal effect occurred at 48 to 72 hr after dosing and the response was prolonged for at least 10 days. The sensitivity to this action seems to vary among species and, perhaps, among organotin agents. No effect was seen on P-450 activity in rabbits and lambs after the sublethal administration of fentin acetate (Nebbia et aI., 1997), but the effect has been confirmed in a number of species [e.g., with tributyltin in fish by Fent and Stegeman (1993) and in the dogwhelk by Spooner et al. (1991)]. Organotins (tributyl, tricyclohexyl, and triphenyl) also act directly to decrease the level of cytochrome P-450 and increase the level of its inactive P-420 form in vitro in microsomes from both mammals and fish. Corresponding losses occur in oxidative activity. This destructive effect is greatest on the ,B-naphthoflavone-inducible CYP1A family (Fent and Buche1i, 1994; Fent et aI., 1998; Kim et aI., 1998; Rosenberg and Drummond, 1983; Rosenberg et aI., 1981). Fent et al. (1998) also reported that the P-450 reducing enzyme, NAD(P)H : cytochrome c reductase, was significantly inhibited in fish exposed to triphenyltin in vivo. The effect on P-450 may occur through interaction with sulfhydryl and/or histidine residues (Fent and Bucheli, 1994; Nebbia et aI., 1999). Studies with fish hepatoma cells by Briischweiler et al. (1996) indicated that the effect of organotins on P-450 levels and activity occurs by a direct interaction with the apoprotein and not through an effect on the synthesis of the protein. However, inhibition occurs only at relatively high concentrations (0.1 to 1.0 mM) and its significance in vivo is unclear. In one case it was concluded that the concentration needed to inactivate P-450 in vitro in liver microsomes from marine mammals was more than 10-fold higher than the levels occurring in the liver in vivo (Kim et at., 1998) and Fent et al. (1998) concluded that inactivation of the CYP1A in the scup, a marine fish, occurred at doses in vivo that would be reached only under conditions of very high contamination with triphenyltin, but that inhibition on NAD(P)H cytochrome c reductase occurred at doses that are more environmentally realistic. The potential toxicological implications of cytochrome P-450 inactivation are considerable but not fully investigated. The possible role of aromatase inhibition in causing imposex has already been described. Cytochrome P-450 monooxygenases are also key enzymes in the biotransformations (both activative and inactivative) of xenobiotics, and self-synergism of organotins seems possible since their degradation occurs primarily through P-450-catalyzed oxidations. 10. Enzymes of xenobiotic biotransformation in addtion to monooxygenase are also affected by organotins in vitro and in vivo. In particular, organotins are very effective inhibitors of glutathione-S-transferases (Henry and Byington, 1976). The inhibitory potency may be high. Triphenylltin inhibited a glutathione-S-transferase from the chicken with an ICso value
57.6 Inhibitors of ATP Synthase
of 100 nM (Thomson et aI., 1998). Triphenyltin was also a strong inhibitor of transferase activity in fish in vitro (George and Buchanan, 1990). Inhibition also occurs in vivo. The effect on glutathione-S-transferase in the liver after subchronic oral exposure to triphenyltin acetate in rabbits was biphasic with an increase at 15 ppm in the diet but a significant decrease at 150 ppm (Di Simplicio et aI., 2000), but it remains to be shown that the inhibition of these enzymes has important physiological consequences. They are an important element in the degradation of electrophilic xenobiotics, and their inhibition could lead to enhanced sensitivity to such agents, but also they play a role in some biosynthetie processes in the cell [e.g., in the conversions of prostaglandins (Thomson et aI., 1998)].
57.6.2.6 Properties of Specific Compounds Azocyclotin: General Properties and Uses Tri(cyclohexyl)IH-l,2,4-triazol-l-yltin (Fig. 57.12), CAS Reg. No. 4108311-8, was discovered by Bayer AG and is described by Kolbe (1977). It exists as colorless crystals, decomposing at 219°C, v.p. 6 x 10- 11 Pa (25°C), W.s. 0.12 ppm (20°C), log P5.3, HLC 7 x 10- 2 Pam3 mol- 1 (20°C), pKa 5.36. In living systems or under alkaline conditions, azocyclotin is readily hydrolyzed to yield tricycohexyltin hydroxide, a related organotin acaricide (Section 57.6.2.6). Chemical hydrolysis is slow at neutral pH (half-life of 81 hr at pH 7) but faster under alkaline conditions (half-life of 8 hr at pH 9).
Q o-SnOH
6
Cyhexatin
Azocyclotin
~ ©-SnOC(0)CH
©
3
Fentin Acetate
CH 3
[ (© -
Tributyltin oxide
Figure 57.12 Pesticides that act as inhibitors of complex V-Organotins.
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Azocyclotin (BAY BUE 1452) is an acaricide used on cotton, fruits, and ornamentals under the trade name of Peropal. Toxicology Profile The toxicology of azocyclotin has been reviewed on seve:al occasions by the Joint FAOIWHO Meeting on Pesticide Residues (e.g., see FAOIWHO, 1982, 1990a). These are the primary data sources for the results presented below. Since azocyclotin is readily converted to tricyclohexyltin hydroxide (cyhexatin) there should be considerable similarities in their toxicological properties. Acute Toxicity Azocyclotin's acute toxicity to terrestrial vertebrates generally lies in the range of 100-400 mg/kg (Table 57.4). The acute oral LDso to the guinea pig is 261 mg/kg and falls within this range (Kolbe, 1977). Symptoms in rats, mice, and guinea pigs began about an hour after oral dosing and persisted for up to 14 days. They were characterized by generalized depression of health, drowsiness and lethargy, and loss of weight, with breathing disorders and diarrhea at higher doses. Azocyclotin's toxicity by inhalation is particularly high. IrritationlSensitization corrosive eye irritant.
Azocyclotin is a strong dermal and a
Subchronic Toxicity When rats were dosed orally for 30 days at 20 mg/kg day, a variety of effects were observed and some deaths occurred. Particular changes of note were leucopenia and decreased thymus weight. The latter effect was also seen at 2 mg/kg day. Changes in the weight and function of other organs were also noted including the liver. There was no evidence of the brain edema that is seen with triethyltin. No histopathology was seen in any organ. In dogs, 500 ppm in the diet for 90 days caused no notable effects other than anemia. Inhalation studies over 3 weeks in rats caused weight decreases in the thymus and liver with an LOEL at about 0.001 mg/l, a level which again indicates the high toxicity of azocyclotin by this route of exposure. Chronic Toxicity In chronic dietary studies in rats and mice, no effects other than growth retardation were seen at levels as high 50 ppm. Dogs tolerated even higher doses without serious consequences. Carcinogenicity No evidence of carcinogenicity was seen in the chronic dietary studies. MutagenicitylGenotoxicity Azocyclotin was uniformly negative in a standard battery of in vitro tests for mutagenicity and chromosomal effects. Reproductive Toxicity A three-generation reproductive study was conducted in rates. At the highest dietary concentration, 50 ppm, body weight depression in the pups during lactation and in the parents were the only effects of note.
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CHAPTER 57 Pesticides Affecting Oxidative Phosphorylation
Table 57.4 Acute Toxicity of Pesticides Acting as Inhibitors of Mitochondrial Complex V (ATP Synthase) to Selected Species LDSO (mg/kg)
LCSO (mg/l)
Acute oral Compound
Rat (M; F)
Azocyclotin
209,363
261
Cyhexatin
540
780
Fenbutatin-oxide Fentin acetate
4400 81-298
Fentin hydroxide
l71,110
Diafenthiuron
2068
Tetradifon
>14,700
G.pig
1450c 20
Acute dermal
Acute inhalation
Quail
Duck
Rat or rabbit
Rat
144
250--375 b
>5000
0.02
654 b
>2000
2510
>2000 2000
77
1600
27
>1500
>1500
LCSO (ppb)
(24--96 hr) Fisha
4-100
Daphnia 40
60--550 0.072-D.23 0.044 0.060
4.8-270 320 42-110
31-80 0.32-32 16.5
>2000
0.56
0.7-3.8
<500
>10,000
>3.0
880-2100
>2000
"Range of values from several species, most commonly including the rainbow trout and carp. hData for chickens. CData for mice.
Developmental ToxicitylTeratogenicity In a study with dietary administration to rabbits during gestation, azocyclotin caused gastrointestinal upsets and decreased weight gain in the does at 0.3 mg/kg/day. Minimal fetotoxicity was seen at 1.0 mg/kg/day, the highest dose tested, but no teratogenic effects were observed. At a higher dose (3 mg/kg/day) in a preliminary study, stomach ulceration, abortions, and deaths occurred. Even at a dose of 30 mg/kg/day in a third study, no teratogenesis was observed although this dose was highly toxic and even lethal to the dams. In a dermal study in rabbits at doses up to 300 mg/kg/day, skin irritation was commonly seen together with reduced feed intake, emaciation, and increased resorptions and abortions at the high dose but no teratogenicity was observed. Immunotoxicity The decreases in thymus weights and leucopenia seen in rats raise suspicions of immunotoxicity which is often seen with other organotins, but no specific studies appear to have been conducted with azocyclotin.
aquatic microcosm it was rapidly hydrolyzed to cyhexatin (Kordel and Stein, 1997). In the aquatic microcosm, azocyclotin, when applied pre-adsorbed onto soil, reduced zooplankton populations at concentrations as low as 45 ppt and severe effects on population structures were seen at 135 ppt (Fliedner et aI., 1997). Cyhexatin: General Properties and Uses Tricyclohexyltin hydroxide (Fig. 57.12), CAS Reg. No. 13121-70-5, was discovered by Dow Chemical Co. and is described by Alison et al. (1968). It exists as co10rless crystals, m.p. 195-198°C, v.p. < 10- 8 Pa (20°C), w.S. < 1 ppm (25°C). It is stable to aqueous hydrolysis under slightly acid to alkaline conditions. Cyhexatin (Dowco 213, TD-2383) is used as a contact acaricide on a wide range of crops. Trade names include Acarstin, Aracnol, Metaran, Mitacid, Oxotin, Pennstyl, Plictran, and Triran. Because of concerns about its potential teratogenicity in rabbits after both oral and dermal administration, cyhexatin was withdrawn from use in a number of countries in the late 1980s.
Biochemical Mechanism of Action No studies seem to have been published regarding biochemical actions of azocyclotin, but since it is converted to cyhexatin, observations with this compound are likely to apply to azocyclotin also.
Toxicology Profile The toxicology of cyhexatin has been reviewed several times by the FAOIWHO Joint Meeting on Pesticide Residues (see FAOIWHO, 1990b, 1992a, 1995). These reviews are the major data source for the results presented below. Additional data are available from Cal DFA (1987a, b).
Absorption, Metabolism, and Excretion Azocyclotin was poorly absorbed after an oral dose in rats and the absorbed dose was cleared relatively slowly from the body, primarily in the feces. Only a very low proportion of the dose was voided in the urine. A special study on liver effects in rats showed that a single dose of 25 mg/kg caused a decrease in some, but not all, cytochrome P-450-catalyzed oxidations which lasted several days (FAOIWHO, 1982).
Acute Toxicity Cyhexatin has moderate acute toxicity for vertebrates by the oral route and a low dermal toxicity (Table 57.4). Relatively small differences in acute oral LDso and inconsistent results were obtained in comparing technical and micronized cyhexatin in rats (FAOIWHO, 1995). Clinical signs of poisoning at 50 mg/kg included anorexia, diarrhea, and Central Nervous System (CNS) depression. After intrarumenal doses of 500 and 750 mg/kg in sheep, severe cardiovascular and respiratory pathology was observed at autopsy (John son et aI., 1975).
Environmental Fate and Toxicity Azocyclotin is extremely toxic to the reference aquatic species (Table 57.4). Its persistence in soil varies from a few days to several weeks. In an
IrritationlSensitization
Cyhexatin is an eye and skin irritant.
57.6 Inhibitors of ATP Synthase Subchronic Toxicity The only subchronic studies reported were in rabbits at oral and dermal doses of 3 mg/kg/day for 13 days. Reduced weight gain and skin irritation in the dermal study were the only effects reported. Chronic Toxicity A 12-month dietary study in dogs used doses from 0.25 to 0.75 mg/kg/day. Only marginal effects relating to organ weight were observed at the highest dose. In a 2-year dietary study in rats at 1, 3, or 6 mg/kg/day, decreased weight gain was seen at 3 and 6 mg/kg/day. Bile duct hyperplasia was observed at all does. Other high dose effects were liver hepatocellular alterations, degenerative myopathy, and radiculomyelopathy of the spinal cord. In a study with B6C3F1 mice using the same exposure protocol, no consistent adverse effects were observed. Carcinogenicity No evidence for carcinogemclty was obtained from the long-term dietary studies with cyhexatin in mice and rats. MutagenicitylGenotoxicity Cyhexatin was negative in the Ames assay (Moriya et aI., 1983). Additional studies in the Ames assay, a Chinese hamster ovary cell forward mutation assay, a mouse bone marrow micronucleus test, and an unscheduled DNA repair assay in rat hepatocytes all proved negative (Cal EPA, 1997). In an additional study (Hrelai et aI., 1994) cyhexatin gave weakly positive results in inducing sister chromatid exchange in human peripheral blood lymphocytes in vitro, but it was inactive in inducing chromosomal aberrations in mouse bone marrow cells in vivo. Cyhexatin was also negative in a study of the ability of pesticides to increase the unwinding rate of DNA in rat hepatocytes in vivo (Grilli et aI., 1991). The weight of evidence suugests that cyhexatin is not genotoxic. Reproductive Toxicity A two-generation reproduction study in rats revealed no reproductive toxicity apart from reduced pup weight even though the highest dose (6 mg/kg/day) caused several forms of maternal toxicity including decreased body weight and biliary hyperplasia. A second dietary study in rats utilized 10 (0.7 mg/kg/day), 30, or 100 ppm of micronized cyhexatin and incorporated a teratology component in the study. At 30 and/or 100 ppm a number of effects were seen including reduced food intake, lower body weights, litter size, pup weight and survival, and sternal ossification defects. An increase in cleft plate and thoracic blood vessel malformations was detected at 30 ppm but this was not dose-related since it was not seen at 100 ppm. Overall it was concluded that there was no evidence of the induction of developmental abnormalities. A subsequent study of the effects of dietary restriction on reproduction and development in rats suggested that many, but not all, of these reproductive effects were probably related to reduced food intake because of the unpalatability of the diet. The decreased pup survival seen at 30 ppm was considered to be a compound-related effect and was used as the LOEL. Because of the unpalatability of diets containing cyhexatin (and other organotins) reduced food intake and consequently
1217
reduced weight gain can complicate the interpretation of dietary studies in experimental animals. To examine this effect further, a pair feeding study in rats was run in which food availability to control animals was matched with that consumed ad libitum by the treated groups. This reproductive toxicity study was continued for one generation at the single dietary cyhexatin concentration of 30 ppm. The only adverse effect of note was reduced weight gain in the pups. Since this effect was decreased but still evident with pair feeding, it was concluded that it is, at least in part, a compound-related effect. Developmental ToxicitylTeratogenicity This area of the toxicology of cyhexatin is controversial. Based on positive studies in rabbits, the compound was withdrawn from many markets. However, the multiple studies conducted on its developmental effects and teratogenicity in rabbits present conflicting results. Between 1990 and 1995 the FAOIWHO JMPR received the results of eight teratology studies of cyhexatin in rabbits conducted under reasonably similar conditions (FAOIWHO, 1990b, 1992a, 1995). In terms of teratology, there were two negative oral and two negative dermal studies compared to three positive oral and one positive dermal one. The panel could not reconcile the results from these studies. After considering the weight of evidence and the possible existence of artifacts and uncertainties in some of the positive studies, they concluded that cyhexatin is not teratogenic in rabbits (FAOIWHO, 1995). Teratology studies in rats were uniformly negative. FAOIWHO (1990b) reported the results of four studies that examined the teratology of cyhexatin in rabbits after oral administration during gestation. Two studies found teratogenic effects in the form of hydrocephalus with a LOEL of 0.75 mg/kg/day in one study and 3.0 mg/kg/day in the other. Dilated cerebral ventricles were seen in a few other animals. Embryo and fetal toxicity were observed at 0.75 andl.O mg/kg/day, respectively. These values are all clearly below the doses causing maternal toxicity. In the other two studies, which appear equally valid, no adverse effects were observed at these doses. The possibilities that the hydrocephalus resulted from infection rather than the treatment or was due to poor mixing and nonhomogeneity of the diets were raised by the review panel, but evidence for either of these explanations is far from conclusive. In a teratology study employing dermal application in rabbits with doses up to 3 mg/kg on days 6 through 18 of gestation, an increase in folded retinas was observed but no clear dose-response relationship emerged and fixation artifacts were suspected. Several instances of hydrocephalus were observed at the highest dose. Skin irritation was also noted at the highest dose, but no systemic toxicity or embryotoxcity was observed. In a subsequent study with a similar protocol, the only adverse effect observed was skin irritation. A further study in rabbits (FAOIWHO, I 992a) compared two samples of cyhexatin from different manufacturing sites with different particle sizes to a very high purity sample (99.7%) for their embryotoxicity and teratology. Oral doses of 0.75, 1.5, and 3 mg/kg/day were administered during gestation. In general
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CHAPTER 57
Pesticides Affecting Oxidative Phosphorylation
the severity of effects was greater with the smaller particle size and included lowered food intake and weight gains, pre- and postimplantation losses, fetotoxicity, and reduced litter size. A high incidence of folded lenses was observed at all doses. These were further evaluated later and the possibility of fixation artifacts was again raised since the folding was slight and there was no clear dose-response relationship in their occurrence (FAOIWHO, 1995). Some rib malformations and a small increase in dilation of brain ventricles occurred at 3 mg/kg/day at a rate above that seen in historical controls. There was no evidence of hydrocephaly. Other than particle size effects there were no obvious and consistent differences between the samples. One further study was reported (FAOIWHO, 1995) with dermal doses of 0.5, 1.0, and 3.0 mg/kg/day. No evidence of teratology or maternal toxicity was obtained. In rats, oral doses of 0.5 to 5.0 mg/kg/day resulted in no maternal toxicity or teratogenicity. An increase in the variability of the number of verterbrae at the highest dose was interpreted as evidence of fetotoxicity. An additional study revealed no notable adverse effects at 10 mg/kg/day but in neither of these studies was a dose given that was sufficient to cause maternal toxicity (Cal EPA, 1997). Immunotoxicity No studies specifically addressing the immunotoxicity of cyhexatin were located.
Human Toxicology No cases of human poisoning by cyhexatin were found. Medical surveillance by ELF Atochem (France) of workers in plants manufacturing organotin compounds, including cyhexatin, revealed no adverse effects on hematological, clinical chemical, or immunological parameters. The urinary tin levels of these workers were no higher than those of unexposed control populations. A study of occupational exposure in workers mixing and spraying Plictran SOW in orchard systems in Michigan determined that dermal exposure ranged from 0.7 to 7 mg/day. Very little exposure occurred through inhalation. Protective clothing and respirators were used. Fruit thinners and pickers had dermal exposures of 21 mg/day when working on the day of spraying and 0.83 mg/kg with a 14-day postapplication interval. In a second study conducted in Europe, dermal exposure in mixer/loader/applicators ranged from 0.8 to 19 mg/day. Blood tin level increased by up to 20-fold after exposure and peaked at 20.5 p,g/l in one individual. The levels had not returned to normal 2 days after exposure ended. Biochemical Mechanism of Action Cyhexatin behaves as a typical inhibitor of mitochondrial respiration acting at complex V. It inhibits Mg2+ -dependent ATPase activity in mitochondria from mouse liver and housefly thoraces in vitro with similar 150 values of about 700 nM (Ahmad and Knowles, 1972). Desaiah et al. (1973) found that cyhexatin inhibits oligomycin-sensitive Mg2+ -dependent ATPase (complex V) from spider mites and fish brain with subnanomolar potencies. In a study with beef heart mitochondria, Mehrotra et al.
(1985) concluded that cyhexatin inhibited oligomycin-sensitive Mg2+ -dependent ATPase with an 150 value of 10-20 nM. It also inhibited Ca2+ -dependent ATPase in mitochondria with a similar potency. The authors concluded that it was likely to interfere with both ATP biosynthesis and Ca2+ transport in mitochondria. Other ATPases also are inhibited in vitro such as the Na+, K+-ATPase of rat brain synaptosomes with an IC50 value of 2 J.!M. This inhibition was completely antagonized by preincubation of the synaptosomes with the dithiol dithiothreitol, but not with the monothiols, glutathione or cysteine (Rao et aI., 1987). The toxicological significance of this observation is unclear. Both the basal and isoproterenol-stimulated Ca2+ATPase from the cardiac sarcoplasmic reticulum was inhibited with high potency (lC50 of 25 nM; Sahib and Desaiah, 1986). This inhibition also occurred in rats in vivo and a decrease in the uptake of Ca2+ was seen. Since Ca2+ uptake by the sarcoplasmic reticulum mediates cardiac relaxation, this inhibitory action on Ca2+ flux could lead to cardiac dysfunction (Sahib and Desaiah, 1987). In totality, these results show that cyhexatin potently inhibits several types of ATPases with differing but potentially vital roles in cellular activities. In another study by this group, Kodaventi et al. (1989) report that cyhexatin inhibits calmodulin-activated adenylate cyclase solubilized from subcellular fractions of rat brain with a threshold at 50 nM. All these types of inhibition may well have consequences for animals exposed to cyhexatin, but their absolute and relative importance is unclear, and it remains difficult to connect in vitro biochemical findings with in vivo toxicology for this and other triorganotins. Cyhexatin has several other potentially important biochemical actions besides the inhibition of complex V and other ATPases. These were included in the discussion of the general biochemical mechanisms of action of organotins (Section 57.6.2.5). Cyhexatin inhibited 5-HT uptake by blood platelets from rats in vitro and moderately in platelets taken from rats 30 min after the intraperitoneal administration of cyhexatin at 5 mg/kg (John son and Knowles, 1983), but the toxicological significance of this action is unclear. Absorption, Metabolism, and Elimination A large number of pharmacokinetic studies on cyhexatin have been presented (FAOIWHO, 1992a, 1995). They show that the compound is poorly absorbed in mice, rats, and rabbits after oral dosing, and in rabbits after dermal application. The bioavailability after an oral dose was much higher in rabbits than rats. In rats, blood tin levels peaked at 3--4 hr after a single oral dose of cyhexatin and then declined to levels near those of controls over 24 hr. The micronized product often gave higher blood levels than technical cyhexatin (and also higher toxicity) but the results were not uniform among species and routes of administration. In pregnant rabbits given 3 mg/kg/day over the course of gestation, the half-life of tin in the maternal blood was about 8 hr. The disposition of cyhexatin in rats was slow with an elimination half-life after chronic dosing was terminated of 10 days for most tissues, but 40 days for brain and muscle. Initial residues were cyhexatin but this was subsequently converted to dicyclohexyltin. The
57.6 Inhibitors of ATP Synthase
major route of elimination of both oral and dermal doses in rats and rabbits was the urine, but the fecal route also contributed significantly to clearance. The typical pathway of metabolism in mammals involves sequential conversion to dicyclohexyltin oxide, cyclohexylstannoic acid, and inorganic tin. The initial attack on cyhexatin involves microsomal oxidase attack to form the 2-, 3-, and 4-hydroxycyclohexyl analogs. The 2-hydroxy metabolite is readily converted to dicyclohexyltin (Kimmel et aI., 1980). Environmental Fate and Toxicity Cyhexatin's acute toxicity (Table 57.4) and dietary toxicity to birds is moderate to low with 8-day dietary LCso values of 3189 ppm for mallard ducks and 520 ppm for bobwhite quail. No reproductive impairment was observed in bobwhite quail fed 5 or 20 ppm prior to egg laying and through the egg laying cycle (Fink, 1975). Cyhexatin is generally much more toxic to aquatic species than terrestrial ones (Table 57.4). Devries et al. (1991) conducted a comparative study of the effects of several organotins on the development of rainbow trout fry. Trout were continually exposed for 110 days to these compounds, beginning at the yolk sac stage. Of the triorganotins, cyhexatin chloride was more toxic than either tributyl or triphenyltin chlorides. It caused 100% mortality of fry within 1 week at a concentration of 3 nM and only a few trout survived exposure at 0.6 nM (0.24 ppb) for the full 11 0 days. Thymus atrophy, seen in mammals with many triorganotins, was not observed, but evidence of immunotoxic effects was obtained even at the lowest concentration tested, as indicated by decreased resistance to a bacterial challenge with Aeromonas hydrophila. In the environment, cyhexatin binds firmly to soils and is converted to di- and monocyclohexyl tins and inorganic tin with half life of about 10 days (Muller and Bosshardt, 1987). This, and its low water solubility, indicate that it has a low potential to leach from soils. It's degradation is enhanced faster by exposure to VV light on soils or in water. In aquatic systems it is mainly adsorbed on sediments. The half-life for degradation in an aquatic microcosm study was 68 days (Kordel and Stein, 1997). Fenbutatin-Oxide: General Properties and Vses Bis[tris(2methyl-2-phenylpropyl)tin] oxide (Fig. 57.12), CAS Reg. No. 13356-08-6, was discovered by Shell Chemical Co. lt exists as colorless crystals, m.p. 145°C, v.p. 8.5 x 10- 8 Pa (20°C), w.S. 12.7 ppb (20°C), log P 5.2. Fenbutatin-oxide hydrolyzes very slowly and reversibly in aqueous solution to the corresponding monotin hydroxide. At pH 5-9, less than 10% conversion occurred in 30 days at 25°C. Fenbutatin-oxide is also very stable to heat, light, and oxidation. Fenbutatin-oxide (SD 14114) is a long-lasting nonsystemic acaricide. lt also controls some sucking insects. Common trade names include Lexitin, Novran, Osadan, Torque, and Vendex. Toxicology Overview The toxicology of fenbutatin-oxide has been reviewed by the Joint FAOIWHO Meeting on Pesticide Residues (FAOIWHO, 1993) and by the V.S. EPA (1994).
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These reports are the primary data sources for the overview below. Acute Toxicity Fenbutatin-oxide has a low acute toxicity to terrestrial vertebrates (Table 57.4). Toxicity by inhalation is considerably higher than by the oral and dermal routes. Inhalation exposure produced necrosis of bronchiolar and renal tubule epithelia, lung congestion, and edema. An intraperitoneal LDso of 33 mg/kg in the rat (FAOIWHO, 1993) suggests that fenbutatin-oxide has a high intrinsic toxicity if it is absorbed systemically. IrritationlSensitization As with other organotins it is somewhat irritating to skin and a severe eye irritant but it is not a dermal sensitizer. Subchronic Toxicity No specific target organ was evident in subchronic and chronic dietary studies. A 3-week dermal application study in rabbits at doses up to 5 mg/kg/day caused no systemic activity but severe local skin irritation and edema occurred. An initial range-finding study in rats showed reductions in food intake, body weight, and some organ weights and changes at dietary concentrations of 300 ppm and above. However, a pair-feeding study established that this was due to unpalatability of the diet rather than systemic toxicity. Chronic toxicity Emesis and diarrhea were common but otherwise no major adverse effects were seen in a 2-year dietary study in dogs in which doses up to 60 mg/kg/day were administered by capsule. No adverse effects except reduced weight gains due to diet unpalatability were recorded in 2-year dietary studies in rats and mice with fenbutatin-oxide at levels up to 600 ppm. In a lifetime feeding study, female rats showed a decrease in the number of leucocytes at higher exposure levels with an LOEL at 15 mg/kg/day. Carcinogenicity Fenbutatin-oxide is classified as a group E carcinogen (evidence of noncarcinogenicity in humans) by the V.S. Environmental Protection Agency since none of the chronic studies above revealed evidence of carcinogenicity. MutagenicitylGenotoxicity No evidence of genotoxicity was obtained in an extensive battery of in vitro and in vivo tests. Additional studies by Moriya et al. (1983) also gave negative results. Reproductive Toxicity In a multi generation study in rats, a LOEL was set at 17 to 20 mg/kg/day based on reduced pup body weight. Parental effects in the form of reduced food intake and weight occurred at the same dose. A special study on the effects of a single high oral dose of febutatin-oxide on the reproductive tract of male rabbits proved negative, even at toxic doses that casued lethality, emaciation, and lesions of the gastric mucosa. Developmental ToxicitylTeratogenicity In standard studies with oral dosing in pregnant females, an increase in abortions
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CHAPTER 57
Pesticides Affecting Oxidative Phosphorylation
and resorptions was observed in rabbits at an oral dose of 5 mg/kg/day. This was accompanied by anorexia and gastric lesions in the does. A similar study in rats demonstrated maternal toxicity at 30 mg/kg/day, but no reproductive endpoints or treatment-related malformations were seen.
Neurotoxicity Because of the brain edema known to be caused by triethyltin, a study was performed in rats with fenbutatin-oxide at a dose of 1000 mg/kg. No evidence of brain pathology was seen although triethyltin bromide, as a positive control, caused the anticipated lesions. No effect was seen on EEG patterns in dogs dosed orally with 30 mg/kg/day for 14 days. Immunotoxicity Results of a 7-day feeding study in mice tentatively indicated that fenbutatin-oxide has a lower potential to cause immunotoxicity than other organotins. Biochemical Mechanism of Action Few investigations have been conducted with fenbutatin-oxide at the biochemical level. In one study, Machera et al. (1996) exposed nauplii of the brine shrimp Artemia to fenbutatin-oxide for 24 hr and found that Mg2+ -ATPase activity in whole tissues was inhibited by 66% at a water concentration of 25 ppb. The Na+ /K+ -ATPase activity was also inhibited but to a lower degree. Studies on platelets conducted by lohnson and Knowles (1983) and Knowles and lohnson (1996) showed that fenbutatin-oxide inhibited 5-hydroxytryptamine uptake by, and enhanced its release from, rat platelets at 10 J.lM in vitro and after intraperitoneal treatment at 2.5 mg/kg. This was attributed to its effects on ATP production. No effect was seen on the aggregation of rat platelets in vitro at 10 J.lM. Absorption, Metabolism, and Elimination Fenbutatinoxide was very poorly absorbed after oral dosing. In rats, 99% of single or repeated oral doses was excreted in the feces, almost entirely as unchanged fenbutatin-oxide. Metabolites identified both in animals and soil include {3, {3-dimethylphenethylstannoic acid and the bis-dealkylation product, 1,3-dihydroxy1,1 ,3,3-tetrakis(2-methyl-2-phenylpropyl)distannoxane. This metabolite showed no appreciable toxicity to rats in a subchronic study at levels of dietary exposure up to 300 ppm and it did not cause brain edema in rats after oral dosing at 100 mg/kg (FAOIWHO, 1993). Environmental Fate and Toxicity The acute toxicity to birds appears to be very low (Table 57.4). Similarly, the dietary 8-day LCso for bobwhite quail is >5620 ppm in the diet. Avian reproduction studies in mallard ducks and bob white quail showed no adverse effect at dietary doses up to 150 ppm. By contrast, fenbutatin-oxide is very highly toxic to both vertebrate and invertebrate aquatic species (Table 57.4). An LCso value as low as 1.7 ppb was obtained with rainbow trout and most LCso values for fish and aquatic invertebrates are below 50 ppb. A developmental study in rainbow trout showed that growth and survival of juveniles was impaired at concentrations above 0.31 ppb, and
a life-cycle study in Daphnia showed that survival is impaired at concentrations greater than 16 ppb. Additional tests on estuarine and marine organisms also established the very high aquatic toxicity of fenbutatin-oxide with LCso values of 2.8 ppb for mysid shrimps and 0.4 ppb for eastern oyster larvae. Machera et al. (1996) obtained an LCso of 50 ppb for brine shrimp nauplii using a 55% formulation of fenbutatin-oxide. Fenbutatin-oxide has a rather high BCF of 490-730 in bluegill sunfish, which in view of its high log P value and resistance to metabolism is not surprising. The true BCF factor is probably higher since the concentration of fenbutatin-oxide did not reach an equilibrium during the course of this study (U.S. EPA, 1994). Depuration was slow with only 50-75% of the tissue residues being cleared in 14 days. Despite its high aquatic toxicity in laboratory tests, the impact of fenbutatin-oxide in actual use may be lower than predicted since it binds very strongly to soil and sediments which limits its availability in the water phase of natural aquatic systems. Binding to soil also results in minimal mobility in the soil and a low potential to leach. Kd values range from 1282 to 2333, depending on soil type. Both hydrolytic and photolytic degradation are very slow with half-lives over 100 days in typicallaboratory studies. Microbial metabolism in the soil under both aerobic and anaerobic conditions is also extremely slow with half-lives of months to years. As a result, fenbutatin-oxide is persistent in soils with a half-life estimated to be slightly less than 1 year based on field measurements (Grey et aI., 1995). Other estimates of its half-life under varying field and environmental conditions range from 271 to 1367 days (U.S. EPA, 1994). This stability gives fenbutatin-oxide the capability to accumulate under conditions of repeated use.
Fentin: General Properties and Uses Triphenyltin is used as a pesticide in any of three forms, as the hydroxide (CAS Reg. No. 76-87-9), as the acetate (CAS Reg. No. 900-95-8), or, to a lesser extent, as the chloride (CAS Reg. No. 639-58-7) (Fig. 57.12). It was discovered by Hoechst (acetate) and N.Y. Philips-Duphar (hydroxide) and is described by van der Kerk and Luijten (1954). A comprehensive monograph on the chemistry, biological properties, toxicology, and environmental fate of triphenyltins is provided by Bock (1981). Fentin hydroxide was first registered in the United States in 1971. In 1985 the U.S. Environmental Protection Agency instituted a Special Review based on its potential developmental toxicity to mixers, loaders, and applicators. This and other toxicological issues were resolved in a recent Reregistration Eligibility Decision (U.S. EPA, 1999c) which included a number of steps to reduce exposure. The hydroxide exists as colorless crystals with m.p. 118120°C, v.p. 6.5 X 10- 6 Pa (20°C), w.s. about 1 ppm (pH 7, 20°C), log P3.43, pKa 5.20. The acetate exists as colorless crystals with m.p. 121-123°C, w.s. about 19 ppm (pH 5, 20°C), v.p. 1.9 x 10-3 Pa (60°), log P3.1-3.4. The chloride exists as colorless crystals, m. p. 106-107°C. The acetate and chloride are quite readily hydrolyzed to the hydroxide (e.g., the half-life of the acetate in water at pH 7 is <3 hr). On gentle heating, the
57.6 Inhibitors of ATP Synthase hydroxide is dehydrated to bis(triphenyltin) oxide. Ultraviolet irradiation gradually dearylates triphenyltins to inorganic tin. The principal use of all forms of fentin is as nonsystemic foliar fungicides with additional activity as antifeeding agents for some insects. Fentin acetate and chloride also control algae and snails in rice and fish ponds and fen tin compounds are used as antifouling agents on marine vessels. Common trade names include Brestan H, Du-Ter, Flo-Tin, Haitin, Photon, Pro-Tex, SuperTin, Suzu H, and Tubotin (hydroxide), Aquatin and Tinmate (chloride), and Brestan, Phytex, Radar, Suzu, Triacetane, and Trimastan (acetate). Toxicology Profile Fentin hydroxide has been very extensively tested for toxicity in experimental animals and a variety of adverse effects have been displayed, although an unexplained lack of reproducibility between some studies is apparent. A wealth of older toxicology results dating back to the 19th century are provided by Bock (1981) but generally are not included here. The primary data sources are FAOIWHO (1992b) and U.S. EPA (1999c) which report a large number of unpublished studies conducted in the 1980s. Acute Toxicity Fentin hydroxide is generally of moderate to high acute toxicity to terrestrial vertebrates by the oral and dermal routes, but it has a much higher toxicity by inhalation (Table 57.4). Results from different studies tend to be rather variable. A value for the dermal LDso as low as 127 mg/kg has been reported for the rabbit although the value for rat is 1600 mg/kg (FAOIWHO, 1992b). Fentin acetate has a roughly similar acute toxicity which is reasonable since it is rapidly converted to the hydroxide in vivo. Sensitivity seems to vary considerably among different mammalian species [e.g., the oral LDso values for mice vary from 81 to 245 mg/kg, but for male guinea pigs the LDso is only 25 mg/kg (FAOIWHO, 1992b)]. For the rabbit a value of 80 mg/kg has been recorded. Reports also vary for the acute dermal toxicity with a value for the rat > 2000 mg/kg, but only 350 mg/kg for the mouse (FAOIWHO, 1992b). Like fen tin hydroxide, the acetate is very toxic by inhalation (Table 57.4). Its intrinsic toxicity is very high as revealed by an intraperitoneal LDso value of 3.6 mg/kg for male rats. Fentin chloride is extremely toxic with an acute oral LDso to the mouse of 18 mg/kg. Signs of intoxication are typical of triorganotins and include anorexia, emesis, tremors, diarrhea, drowsiness, and ataxia. IrritationlSensitization As with other organotins, fentin hydroxide is a severe eye irritant. Skin irritation is mild to moderate and it is not a dermal sensitizer. Fentin acetate has similar properties but did give a positive Buehler test for skin sensitization in guinea pigs (FAOIWHO, 1992b). Subchronic Toxicity In subchronic feeding studies in rats, multiple hematologic and blood chemistry changes were observed at higher doses of 7 to 8 mg/kg/day, The most sensitive endpoint was a decrease in IgG antibodies observed at the
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lowest dose tested, 0.75 mg/kg/day. These levels were still depressed after a 4-week recovery period. A similar result was obtained in a mouse subchronic feeding study with decreased levels of immunoglobulin antibodies seen at the lowest dose, 0.75 mg/kg/day, more extensive antibody effects and decreased adrenal weights at 3.8 mg/kg/day, and decreased ovary weights and hemoglobin at 19.5 mg/kg/day. In a parallel study in guinea pigs, decreased leucocyte counts were observed at the lowest dose tested, 0.1 mg/kg/day. In a subchronic dermal studies in Wistar rats only skin irritation without systemic toxicity was seen at 10 mg/kg/day, but several effects on biochemical parameters in the blood and clinical signs (piloerection, mydriasis, dyspnea) and some deaths were seen at 20 mg/kg/day. A second dermal study using Charles River rats failed to reveal any systemic toxicity at 20 mg/kg/day. A 13-week subchronic inhalation study in rats had an LOEL at 0.002 mg/l. Adverse effects at this very low exposure level included reductions in white blood cells, lung and respiratory irritation and edema, and death, particularly in males. Additional 70-day subchronic dietary studies have been conducted with rabbits and lambs at concentrations from 15 to 150 ppm (Dacasto et aI., 1994). In both species, decreased body weight and decreased relative thymus weight were recorded. In the rabbit, the major lesions were found in the thymus and lymph nodes. Lambs showed similar but less severe lesions.This immunosuppressive activity echoes that seen in rodents, but the lamb and rabbit are much less sensitive. Chronic Toxicity A I-year feeding study in dogs showed no major effects at the highest dose, about 0.6 mg/kg/day. However, a lifetime study in mice had an endpoint based on a decreased leucocyte count at 0.25 mg/kg/day. In a chronic dietary study in rats, decreases in immunoglobulins and increased deaths were seen at doses as low as 0.3-0.4 mg/kg/day and decreased immunoglobulins were also found in a parallel mouse study with doses as low as 0.85 mg/kg/day. These were the lowest doses tested and no NOEL was established. MutagenicitylGenotoxicity Fentins (both as the acetate and hydroxide) are generally negative in standard genotoxicity tests, but fentin hydroxide gave a positive result for chromosomal aberrations after metabolic activation in human lymphocytes (Moriya et aI., 1983). A similar positive result with activation is reported by FAOIWHO (1992b). Positive result were also obtained in two mouse lymphoma mutation assays. These positives were tentatively interpreted by FAOIWHO (1992b) in terms of the toxicity of fentin to the lymphocytes rather than as a specific clastogenic action. In any case, the results of clastogenicity studies in vivo have been negative. Carcinogenicity Fentin hydroxide caused pituitary and testicular (Leydig cell) tumors in rats 1-2 mg/kg/day and hepatocellular adenomas and carcinomas in mice at about 15 mg/kg/day in standard long-term dietary studies. Two previous chronic dietary studies at comparable dose rates in mice, and one in rats, had proved negative for carcinogenicity (FAOIWHO, 1992b).
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CHAPTER 57 Pesticides Affecting Oxidative Phosphorylation
Fentin hydroxide was classified as a B2 carcinogen (probable human carcinogen) by the U.S. Environmental Protection Agency. A Qt value of 18.3 (mg/kg/day)-l was calculated. Reproductive Toxicity Fentin hydroxide causes reproductive toxicity. Decreases in litter size, pup weight, and the relative sizes of the spleen and thymus of weanlings were observed in rats in a multigenerational feeding study at a dose (0.9 mg/kg/day) that did not cause maternal toxicity. A study of the effects of fen tin hydroxide on male fertility in rats when given orally at 20 mg/kg/day for 25 days showed a reduction in spermatazoa count that was reverscd aftcr a 70-day recovery period. Ema et al. (1997) observed that an oral dose of triphenyltin chloride of either 4.7 or 6.3 mg/kg given to female rats on days o to 3 of pregnancy induces implantation failures by enhancing preimplantation embryonic loss. No subsequent adverse effects were seen at these doses on the fetuses where implantation was successful. By using a pair-feeding approach, they concluded that this was not due to the decreased food consumption commonly seen in studies with organotins (Harazano et aI., 1998). In further exploring the mechanism for this effect, Ema et al. (1999) found that oral doses that caused early implantation losses also suppressed uterine decidualization as indicated by greatly decreased uterine weights. At these doses, fentin chloride also significantly reduced serum levels of progesterone, which is required for decidualization. Ema et al. therefore conclude that fentin in some way interferes with progesterone biosyntheses which decreases the receptivity of the uterus and leads to implantation failures. Developmental Toxicity/Mutagenicity In dietary studies with pregnant rabbits, fen tin hydroxide caused both fetotoxicity and teratogenicity. These were detected in the form of decreased implantations and increased resorptions, decreased fetal weight, and various malformations including poorly ossified skeletal elements at an LOEL of 1 mg/kg/day. Fetotoxicity, but only relatively minor effects on ossification, were seen at 8 mg/kg/day in several studies in rats. Hamsters at 12 mg/kg/day showed decreased pup weight and various malformations including skeletal and other minor anomalies. In every case, these doses also caused maternal toxicity in the form of reduced feeding and weight gain and sometimes more severe toxic effects. A dermal study in rats gave negative results at the highest dose tested, 3 mg/kg/day. Reports of occasional hydronephrosis, hydrocephalus, omphalocele, and hydroureter in some previous studies led to the establishment of a special study of the induction of these effects in rats. However, no evidence for such irreversible structural effects was seen at 8 mg/kg/day which was clearly toxic to the dams. Immunotoxicity There is strong evidence for immunotoxicity. Lymphocytes and immunoglobulins were decreased after subchronic or chronic dietary administration of fentin in guinea pigs, mice, and rats at low doses. In specific tests for immunotoxicity, decreases in spleen weights, leucocytes, lymphocytes,
and antibody immunogloblins, and lymphoid depletion in the thymus and spleen were observed in mice. Effects on the spleen were seen with doses as low as 1.2 mg/kg/day in mice. Similar, but less extensive, effects were observed in rats beginning at a dose of 3.4 mg/kg/day. Ishaaya et al. (1976) showed that fentin acetate fed to mice for 4 days at dietary levels of 30 to 300 ppm caused a dose-related decrease in spleen weight and in the number of blood leucocytes. A review of the literature on this topic (FAOIWHO, 1992b) indicates that short-term (up to 13 weeks) exposure to fentins caused immunosuppression marked by lymphopenia and lymphocyte depletion in the spleen and thymus resulting in altered cellular and humoral immunity, but these effects are transient and tend to diminish on longer exposure. Some tests of immune function show no response to organotins. Overall it appears that fen tins display weak immunosuppressive actions and that these appear at doses that reduce levels of circulating lymphocytes and the weight of the lymphoid organs. Endocrine Effects Fentin hydroxide appears to be an endocrine disrupter. It causes testicular and pituitary tumors in rats and changes in both adrenal and ovary weights in mice. Embryo implantation losses caused by fentins may involve decreased progesterone levels (Ema et aI., 1999). Fentins also induce masculinization of female mollusks (imposex) as describe later. Further mechanistic studies are needed to determine its status in this respect.
Human Toxicology According to U.S. EPA (1999c) there have been very few, if any, well documented cases of human poisoning by triphenyltin hydroxide in the United States. However, FAOIWHO (1992b) provides several examples of fentin poisoning. Two cases involved occupational inhalation of Brestan 60 (60% fentin acetate with 15% maneb) during preparation of spray solutions by farmers. Signs of poisoning were nausea, dizziness, transient loss of consciousness, convulsions, persistent headache, photophobia, and impaired liver functions. Recovery occurred after 10-15 days (Manzo and Richelmi, 1981). A further example of human poisoning by triphenyltin acetate with neurological signs involved a 36-yearold male who developed dizziness and generalized paroxysmal abnormalities with slowed EEG rhythms after dermal exposure. Skin injury was also reported (Colosio et aI., 1991). Wu et al. (1990) described a case based on a suicide attempt by a 23-year-old male who consumed a fentin acetatecontaining molluscicidal preparation (amount unstated). Abdominal pains, diarrhea, vomiting, severe ataxia, and coma resulted. The patient also showed neurological effects including a reversible sensorimotor polyneuropathy that developed 2 months after exposure. A full recovery was attained in 3.5 months. This diagnosis has been questioned by Cavanaugh (1995) on the basis of the long delay in the appearance of some neurological symptoms and the lack of previous evidence for triphenyltin-induced neurotoxicity in humans. In the same issue, Wu et at. defended their diagnosis. Subsequently, a second
57.6 Inhibitors of ATP Synthase case of human poisoning by fen tin acetate with substantial neurological involvement has been reported by Lin et al. (1998). In this case, the patient, an 18-year-old female who intentionally consumed about 33 g of a 45% triphenyltin acetate formulation, felt weakness and nausea 3 days later despite previous gastric lavage and treatment with activated chrcoal. After hospitalization, she lost full consciousness for 9 days, showing sponteous involuntary movement of the hands, facial twitching, and emotional instability. Diplopia, drowsines, giddiness and vertigo, bidirectional nystagmus, confusion, and disorientation also developed. An EEG showed mild cortical dysfunction wihout seizures. Scans by MRI and Tc-99m HMPAO brain SPECT to assess regional blood flow were unremarkable. Thus fentin may cause cellular dysfunction in the brain without evident structural damage. Leucopenia was noted on the sixth day and liver impairment occurred on the ninth day. Borderline delayed peripheral neuropathy developed on day 53. Recovery of normal neurological functions took 1 year. Further examples of fentin poisoning with neurological involvement, in addition to those cited by FAOIWHO above, have been reported by Lin and Hsueh (1993) involving exposure by ingestion; included irritability, blurred vision, headache, and disturbances of consciousness as well as reversible acute nephropathy (proximal tubule necrosis) and hepatitis. Biochemical Mechanism of Action Fentin compounds inhibit mitochondrial ATPase activity with considerable potency. Pieper and Casida (1965) found an Iso value of 100 nM for the inhibition of ATPase by triphenyltin chloride in a housefly thoracic particulate preparation. With triphenyltin hydroxide, Ahmad and Knowles (1972) calculated an Iso values slightly below 1 J.lM for Mg2+ -stimulated ATPase activity in both mouse liver and housefly thoracic mitochondria. Several other mechanistic studies with fen tin were outlined in Section 57.6.2.5. The effects of organotins on cytochrome P450 and its toxicological implications were described. Enough work has been conducted with fen tins in this regard to indicate that they are relatively potent in this action. Further studies that shed additional light on the mechanisms of toxicity of fentin are presenterd below. Zazueta et al. (1994) studied the effect of fentin on membrane permeability and calcium regulation in mitochondria from rat kidney. At low micromolar concentrations it induced a rapid increase in permeability with a fall in membrane potential and a loss of calcium and matrix proteins as it induced the mitochondrial pemeability transition. This was not due to a general detergent effect since it was efficiently reversed by sulfhydryl reagents. This opening of the permeability transition pore was not reversed by cyclosporin A, but EDTA was effective. Several potential mechanisms for this effect are considered including a role for fentin in enhancing calcium cycling across the membrane, but the authors suggest that the most likely is the binding of fentin to sulfhydryl groups that trigger the transition pore to open. Tributyltin causes similar effects (Stridh et aI., 1999b). In a recent study of the effect of phenyltins on human NK (natural killer) cells in vitro, Whalen et al. (2000) found that
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I J.lM triphenyltin almost eliminated their tumor killing capability without reducing their general viability. Over 24 hr fen tin decreased the ability of NK cells to bind to tumor cells, but this is a secondary effect since a dramatic decrease in NK cell killing efficiency was seen within 1 hr of exposure to tin compounds when cell binding was still unaffected. The order of potency among phenyltins was triphenyl>diphenyl>monophenyl, but triphenyltin was slightly less active than tributyltin. Since NK cells are a primary defense against tumor and virus-infected cells, a decrease in their efficiency in vivo could lead to enhanced sensitivity to tumor formation and viral infections. Absorption, Metabolism and Elimination Clearance of an oral dose of triphenyltin hydroxide in rats was primarily by the biliary route. In different studies 80-100% of the dose was recovered in the feces with much being in the form of the parent compound, indicating limited uptake, averaging 40% of the dose. Successive oxidative removal of the phenyl groups eventually produces inorganic tin. Because of slow clearance, on repeated dosing cumulation of residues occurs in tissues giving levels up to sevenfold above those found with a single dose. The significance of cytochrome P-450 oxidations in the degradation of fentin is underlined by the observation of Ohhira et al. (2000) that when mice are administered SKF-525A, a P-450 inhibitor, together with fentin, the levels of fen tin in the tissues are increased about threefold after 24 hr and the amount of metabolites is reduced. Because of the elevated tissue levels of fentin, hyperglycemia was observed in the SKF-525A-treated animals but not in the controls. In hamsters, oxidative dearylation of fentin is slower, the levels in the tissues, including the pancreas, reach a higher level than in rats, and hamsters show a hyperglycemic response to trifentin without manipUlation of P-450 activities (Ohhira et aI., 2000). Fentin hydroxide binds strongly to the skin making dermal absorption studies difficult. Little of the dermal dose is absorbed in 10 hr, but slow uptake may continue for some time from reservoirs held in the skin (U.S. EPA, 1999c). Environmental Fate and Toxicity Fentin hydroxide has a moderate to high acute toxicity to birds (Table 57.4). It is very highly toxic to aquatic species in general. In a study of the effects of organotins on the development of rainbow trout fry by Devries et at. (1991) fentin chloride caused acute mortality at 15 nM. The no-effect concentration on juvenile development was very low at 0.12 nM (0.05 ppb). Diphenyltin was about lOO-fold less active. Thymus atrophy, seen with fentin in mammals, was not observed in the fish, but evidence of immunotoxic effects was obtained even at the lowest effect level as indicated by decreased resistance to a bacterial challenge. As noted previously, fentin has a very strong capability to cause the imposex phenomenon in mollusks by its action as a xenoandrogen.lts potency in this regard in the rock shell (Thais clavigera) was about the same as that of tributyltin (Horiguchi et aI., 1997). As usual, di- and monophenyltins were much less active. Fentin induced imposex in the female ramshom snail (Marisa comuarietis) with an EClO after exposure for 4 months
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Pesticides Affecting Oxidative Phosphorylation
of 12.3 ppt. Fecundity was reduced at an even lower concentration (5.4 ppb) with a complete lack of spawning at 163 ppt (Schulte-Oehlmann et aI., 2000). Fentin hydroxide binds very strongly to soils (Koc values range from 1900 to 54,000 ml/g; U.S. EPA, 1999c) and to humic substances in water (Looser et aI., 1998). It is resistant to both hydrolysis and photolysis. Estimates of the half-life for its degradation in soil range from 21 to about 140 days. A BCF of 3700 for whole body accumulation in fish indicates a significant capability for bioaccumulation. Larvae of the fish (Thymallus thymallus) showed a similar BCF value of 2200 (Looser et aI., 1998). It is also bioconcentrated by aquatic invertebrtaes with BCF factors ranging from 190 for Daphnia to 680 for Chironomus. Uptake was higher at pH 8 than pH 5, as expected from the dissociation of the hydroxide at acidic pH, but uptake seemed to occur even with the triphenyltin cation.
57.6.3 OTHER INHIBITORS OF COMPLEX V 57.6.3.1 Diafenthiuron 1- Tert - buty1- 3 - (2 ,6- diisopropy1-4 -phenoxypheny1)thiourea (Fig. 57.13), CAS Reg. No. 80060-09-9, was discovered by Ciba-Geigy Corp. and is described by Streibert et al. (1988). Its development and properties are described by Drabek et al. (1990).
General Properties It exists as a white powder, m.p. 150°C, v.p. 2.2 x 10-7 Pa (20°C), solubility in water 0.06 ppm (25°C), log P5.76, HLC 1.28 x 10-2 Pam3 mol-I. It is converted by ultraviolet light to the corresponding carbodiimide CGA 140400 (Drabek et aI., 1990). Uses Diafenthiuron (CGA 106,603) is used as an insecticide, particularly against whiteflies and aphids, and as an acaricide. Trade names include Pegasus and Polo.
Diafenthiuron
CGA 140408 Figure 57.13 Pesticides that act as inhibitors of complex V-Diafenthiuron and its active metabolite.
Toxicology Profile The sources of data are Streibert et al. (1988) and Anonymous (1992b). Only limited data are available. Acute Toxicity The acute toxicity of diafenthiuron to terrestrial vertebrates is low (Table 57.4). The target tissue in rats fed large doses of diafenthiuron is the lung where it increased lung weight associated with an increased incidence of alveolar foam cells. At high lethal doses given intraperitoneally, mice die within 4-24 hr with symptoms of lethargy, respiratory distress, and intermittent tonic contractions of the limb extensors, but at lower doses only lethargy, anorexia, and dehydration are seen and death may be delayed for many days (Petroske and Casida, 1995). IrritationlSensitization Diafenthiuron is only slightly irritating to the skin or eye and it is not a dermal sensitizer. Chronic Toxicity Reversible edema of the pancreas was observed in dogs in a I-year study. Chronic feeding studies in mice caused proliferative lesions (focal hyperplasia, adenoma, and carcinomas) which were considered to be secondary to the cytotoxicity caused by the high dose and not predictive of carcinogenicity in humans. Carcinogenicity No additional results on the carcinogenicity of diafenthiuron were found. MutagenicitylGenotoxicity Diafenthiuron has not been found to be mutagenic in a suite of standard tests. Reproductive Toxicity reproductive toxic ant.
Diafenthiuron was not found to be a
Developmental ToxicitylTeratogenicity negative in studies of teratogenicity.
Diafenthiuron was
Biochemical Mechanism of Action Diafenthiuron is a propesticide. The active metabolite is the corresponding carbodiimide (CGA 140,408) which is produced on the leaf surface by sunlight and by oxidative metabolism in living organisms (Petroske and Casida, 1995; Ruder and Kayser, 1992; Ruder et aI., 1991). This conversion is shown in Fig. 57.13. CGA 140,408 is stabilized against nucleophilic attack through steric shielding of the highly reactive carbodiimide center by the surrounding bulky and hydrophobic alkyl groups. It does react with water slowly, producing the urea analog, and with carboxylic acid groups in fatty acids and some amino acids. This reactivity is central to its mechanism of action since the carbodiimide reacts covalently with a critical aspartate residue in the Fo portion of mitochondrial ATP synthase, thus blocking ATP biosynthesis (Ruder and Kayser, 1993). This parallels the action of the well-known ATP synthesis inhibitor, dicyclohexy1carbodiimide (DCCD). This compound has been shown to react with a specific aspartate residue in the "e" subunit of the Fo component (Fig. 57.5). Only 1 "e" subunit of the 9-12 present in each Fo
57.7 Mitochondrial Uncouplers unit needs to be derivatized in order to fully block ATP synthase activity. It is postulated that this carboxylic acid group is intimately concerned with proton transport through the Fo system by protonation on the cytoplasmic face and subsquent deprotonation toward the matrix side (Fillingame, 1992). Additional evidence that the carbodiimide is the active agent and the site of action in insects is ATP synthase was derived by Ruder and his colleagues (Ruder et aI., 1991; Ruder and Kayser, 1992, 1993) and is reviewed by Hollingworth and Gadelhak (1998). This includes the observation that although both compounds are ATPase inhibitors in vivo, only the carbodiimide is an active inhibitor in vitro, that the carbodiimide acts more rapidly and potently in insects than its parent, and that in insects the toxicity of diafenthiuron is antagonized by piperonyl butoxide, a microsomal monooxygenase inhibitor. After in vivo exposure, diafenthiuron caused a significant decrease in ATP levels in the nervous system of locusts, the apparent target tissue in insect, and the severity of poisoning correlated well with the degree of inhibition ATP synthase in this tissue. Vetebrate mitochondrial ATP synthase is also sensitive to inhibition by the carbodiimide both in vitro and in vivo, but piperonyl butoxide does not antagonize the acute toxicity of diafenthiuron in mice (Petroske and Casida, 1995). Even so, these authors concluded that the levels of the carbodiimide produced in mice in vivo were sufficient to cause severe inhibition of ATPase activity, as judged by the sensitivity of the enzyme in vitro. They also observed that whereas the intraperitoneal LDso in mice for diafenthiuron was 15 mg/kg, that of the carbodiimide was only 0.3 mg/kg, although the same symptomatology was observed. This again supports the activation hypothesis. In insects, but not in rats or plants, the carbodiimide metabolite also reacts with porin (the voltage dependent anion channel) of the outer mitochondrial membrane (Ruder and Kayser, 1993; Wiesner et aI., 1996). The toxicological significance of this reaction is unclear, in part because the functions of porin are unclear. The open porin channel is large, nonselective, and appears to allow a range of solutes to enter the mitochondrion which may then be taken up through the inner membrane. It also is an attachment site for several cytoplasmic enzymes such as hexokinase (Crompton et aI., 1999). Finally, as already described, it is probably a critical component of the mitochondrial permeability transition pore that is central to the role of mitochondria in apotosis. The reaction with carbodiimides does not change the conductance of this channel, but it does alter its voltage dependence, leading to a shift moring the open the closed state (Wiesner et aI., 1996). It also seems reasonable to suppose that the extensive production of a carbodiimide in vivo, even one of reduced reactivity, would lead to the alkylation of other cellular macromolecules, as is true for DCCD (e.g., see Solioz, 1984). If so, it has not been reported except for the finding of derivatized fatty acids as metabolites in vivo. However, Petroske and Casida (1995) concluded that although the rapid acute toxicity to mice at high doses of diafenthiuron could be explained by the inhibition of ATP synthase, the delayed toxicity seen at lower doses was not related to such inhibition since this enzyme had already
1225
recovered well before death occurred. Other sites of action may therefore be significant in the causation of this type of toxicity. It has been suggested that diafenthiuron and its carbodiimide may act in insects by stimulating biogenic amine (octopamine) receptors (Kadir and Knowles, 1991). This capability often correlates with activity on GY2-adrenergic receptors in vertebrates. However, the octopaminergic action of these compounds could not be confirmed in subsequent studies (Hollingworth and Gadelhak, 1998). Metabolic Fate Diafenthiuron is rapidly metabolized and excreted mainly in the feces as urea and fatty acid derivatives formed from the carbodiimide intermediate. In mice treated with diafenthiuron intraperitoneally, the carbodiimide was found in several tissues. It was also proposed that a thiourea sulfoxide was formed, leading to the production of the corresponding formamidine after desulfoxidation (Petroske and Casida, 1995). Environmental Fate and Toxicity In birds, 8-day feeding studies indicated low toxicity with LCso values> 1500 mg/kg in bobwhite quail and mallard ducks. Diafenthiuron is strongly absorbed on soils and degradation is rapid varies from (half-life <1 hr to 1.4 days). It is much more toxic to fish in standard tests than to terrestrial vertebrates (Table 57.4). However, it is claimed that the rapid degradation of diafenthiuron in the field means that there is very little practical hazard to aquatic species (Anonymous, 1992b).
57.7 MITOCHONDRIAL UNCOUPLERS 57.7.1 INTRODUCTION
Mitochondrial uncouplers are a group of relatively simple but very effective toxicants. They act by discharging the proton gradient across the inner mitochondrial membrane and bypassing the concomitant phosphorylation of ADP to ATP. Thus the potential energy represented by the proton-motive force, i:::.p, is dissipated as heat and ATP synthesis ceases. Even worse from the point of view of cellular disruption is that as t::.p is lowered, the ATP synthase (complex V) operates in reverse to hydrolyze ATP. Thus even the ATP created by the continuation of glycolysis is rapidly destroyed with dire implications for cellular survival. The mitochondrion has become a machine for destroying rather than synthesizing ATP. The mechanism by which typical uncouplers achieve this effect is by acting as protonophores which shuttle protons across the impermeable inner mitochondrial membrane along their electrical and chemical gradients (McLaughlin and Dilger, 1980; Wallace and Starkov, 2000). This is a purely physicochemical process. The requirements for a molecule to be an effective uncoupler are only that it should be a lipophilic weak acid. In the simplest form of this mechanism, shown in Fig. 57.6, the uncoupler (generically shown as HA) tends to locate within the phospholipid bilayer membrane because of its
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CHAPTER 57
Pesticides Affecting Oxidative Phosphorylation
lipophilic nature. At the matrix face of the membrane it encounters an elevated pH and, as a weak acid, will tend to dissociate and release a proton to the matrix. The resulting anion (A-) passes back to the cytoplasmic face of the membrane along its concentration gradient and according to the charge differential across the membrane. The most powerful uncouplers have anions that are also lipophilic. Molecular features encouraging this are the presence of aromatic systems in the molecule conjugated with the acidic center which allow dispersion of the negative charge across the molecule, and shielding of the negative charge by bulky, lipophilic neighboring groups. At the cytoplasmic face, the anion (A -) encounters a more acidic environment, tends to become protonated, and, again according to its chemical gradient within the membrane, will diffuse back across to the matrix face where the proton is discharged. A proton is translocated across the membrane in each turn of this cycle. Additional, more complex variations on this basic theme are possible and are described by Wallace and Starkov (2000). Other mechanisms of achieving uncoupling exist such as redox cycling (as reviewed in Section 57.1.6.4). Organotin pesticides which have both respiratoy uncoupling and inhibitory properties are covered in Section 57.6.2. A protonophoric uncoupler thus acts as a catalyst which creates a proton short circuit across the inner membrane. It is not consumed as it uncouples, nor is a binding site required for its action. Indeed, protonophoric uncouplers are capable of discharging ion gradients and increasing the conductivity of simple artificial phospoholipid bilayer membranes (Bakker et aI., 1973; Cunarro and Weiner, 1975). This means that uncoupIers generally show, at most, very modest selectivity between mitochondria from different organisms, and useful selective toxicity is likely to be achieved primarily through differences in the pharmacokinetics of uncouplers between target and nontarget species. "Uncoupling" refers to the fact that in the uncoupled state, the rate of oxidation of substrates and passage of electrons down the transport chain to oxygen is no longer tightly coupled to and controlled by the rate of synthesis of ATP. Respiration is uncoupled from phosphorylation. In this condition, the electron transport chain is free to run at its maximum rate which is typically several-fold faster than in the coupled state. Thus, in the presence of uncouplers, substrates are oxidized and oxygen is consumed at a much higher rate than in the normal cell. The large amounts of energy released in this process are not conserved and are wasted as heat. The implications of the above mechanism of action are confirmed by the signs and symptoms of poisoning caused by typical uncouplers in both experimental animals and humans. These include fatigue, nausea and vomiting, hyperactivity, flushed skin, sweating, fever (particularly in larger animals including humans), dehydration and thirst, dyspnea, deep and rapid respiration, tachycardia, cyanosis, asphyxial convulsions, and coma. The very rapid onset of rigor mortis after death is an indicator for this mode of poisoning. For some uncoupIers, but not others, edema and spongy degeneration of the white matter of the brain and spinal cord are induced as ATP-
dependent ion pumps are disrupted and osmotic effects cause swelling of the myelin sheath (e.g., see bromethalin and chlorfenapyr, Section 57.7.2.3). Most of these signs and symptoms can be readily related to mitochondrial uncoupling action which causes decreased ATP levels, increased respiratory chain activity and oxygen consumption, the generation of heat, failed ionic regulation and consequent osmotic swelling, and, over time, the consumption of energy reserves within the body. Considerable weight loss is seen under conditions of chronic but sublethal exposure, and it is interesting that in the 1930s, the uncoupler 2,4-dinitrophenol (2,4-DNP) was quite widely employed as a weight reduction drug. Although effective in such a use, the several severe side effects, including peripheral neuritis, liver injury, and the induction of cataracts in perhaps 1% of those taking this therapy, led fairly rapidly to its discontinuation (Gasiewicz, 1991; Kurt et aI., 1986). In a study of the activities of a broad range of uncouplers in mice, Ilivicky and Casida (1969) found that they produced very similar signs of poisoning of the type described above, and these signs could readily be distinguished from those of mitochondrial respiratory inhibitors or nonmitochondrial poisons. There was a rough but persuasive correlation between the ability of these varied compounds to uncouple mitochondria in vitro and their toxicities to mice when administered intraperitoneally. Toxicity also correlated well with the degree of uncoupling of the mitochondria in brain in vivo but less well with the degree of uncoupling of liver mitochondria. Brain mitochondria were generally somewhat more sensitive to the action of these uncouplers than liver mitochondria when tested in vitro. On the basis of this work it seems reasonable to conclude that pesticidal compounds that are strong uncouplers in vitro cause most or all of their acute toxic effects in vivo through uncoupling, and that the nervous system is a particularly sensitive target for their actions. Most uncouplers are also inhibitors of the respiratory chain at somewhat higher concentrations than those needed for uncoupling, and a biphasic effect on respiration is typical with stimulation turning to inhibition as the concentration is increased. This inibition of respiration probably arises because uncouplers also inhibit specific steps in the respiratory chain at concentrations higher than those that cause uncoupling. There are probably compounds with the reverse properties (i.e., they are weaker as uncouplers than inhibitors, but the uncoupling phase will not seen in this case because respiratory inhibition predominates). Probable examples of this are some 2,6-dinitro-4-alkylphenols which clearly are potential uncoupIers but which generally are found to be respiratory inhibitors with mitochondria in vitro (Ilivicky and Casida, 1969). The ability of the isomeric 2-alkyl-4,6-dinitrophenols, which are potent uncouplers in vitro, to also inhibit respiratory complexes I and III (and also photosystem 11 in the photosynthetic apparatus) at higher concentrations has been studied by Saitoh et al. (1992) and Singer and Ramsay (1994). All three of these electron transport components have Q binding sites for which these phenols compete. The structural analogy between these phenols and Q was stressed by Singer and Ramsay (1994), particularly
57.7 Mitochondrial Uncouplers
the branching pattern in the alkyl chain that mimics the initial isoprenyl group in Q. In mitochondria, alpha branching of the alkyl group with a chain length of four or five carbon atoms gave optimal inhibition of respiration with Iso values between 0.1 and 1 I-lM. Saitoh et al. (1992) suggested that the function of the alpha branch was to hold the alkyl chain perpendicular to the plain of the phenyl ring.
57.7.2 PROPERTIES OF SPECIFIC COMPOUNDS 57.7.2.1 Dinitrophenols The toxicology of dinitrophenols has been reviewed by Gosselin et al. (1984) and Gasiewicz (1991).
Status and Uses These compounds were reviewed in considerable detail in the first edition of this handbook (Gasiewicz, 1991). Dinitrophenol derivatives were used as insecticides as early as 1892 but it was not until the 1930s that their value as herbicides was discovered. At one time, a considerable number of dinitrophenol derivatives were registered for use in the United States with a broad range of pesticidal activities including herbicidal, acaricidal, insecticidal, and fungicidal applications. These were typically 2,4-DNP derivatives with an added alkyl chain ortho to the hydroxy group and included such pesticides as binapacryl, DNOC, dinocap, dinobuton, dinoseb, dinocturon, and dinoterb. Both the free phenols and various metallic or amine salt forms have been used. In a number of cases the phenolic group was esterified with an acidic group which tends to decrease their phytotoxicity and mammalian toxicity and to increase formulation options. Such dinitrophenolic esters are regarded as propesticides and the 2,4dinitrophenolic component is liberated more or less rapidly in vivo to carry out its toxic uncoupling actions. As a result of their adverse toxicology, particularly their high acute toxicity and tendency to cause birth defects and chronic reproductive effects, and the advent of more desirable and selective pesticides, the dinitrophenols have been largely superceded in the United States and many developed countries, and only a few registered uses for these compounds now remain (e.g., in the United States only 2,4-DNP and dinocap are still registered for pesticidal use). However, other compounds in this class such as DNOC, dinocap, dinoterb, and dinobuton are still in use in some countries (Tomlin, 2000). Because of this steep decline in use and the fact that no important new or changed views of their toxicology have appeared in the last decade, the section here is limited to specific information on dinocap and an overview of the general properties of these pesticides. For more extensive coverage of the properties, and toxicology of other individual dinitrophenolic pesticides, the reader is referred to the previous edition of this handbook (Gasiewicz, 1991). Overview of Toxicology The dinitrophenolic pesticides are all derivatives of 2,4-DNP, a compound which itself has minor
1227
uses as a fungicide in fabric and leather preservation, but which is more familiar as the archetype uncoupler of mitochondrial oxidative phosphorylation. Pesticidal analogs typically have an additional alkyl substituent in the ring "ortho" to the hydroxy moiety which increases the lipophilicity of the molecule and thereby considerably increases uncoupling activity and toxicity compared to 2,4-DNP itself in both vertebrates and insects (Ilivicky and Casida, 1969; Miyoshi et aI., 1987; Miyoshi and Fujita, 1988). However, all these compounds cause their major toxicological actions in the same way as 2,4-DNP. The variations in lipophilicity and in the form of the phenol employed (free acid, alkali metal salt, alkoxyamine salt, or esterified) lead to variations in the rate of absorption and this impacts the toxicity of the different pesticidal dinitrophenol derivatives. Also, in the phenolic ester forms, the rate at which the ester is hydrolyzed in vivo also will impact the speed of action and ultimate toxicity of these propesticides. In general the dinitrophenolic pesticides tend to have relatively high acute toxicities by the standard of modern pesticides and lack sufficient selective toxicity due to the universality of uncoupling as a process which is deleterious to all multicellular organisms. The hazardous nature of some of these compounds has resulted in numerous examples of human poisoning and some deaths in addition to those observed during the use of these compounds as slimming agents. These adverse responses are particularly frequent with the more acutely toxic analogs such as DNOC (Gasiewicz, 1991). Other specific toxicological actions including the induction of cataracts by some members of the group (Gasiewicz, 1991) and the clear ability of others [e.g., dinocap (Section 57.7.2.1) and dinoseb (U.S. EPA, 1989c)] to act as teratogens, has raised particular concern over worker exposure and has led to severe regulatory actions. In general, these compounds do not appear to be general genotoxicants though there are scattered literature reports of positive tests for genotoxicity with dinitrophenolic pesticides. The dinitrophenols as a class do not appear to be carcinogenic.
Human Toxicology and Therapy The signs, symptoms and treatment of dinitrophenol poisoning are reviewed by Hallenbeck and Cunningham-Burns (1985), Gasiewicz (1991), and Reigart and Roberts (1999). In addition to the generalized signs of poisoning described above (Section 57.7.1), in which a range of neurotoxic effects are prominent, renal failure may occur rapidly with high doses. Hepatotoxicity marked by jaundice is also a possible sequel. Exposure by inhalation can lead to tightness of the chest and pulmonary edema in addition to systemic effects. The onset of symptoms is often rapid after exposure but can be delayed for 1-2 days. The toxic effects tend to be more severe at elevated environmental temperatures (Gasiewicz, 1991). Death may arise from hyperthermia, or respiratory or cardiovascular failure, and can occur within 24-48 hours. After exposure ceases, recovery tends to be protracted because of the relatively slow rate of clearance of many dinitrophenolic pesticides. Half-lives in human blood may be in the range of 5 to 7 days (Gasiewicz, 1991). The bright yellow
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CHAPTER 57
Pesticides Affecting Oxidative Phosphorylation
D i nocap
Cl
A H Cl 02N\QI~>i5> HO
Bramaxynil
Figure 57.14
I axyn i I
Niclasamide
Uncouplers of oxidative phosphorylation-PhenoIs.
color of these compounds leading to staining of sites of exposure (skin, hair, or clothing), and the urine and sclerae, is an important (but not absolutely definitive) indicator of exposure. Decontamination of sites of exposure and supportive treatment (control of fever by physical means, fluid, electrolyte, and energy replacement, and oxygen administration) are indicated. Salicylate antipyretics are contraindicated since these too can act as uncouplers and there is evidence that they may exacerbate dinitrophenolic poisoning in experimental animals. Other antipyretics are ineffective since the hyperthermia is a general tissue response rather than a disorder of central temperature regulation. Forced diuresis has been recommended for some phenolic uncouplers (Gasiewicz, 1991). No specific antidotes are available. Dinocap: General Properties and Uses Dinocap is a complex mixture of 2,6-dinitro-4-octylphenyl and 2,4-dinitro-6octylphenyl crotonates where the octyl moiety is a mixture of 1-methylheptyl, 1-ethylhexyl, and 1-propylpentyl isomers (Fig. 57.14). The commercial material has about a 2: 1 ratio of 6-octyl to 4-octyl isomers. The CAS Reg. No. is 131-72-6 (6-octyl, I-methylheptyl single isomer) or 39300-45-3 (mixed isomers). It was discovered by Rohm & Haas Co. The chemical composition of dinocap is described by Kirby and Hunter (1965) and Kurtz et aI. (1970). A series of samples of the technical product produced around 1970 were analyzed by Kurtz et al. They contained 72-77% 2,4- and 2,6-dinitrooctylphenyl crotonates, 4-7% mixed dinitrooctylphenols, 0.5-1 % mononitrooctylphenols, 1-4% crotonic acid, 2-6% octenes (used in the synthesis of the crotonates), several unknown constituents at less than 1% each, and 6-13% nonvolatiles which was mainly complex polymeric material without pesticidal activity. The samples averaged 83% as active ingredients (crotonates and phenols). Total6-octyl isomers constituted 68% of the active ingredients and 4-octyl isomers were the remaining 32%. Of this 68% as 6-alkyl components, 26% had 1-methylheptyl chains and the other 42% had either l-ethylhexyl or 1-propylpentyl (ratios not determined). The complexity of this mixture of isomeric esters and free phenols creates a considerable challenge for analytical chemists in assessing environmental residues of dinocap (Heimlich et aI., 1995).
Dinacap exists as a dark brown liquid, b.p. 138-140°C at 0.05 mm Hg, v.p. 7.5 x 10- 8 Pa (25°C), log P4.54, w.s. 4 ppm. Estimated half-lives for hydrolysis to the phenol are 3.5 years, 129 days, and 12.9 days at pHs 7, 8, and 9, respectively (HSDB, 2000). Dinocap is a foliar fungicide with acaricidal activity. It was first used in the 1930s. Trade names include Caprane, Crotonate, Crotothane, Karathane, Mildane, Mildex, and Sialite. Toxicology Overview A Special Review of all pesticide products containing dinocap was initiated in 1985 by the V.S. Environmental Protection Agency, primarily in response to studies with rabbits that revealed birth defects and chronic reproductive effects. This was completed in 1989 and the conclusion were reviewed (V.S. EPA, 1989b). Measures to reduced applicator exposure and warnings on the label regarding its potential teratogenicity were required for continued registration. The toxicology of dinocap from a regulatory standpoint has also been reviewed on several occasions by the FAOIWHO Joint Meeting on Pesticide Residues [e.g., see (1990c, 1999a)]. A brief addendum to the major report in FAOIWHO (1999a) is provided in FAOIWHO (1999b). These reports and the review by Gasiewicz (1991) lead to the conclusions below. Most studies were performed on technical dinocap mixture. Some recent studies have used purer materials (90-95%), and, in some cases, the single 6- and 4-(1-methylheptyl) isomers have been employed. These differences in isomer mix and purity may have led to differences in the toxic effects observed, since it appears that specific isomers or impurities are responsible for at least some toxic effects. Acute Toxicity This varies considerably with species (Table 57.5; FAOIWHO, 1990c). Rats and rabbits are relatively insensitive (acute oral LD50s vary from 510 to 3100 mg/kg in a range of reports) whereas mice and dogs are approximately lO-fold more sensitive (acute oral LDsos ranging from 50265 mg/kg). The intravenous LDso in rats is only 2.5 mg/kg, illustrating the high intrinsic toxicity of dinocap. The acute oral toxicities provided in FAOIWHO (1999a) are several-fold lower than those in Table 57.5 (mainly from Larson et aI., 1959), at least in part because dinocap of higher purity was utilized. Clinical signs of poisoning are typical of those already described for dinitrophenolic uncouplers in general. IrritationlSensitization Dinocap is an irritant to the eye, skin, and mucous membranes and a dermal sensitizer in both test animals and humans (FAOIWHO, 1999a; Gasiewicz, 1991). Subchronic Toxicity In 90-day dietary studies, pathological changes including hepatic necrosis were observed in dogs (6.25 mg/kg/day) and rabbits (30 mg/kg/day). Degenerative changes were also seen in the gastrointestinal tract and kidney in the rabbits. In a 28-day dietary study in mice, hepatocellular necrosis was similarly observed but only at doses (500 ppm and higher) that also caused deaths.
57.7 Mitochondrial Uncouplers
1229
Table 57.5 Acute Toxicity of Pesticides Acting as Mitochondrial Uncouplers to Selected Nontarget Species LDSO (mg/kg) Acute oral Compound
Rat (M; F)
Mouse
Quail
Duck
LCso (ppb) (24-96 hr)
LCso (mgll) Acute dermal
Acute inhalation
Rat or rabbit
Rat
Daphnia
2,4-Dinitrophenol
30-71
72
Dinobuton
140
2540
Dinocap
980; 1190
180
>4700 150d
3.4
23
200-1000
450-6000
5700
Dinoterb
62
25
DNOC
25-40
16-47
Bromoxynil
81; 93
110-160
Bromoxynil octanoate
400;238
Ioxynil
110
Ioxynil octanoate
190-390
200-230
Niclosamide
>5000
1500
Pentachlorophenol
146; 175
74-177
TFM
160; 141
150b
15.7
>5000 3.0
15-33
75 C
100-193
200
>3660
0.27
63-5000
19,220e
148
2050
>2000(M); 131O(F)
0.81
53-150
96
30!
1200
>2000
0.4
3300-8500
3900
1000!
1200
>912; 1240g
60 h
>1000
>1000
20
13-230
380
320
0.2
32-205
4000
>2000
200
0.6--37
Bromethalin
10.7; 9.1
5.0
4.6--11
2000
0.024
38-598
2-5
Chlorfenapyr
441; 1152
55
34
10
>2000
1.9
7.4-500
6.1
Fluazinam
>5000
>5000
1782
>4190
>2000
0.47
110-150
190
Sulfluramid
5000
473
LPOS
154; 154
42
81
>2000
>4.4
>10,000
210-390
>2000
0.21
4200-49,000
67,000
aRange of values from several species, most commonly including the rainbow trout and bluegill. bData for hen. CData for the sideswimmer (Gammarus fasciatus). dData for guinea pig. eData vary widely with source; see text. ! Data for pheasant. 8Data for mouse. h Species unknown.
Chronic Toxicity At a sublethal dose (150 ppm), weight reductions and atrophy of the testes was seen in an 18month feeding study in CD-l mice but no hepatotoxicity was noted in this case. In a 2-year dietary study in rats, doses of 125 mg/kg/day caused reduced growth and survival and spleen enlargement in males. The significance of the cataracts reported in a dietary study in white Peking ducklings at doses of 502500 ppm (Larson et aI., 1959) is difficult to ascertain since this action was inconsistent and not clearly dose-dependent (Gasiewicz, 1991). In a 2-year dietary study in dogs, retinal atrophy was seen at 60 ppm. However, this is not considered relevant for human risk assessment since it occurs secondarily to damage to the tapetum lucidum which is absent in humans. No retinal effects were seen in rats and mice which also lack this structure (FAOIWHO, 1999a). Carcinogenicity Dinocap was not carcinogenic in lifetime feeding studies in rats at dietary concentrations up to 2000 ppm, or in mice at dietary concentrations up to 200 ppm (about 35 mg/kg/day).
MutagenicitylGenotoxicity Dinocap showed no evidence of genotoxicity in a typical battery of tests in vitro and in vivo. However, in one study, a positive result was obtained in the Ames Salmonella assay (Moriya et aI., 1983). Reproductive Toxicity No specific reproductive toxicity was seen in rats in a multigenerational dietary study at concentrations up to 1000 ppm which were reduced to 400 ppm in the second generation due to high mortality in pups (FAOIWHO, 1999a). In another study in rats, decreased growth rates and survival of offspring in the second generation were recorded at dietary doses of 104-126 mg/kg/day (Fraczek, 1979 in Gasiewicz, 1991). Developmental ToxicitylTeratogenicity Based on a substantial series of developmental toxicity studies, the V.S. Environmental Protection Agency (V.S. EPA, 1989b) and the FAOIWHO JMPR panel (FAOIWHO, 1999a) both concluded that dinocap is teratogenic in animals, particularly in mice, and that it therefore poses a risk of teratogenicity to humans. In an initial exploratory study, Gray et al. (1986) assessed the teratogenicity of technical dinocap (84% active ingredi-
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CHAPTER 57 Pesticides Affecting Oxidative Phosphorylation
ents) in CD-1 mice. Cleft palate, extreme abdominal distension (ballooning) and twisting of the neck and tilting of the head (torticollis) were observed in the offspring when mice were dosed orally during pregnancy at 25 mg/kg/day. This dose also caused high postn3t::t1 mortality but was not toxic to the mothers. Torticollis was also seen at a lower incidence at 12 mg/kg/day and caused a variety of locomotor aberrations including repeated circling and rolling over. These behavioral effects were tentatively attributed to a reduction in the formation of otoliths in the inner ear. By contrast, teratogenicity was lacking in a parallel study in Sprague-Dawley rats at doses as high as 150 mg/kg/day, although an increase in the incidence of supernumerary ribs was observed at this dose (FAOIWHO, 1999a). In hamsters, only slightly retarded growth was observed which occurred at or near maternally toxic doses of 50 mg/kg/day or higher. The lack of clear teratogenic effects in rats and hamsters was confirmed by Rogers et al. (1988) who noted significant differences in the ratios of adult to developmental toxicities (AID ratios) between the species tested. For mice, an AID ratio of about 8.6 was calculated compared to the hamster and rat with AID ratios of 1 or less, indicating that developmental toxicity only occurred in the presence of maternal toxicity. In a follow-up study in mice using a standardized experimental protocol for assessing developmental toxicity, Rogers et al. (1986) observed developmental effects after oral administration at the lowest dose of 5 mg/kg/day. Significant maternal toxicity only occurred at 80 mg/kg/day. A low incidence of cleft palate in mice at 5 mg/kg/day increased to an incidence of 75% at 40 mg/kg/day. Reduced fetal weight and moderate to severe hydronephrosis was also recorded beginning at 5 mg/kg/day. Increases in supernumerary ribs appeared at 20 mg/kg/day and a few examples of exencephaly and umbilical hernias were also observed in mice at higher doses. Otolith development was not studied in this protocol. Clearly dinocap causes mUltiple types of teratogenic effects in mice. A further investigation of the unusual behaviora1 effects seen in weanling mice including a study of the effect of prenatal exposure to dinocap on the swimming ability in the offspring, head-tilting (torticollis), and reductions in otolith development was conducted by Gray et al. (1988). Considerable difficulties in swimming were observed in mice with torticollis and even in some animals without this postural abnormality. The authors confirmed that the behavioral deficits arise because of a partial or complete failure to develop oto1iths and concluded that the otolith status is the most sensitive endpoint in detecting this type of teratogenesis. In a special study with CD-1 mice conducted by Rogers et al. (1989), the reduction in fetal otolith formation was found to be dose-dependent with a threshold at 10 mg/kg/day. Maternal toxicity was not observed until the dose reached 60 mg/kg/day. By contrast, otoliths were affected in the Syrian golden hamster only at 100 mg/kg/day, a dose that also caused severe maternal and fetal toxicity. This again demonstrates the relatively high sensitivity of mice to the teratogenic actions of dinocap. The induction of effects on otolith production has rarely been reported before in teratogenicity studies,
but as Rogers et al. point out, it would not be observed directly in standard testing protocols. No mechanism has been proposed by which dinocap might affect otolith development. Interestingly, Rogers et al. (1987) discovered that two major ingredients of dinocap, the individual isomers 2,6dinitro-4-(1-methylheptyl)phenyl croton ate and 2, 4-dinitro-6(l-methylheptyl)phenyl crotonate (each at 95% purity) were not active as teratogens in mice either alone or in combination. At the same dose (25 mg/kg/day) technical dinocap gave the typical range of teratogenic effects seen in other studies. The technical dinocap had 84% active ingredients but the impurities were not characterized. Presumably the composition of this sample resembles that described by Kurtz et al. (1970) quite closely. This raises a critical question regarding the nature of the teratogenic agent(s) in this mixture and the possible role of other alkyl chain isomers or minor impurities in the technical material. This significant question does not appear to have been resolved. However, the positive teratogenic respone in CD-1 mice was confirmed in all essential details by an additional study conducted by the registrant using a higher purity sample of dinocap [94.4% compared to 84% in the studies of Gray et al. (1986) and Rogers et al. (1986)]. The LOEL for developmental effects in this study was 10 mg/kg/day (FAOIWHO, 1999a). Unfortunately, the detailed compostion of the purified material is not provided in this publication. Similar teratogenic effects including cleft palate and decreased otolith production were seen in CD-1 mice after dermal exposure to a formulated version of dinocap (Karathane LC XF) at 25 mg active ingredient/kg/day. This dose caused no maternal toxicity (FAOIWHO, 1999a). In rabbits, teratogenicity studies revealed an increased incidence of skeletal malformations such as vertebral asymmetry and malformed ribs at 48 mg/kg/day, but this dose also caused obvious maternal toxicity. In a previous study (Costlow et aI., 1986; FAOIWHO, 1990c), hydrocephaly and neural tube defects were found in rabbits at doses of 3 mg/kg/day and higher. Such effects were completely lacking in the second study, perhaps because the purity of the dinocap had been increased from 84% to 95.4% (FAOIWHO, 1999a). Costlow etal. (1986) found no developmental or teratogenic effects in rabbits when dinocap was applied demally. The mouse teratogenicity data were used by the FAOIWHO JMPR panel to set an AD! of 0.008 mg/kg using an elevated 500-fold safety factor since malformations were seen after both dermal and oral exposures and in three test species (FAOIWHO, 1999a). Immunotoxicity An assessment of the immunotoxicity of dinocap in female C57BLl6J mice administered by gavage for 7 to 12 days revealed adverse effects at doses of 25 mg/kg/day. These effects included decreased thymus weight, increased spleen weight, a reduction in the proliferative responses of lymphocytes to concanavalin A and phytohemaglutinin, and a suppression of the IgM and IgG plaque-forming response to sheep red blood cells. Using cultured mouse thymocytes, dinocap at 10 I-lg/ml also caused a depression in the proliferative
57.7 Mitochondrial Uncouplers and mitogen-stimulated responses with no evidence of cytotoxicity, Although adverse immunological effects were seen in this study, it was concluded that these were relatively modest and only seen at high doses in vivo (Smialowicz et al., 1992).
Human Toxicology No serious human poisonings by dinocap appear to have been reported (FAOIWHO, 1990c, 1999a), although Gasiewicz (1991) cites a single case of dinocap poisoning in an agricultural worker which was marked by allergic dermatitis, dyspnea, and thirst, Nine days after these symptoms an outbreak of vesicles over the entire body followed. Complete recovery was attained in 1 month after cortisone treatment. The level of exposure is undefined but it must have been considerable in view of the symptoms described. The estimated daily human intakes of dinocap produced by several dietary models represent 0-2% of the ADI which indicates a very low degree of risk to the general population through food consumption (FAOIWHO, 1999b). A study of worker exposure to dinocap in California estimated joint dermal and inhalation exposures ranging from 0.028 to 0.49 mg/kg/day depending on the type of protective clothing worn (Wang, 1988). Biochemical Mechanism of Action It is reasonable to assume that dinocap requires hydrolysis to release the free phenol in order to cause uncoupling and that this occurs rapidly in vivo [e.g., Ilivicky and Cas id a (1969) found strong uncoupling of brain and liver mitochondria in vivo within 30 minutes of administering a toxic intraperitoneal dose of dinocap to mice]. By comparison with 2,4-DNP, the predominant phenolic isomer of dinocap, 2,4-dinitro-6-octylphenol, is likely to be a strong uncoupler [e.g., Hemker (1962) found that one phenolic component of dinocap had an uncoupling activity 7- to 25fold higher than that of 2,4-DNP in rat liver mitochondria]. Similarly, Ilivicky and Casida (1969) concluded that 6-secbutyl-2,4-DNP is an uncoupler with an activity about 50-fold greater than that of 2,4-DNP. However, there may be complications in this apparently simple story. As pointed out by Corbett et al. (1984), the 4-alkyl-2,6-dinitrophenol components of the dinocap chemical mixture may act as respiratory inhibitors rather than uncouplers. Ilivicky and Casida (1969), studying the related sec-butyl analogs of 2,4-DNP, found that the 4-alkyl2,6-dinitro analog was not an uncoupler at any concentration, but acted as a strong inhibitor of oxphos with appreciable activity at 0.1 f.1.M. In mice, it caused an entirely different suite of poisoning signs than uncouplers that were characterized by ataxia and, sometimes, convulsions, but no immediate rigor mortis after death. As discussed in the Introduction to this section, even the 2,4-dinitro-6-alkyl components of the mixture can inhibit respiration if concentrations rise further beyond those that cause uncoupling. The effects of dinocap in vivo may therefore tend to lie in balance between respiratory uncoupling and inhibition, depending on the relative contents of the 4- and 6-alkyl phenolic forms, their relative rates of release by ester hydrolysis and removal by further metabolism, and the concentrations that they achieve in the tissues. In the case of the dinocap, the signs of poisoning in animals clearly
1231
indicate that uncoupling predominates in most situations. However, there may be exceptions. It is interesting to note in this context that: (1) Larson et al. (1959) observed that dinocap caused an increase in respiration rates typical of an uncoupler after oral adminsitration in female rats but no increase in male rats, whereas 2,4-DNP stimulated respiration in both sexes, and (2), that whereas the 6-octyl isomers of dinocap are better as acaricides, the 4-octyl isomers are superior as fungicides (Tomlin,2000).
Absorption, Metabolism, and Excretion Dinocap [as the purified 2,4,-dinitro-6-(1-methy1heptyl) isomer] was well absorbed (60-70%) after a single oral dose of 25 mg/kg in mice and reached peak plasma concentration at 2 to 6 hr after dosing. The half-life for plasma clearance was about 6 hr. After dermal application at 25 mg/kg, the peak plasma concentration was about 25% of that with the oral dose and was reached after 6 to 8 hr, but the clearance rate was the same. This indicates a rapid but incomplete (25% in 4 hr) absorption through the skin. A study of the penetration of dinocap through isolated mouse and human skin samples indicated a more rapid penetation in the mouse, which is typical, but it did not model the in vivo data closely and the results are therefore of dubious relevance quantitatively (FAOIWHO, 1999a). These pharmacokinetic results after oral dosing in mice were quite similar to those obtained in previous studies in rabbits, using the same single isomer of dinocap (FAOIWHO, 1990c). This is extensively metabolized and excreted in rabbits with biphasic elimination kinetics, giving half-lives of about 3 and 44 hr after oral dosing. The elimination half-lives after dermal dosing were more variable but were approximately 10fold higher than those determined orally. Total absorption was 60-69% of the dose orally and 4 to 9% dermally. A dermal penetration study in rhesus monkeys also indicated incomplete absorption with 5-20% of the dose recovered in the excreta depending on the application conditions (FAOIWHO, 1999a). The metabolic fate of the l-methylheptyl component of dinocap in rats and mice is reviewed in detail in FAOIWHO (1999a). After an oral dose of 100 mg/kg, rats eliminated 30% of the dose in the urine in 24 hr. After an oral dose of 25 m/kg mice, eliminated 58% in the 24 hr urine. At least 12 metabolites were identified in each species. Of the metabolites found in rat, 85% were also found in the mouse, and 70% of those found in the mouse were also found in rat. Dinocap was metabolized by extensive initial hydrolysis, oxidation of the two terminal carbon atoms of the side chain, and ,B-oxidative chain shortening. Sulfate conjugation products of the phenol were found in mice and products of nitroreduction followed by N -acety lation in rats. Environmental Fate and Toxicity Dinocap is moderately toxic to birds but its toxicity to fish and aquatic invertebrates is extremely high (Table 57.5). It binds strongly in soils with a high clay and organic matter content with an estimated Koc value of 44,000. It therefore has a low potential to move into ground water. In soil it is degraded both by microbial and pho-
1232
CHAPTER 57
Pesticides Affecting Oxidative Phosphorylation
tochemical actions with a field half-life of 4-6 days (HSDB, 2000). 57.7.2.2 Other Phenolic Pesticides with Uncoupling Activity
A number of other phenolic pesticides have uncoupling activity which plays a significant part in their toxicity. Since it is only necessary to develop a weak acid with lipophilic properties to obtain an active uncoupler, a variety of phenolic structures with a combination of electron withdrawing and lipophilic groups fit within this category. These compounds include nitrophenols such as the sodium salt of p-nitrophenol itself, which, in combination with sodium nitroguicaolate, has uses as a plant growth regulator under the trade name of Atonik, and 3-trifluoromethyl-4-nitrophenol, a compound with limited uses as a specific lampreycide in the Great Lakes region of the United States. Further information on its uses and toxicology is available from a recent U.S. Environmental Protection Agency publication (U.S. EPA, 1999d). A detailed survey of the toxicological properties of p-nitrophenol is also available in U.S. EPA (1998e). Several halophenols such as pentachlorophenol and herbicides in the bromoxynil family are also powerful uncouplers. Since the 1930s, pentachlorophenol has been used in huge amounts for the preservation of wood products against fungal and insect pests, as a nonselective herbicide, as a cotton defoliant, as a molluscicide, and as a wide spectrum fungicide and bactericide. In the United States it now retains only its use in wood preservation. It is used both as the free phenol and its sodium salt. The compound is covered in detail in Chapter 65. Halogenated salicylanilides such as niclosamide, a widely used common anthelmintic and molluscicide, also have strong uncoupling properties. Acute toxicity data for some of these other phenolic uncouplers are included in Table 57.5 for comparative purposes. The more economically important ones are described in further detail below. Bromoxynil: General Properties and Uses 3,5-Dibromo-4hydroxybenzonitrile (Fig. 57.14), CAS Reg. No. 1689-84-5, was developed by Amchem Products Inc. and May & Baker Ltd. The herbicidal activity of bromoxynil and the closely related herbicide ioxynil were described independently by Carpenter and Heywood (1963) of May & Baker, and by Wain (1963) of Wye College, while they were also under study by Amchem. The chemistry, acute toxicity, and biological properties of bromoxynil have been reviewed by Carpenter et al. (1964) and its development has been described by Heywood (1966). Bromoxynil is also utilized as its octanoate ester (CAS Reg. No. 1689-99-2) and its potassium salt (CAS Reg. No. 296168-4). Use has also been made of the heptanoate ester. The esters are propesticides which are readily converted to bromxoynil in vivo by esterase activity. Bromoxynil has also been used as its butyrate ester but this was voluntarily withdrawn from the U.S. market in 1989 due to concerns regarding its developmental toxicity (U.S. EPA, 1998d).
Bromoxynil is an active metabolite of the herbicide bromofenoxim (3,5-dibromo-4-hydroxybenzaldehyde 2,4-dinitrophenyloxime, CAS Reg. No. 13181-17-4) which is metabolized in mammals to a mixture of2,4-DNP and bromoxynil, both active as mitochondrial uncouplers. This compound (trade name Faneron) now appears to no longer be marketed (Tomlin, 2000). Bromoxynil is a white crystalline solid, m.p. 194-195°C, v.p. 6.3 x 10-6 Pa (20°C), w.s. 130 ppm (20-25°C), log P2.8 (unionized phenol), HLC 1.34 x 10-5 Pam3 mol-I, pKa 3.94.2. The octanoate ester is a waxy cream solid, m.p. 45-46°C, v.p. 1.9 x 10- 4 Pa (25°C), w.s. 3 ppm (25°C) (Tomlin, 2000) [also cited as 80 ppm (25°C) by Wauchope et al. (1992)], log P5.4. The hydrolysis half-lives at pH 5, 7, and 9 are 34.1, 11.5, and 1.7 days, respectively. The properties of the heptanoate are very similar. The potassium salt has a water solubility of 61,000 ppm (20-25°C). Bromoxynil and its esters are widely used as selective contact herbicides for postemergent control of broad-leaved weeds. They are often included in mixtures with other herbicides to broaden the spectrum of control. Common trade names include Brominal, Bromotril, Buctril, Certrol, Combine, Connect, Emblem, Labuctril, Merit, Pardner, Sabre, Terset, Toplan, and Torch. An estimated 2.5 to 3 million pounds of all forms of the active ingredient are used annually in the United Sates (U.S. EPA, 1998d). Toxicology Profile The primary sources of data are Cal EPA (1995) and U.S. EPA (1998d). The U.S. Environmental Protection Agency (U.S. EPA, 1998d) concluded that there are no important toxicological differences between bromoxynil and its esters since the esters are rapidly cleaved to the free phenol in vivo. However, this conclusion has been questioned by the California Environmental Protection Agency (Cal EPA, 1995) based on differences in their pharmacokinetics, toxicological effects, and potencies. Acute Toxicity There are several sources of data for these values [e.g., Carpenter et al. (1964), Ahrens (1994), U.S. EPA (1998d), and Tomlin (2000)]. While they are in reasonable agreement in most cases, they are rarely identical and there are a few instances where the values offered appear to be incompatible. These are noted below. The data in Table 57.5 include the range of values recorded in these sources where they are reasonably close. Bromoxynil is highly toxic to mammals and birds (Table 57.5). The data for the acute oral toxicity to the rat are from U.S. EPA (1998d). Alternative values for the acute oral LD50 in rats range from 190 mg/kg (Carpenter et aI., 1964) up to 440 mg/kg (Ahrens, 1994). The acute oral LD50 values for guinea pigs, cats, and rabbits are 63, 75, and 260 mg/kg, respectively (Carpenter et aI., 1964). The acute toxicity of the potassium salt of bromoxylin is not notably different from that of the free phenol (Tomlin, 2000). The octanoate ester has a lower acute oral toxicity to rats than the phenol but tends to have a greater acute dermal toxicity since it is absorbed more readily. Its acute oral LD50 to
57.7 Mitochondrial Uncoup1ers rabbits is 325 mg/kg, only slightly more than that of the phenol at 260 mg/kg (Tomlin, 2000). IrritationlSensitization The phenol is a moderate eye irritant but not a skin irritant or sensitizer. The octanoate is a mild skin and eye irritant and also acts as a skin sensitizer. Subchronic Toxicity In subchronic (90 day) dietary studies in rats, very high mortality was observed at 1456 ppm [168 mg/kg/day (M) and 250 mg/kg/day (F)]. At the next lower dose (58 or 76 mg/kg//day) decreased body weight and changed clincial blood chemistry were observed. In a similar study with mice, hepatoxocity in the form of hepatocellular hypertrophy, degeneration, and vacuolization was seen with a LOEL of l3 mg/kg/day for males and 39 mg/kg/day for females. In dogs, doses of 16 mg/kg/day caused mortality. At lower doses, diarrhea, panting, salivation, an unsteady gait, and hematological effects were recorded. The LOEL was 1 to 5 mg/kg/day. Parallel subchronic studies with the octanoate ester revealed a possible increase in necrosis of thymic lymphocytes and cardiac myofibers in rats at a dietary concentration of 1100 ppm [91 mg/kg/day (M), III mg/kg/day (F)]. A subchronic dietary study of the octanoate in dogs gave results qualitatively similar to those with the free phenol. Chronic Toxicity In a I-year study in dogs with the phenol, results similar to those in the subchronic study were obtained with a LOELINOEL of 1.5 mg/kg/day. No gross or histopathological changes were seen. This study has been used by the U.S. Environmental Protection Agency to set the reference dose (AD!) for bromoxynil at 0.015 mg/kg/day. A 2-year study in Fischer 344 rats gave no toxicological endpoints at the highest dose administered (5 mg/kg/day). Using Sprague-Dawley rats and higher doses, spongiosis hepatitis was seen in males at 8.2 mg/kg/day together with foci of cellular alteration at 28 mg/kg/day. No other notable toxic effects were seen in either sex. Lifetime feeding studies in CD-l mice resulted in a variety of signs of hepatotoxicity at 12-14 mg/kg/day and an increase in neoplasms at 46-53 mg/kg/day. Carcinogenicity Bromoxynil did not act as a carcinogen in the long-term rat studies. However, it was positive for liver tumors in two chronic dietary studies in mice. In Swiss white mice, males showed a dose-resulted increase in incidence of liver adenomas and carcinomas (approximately 50% of each type at higher doses). The effect was statistically significant at l3 mg/kg/day. The CD-l mice showed a similar increase in hepatocellular adenomas and carcinomas. This was observed in both sexes but the effect was much more severe in males with an increase occurring at the lowest dose administered (3.1 mg/kg/day). Based on these results, and the positive results in some genotoxicity tests, the U.S. Environmental Protection Agency has classified bromoxynil as a Class C (possible human) carcinogen with a value ofO.103 (mg/kg/day)-l (U.S. EPA,1998d).
QI
1233
MutagenicitylGenotoxicity Bromoxynil gave primarily negative results in a series of 10 in vitro and in vivo tests for mutations and chromosomal effects. However, it was positive in a DNA repair test in E. coli and in a forward mutation test in mouse lymphoma cells. It also caused chromosomal aberrations in Chinese hamster ovary cells in vitro. The latter two response were seen only with metabolic activation. The octanoate was negative in a smaller array of assays. Developmental ToxicitylTeratogenicity Rats demonstrated reduced ossification and an increased incidence of supernumerary ribs at an oral dose 12.5 mg/kg/day which was not toxic maternally. Similar responses were seen in rats in another oral studies at 5 mg/kg/day and in a dermal study at 50 mg/kg/day. Studies with rabbits gave the same effects with a LOEL of 15 mg/kg/day for the increased incidence of supernumerary ribs. In mice the same result was obtained but only in the presence of maternal toxicity. Other malformations, some severe, and increased resorptions were observed in all these studies but only at higher doses with clear maternal toxicity. The octanoate ester gave generally similar results with a LOEL for increased supernumerary ribs in a dermal study in rats of 15 mg/kg/day; however, a dermal study in rabbits was negative for developmental toxicity although significant skin irritation was observed. Developmental toxicity studies with mice, using the free phenol, and in rats, using both the phenol and the octanoate, have also been described by Rogers et al. (1991). The only developmental effect observed was an increased incidence of supernumerary ribs. This occurred with the phenol in mice at 32 and 96 mg/kg/day and with both the phenol and the octanoate in rats at an equimolar dose of 15 or 22 mg/kg/day. In each case the extra ribs were seen only at doses that were also maternally toxic. The difficulties in interpreting the observation of supernumerary ribs in developmental toxicity studies are discussed by Chernoff (1990) and Chernoff et al. (1991) in the context of these studies with bromoxynil. Supernumerary ribs are seen in control animals and an elevated incidence is often observed in reproductive toxicology studies. This increase may be related to maternal toxicity, particularly in mice. The extra ribs tend to disappear during subsequent development in rats but not mice The authors conclude that their significance as an indicator of developmental toxicity and the extrapolation of results to other species remains problematical. Based on these indications of teratogenicity, the U.S. Environmental Protection Agency has maintained a lO-fold additional safety factor for bromoxynil to protect females of reproductive age (U.S. EPA, 1998d). Reproductive Toxicity No adverse effects on reproductive performance were observed with bromoxynil in a two-generation study in rats at doses up to 21 mg/kg/day. Decreased body weight gain during lactation and delayed eye opening were seen as developmental endpoints at this dose but this also caused lowered weight gain in the parents. A dermal study with bromoxynil octanoate in rats did not result in reproductive
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Pesticides Affecting Oxidative Phosphorylation
or developmental effects at doses that were not also systemically toxic to the parents. Significant skin irritation did occur at 100 mg/kg/day. Endocrine Effects Van den Berg et a!. (1991) investigated the potential of a number of halogenated compounds to lower plasma thyroid hormone level through competiton with hormone transporters. Bromoxylin was a strong competitor for the thyroxine binding site of transthyretin and reduced the levels of both thyroxine (T4) and triiodothyronine (T3) in the plasma of rats. The interpretation of this observation in terms of human risk is complex and the effects on transporter binding and thyroxine levels at least may not be clinically significant (Brucker-Davis, 1998). Human Toxicology Only a very limited number of instances of human poisoning by bromoxynil have been reported (U.S. EPA, 1998d). Eye and skin illnesses predominated in California's Pesticide Illness Surveillance Program from 1982 to 1993 and these averaged about one per year for bromoxynil. Four workers in a manufacturing plant making both bromoxynil and ionoxynil developed typical symptoms of uncoupler poisoning including excessive perspiration, thirst, fever, emesis, myalgia, and weight loss. This was attributed to an increase in the production levels of the herbicides without adequate increases in ventilation which created excessive exposure to fine dusts. The effects reversed rapidly after exposure ceased (Con so et a!., 1977). Based on current data, no significant concerns were expressed by U.S. Environmental Protection Agency regarding dietary risks to human populations from anticipated bromoxynil residues in food and water (U.S. EPA, 1998d). Biochemical Mechanism of Action Bromoxynil's primary toxic effects in plants are produced through the inhibition of photosynthesis in photosystem 11 with uncoupling of photophophorylation as a second possible contributing effect (Ahrens, 1994). It is a fairly potent mitochondrial uncoupler in vitro (Parker, 1965) with a UC50 value in rat liver mitochondria of 3.2-5 Il-M, an activity which is about 5- to lO-fold higher than that of 2,4-DNP. The signs and symptoms of acute poisoning in vertebrates (including humans) are in reasonable accord with uncoupling being the primary mechanism of toxic action. Absorption, Metabolism, and Elimination A study with the octanoate ester in rats showed that after an oral dose, uptake was moderate. Peak plasma levels were attained 7 to 10 hr after dosing. Most of the dose was eliminated within 7 days, primarily in the urine. The ester was rapidly and completely converted to the free phenol and excreted in the urine either as the phenol or its conjugates. The heptanoate is also converted to the phenol rapidly and completely in vivo. Dermal uptake studies in rats show an absorption in 24 hr of 11 tol8% for the octanoate depending on concentration, and about 3% for the phenol. A toxicokinetic study with bromoxynil itself was conducted in rats by Stahler et a!. (1991).
Environmental Fate and Toxicity Among the bird species tested, the pheasant appears to be the most sensitive to bromoxynil with an acute oral LD50 of 50 mg/kg (Table 57.5). Hens are less sensitive with an LD50 of 240 mg/kg (Carpenter et a!., 1964). Although bromoxynil octanoate has about the same acute toxicity as the phenol to quail it is much less toxic to mallards (Table 57.5). Short-term feeding studies in birds with both bromoxynil and its octanoate ester revealed only slight toxicity with LC50 values above 1000 ppm in the diet in all cases. In an avian reproduction study in mallard ducks a NOEL of 102 ppm was determined with adverse effects including a lower number of eggs laid, fewer live embryos, and regression of the ovary. Bobwhite quail were less sensitive than the mallards. Reported values for the toxicity of the phenol (and its potassium salt) to fish vary widely from an LC50 of 63 ppb for catfish to 4000 ppb for the bluegill sunfish and 5000 ppb for harlequin fish which seem to be unusually resistant to chemicals in this class. The values reported by Ahrens (1994) and Tomlin (2000) are generally much lower (50-500 ppb) than those in U.S. EPA (1998d) (2000-4000ppb). The LC50 for Daphnia is given as 12,500 ppb in Tomlin (2000) and 19,200 ppb in U.S. EPA (1998d) but as 110 ppb in Ahrens (1994). The reason for these large differences is unclear. Carpenter et al. (1964) note that the toxicity of bromoxynil to fish is very dependent on the hardness of the water, presumably because of decreased solubility of the calcium salt. In soft water the LC50 to harlequin fish was 5000 ppb. In hard water the LC50 rose to 63,000 ppb. The octanoate is uniformly more toxic to aquatic species than bromoxynil phenol, probably because its notably higher lipophilicity leads to faster uptake (Table 57.5). The high aquatic toxicity of the octanoate is also supported by data for other estuarine and marine organisms (e.g., it has an LC50 for mysid shrimps of 65 ppb and an EC50 for the inhibition of shell formation in the Eastern oyster of 155 ppm). The BCF for the octanoate in bluegill sunfish was 230-fold and depuration occurred in 14 days which indicates a limited capability for bioconcentration. Decreased juvenile survival was seen in a chronic exposure study with fathead minnows at concentrations as low as 18 ppb for the octanoate, and life-cycle studies with Daphnia magna indicated that survival, reproduction, and growth were diminished at 5-6 ppb. Despite this high toxicity, the U.S. Environmental Protection Agency concluded that under normal conditions, the risk to birds and aquatic vertebrate from use of the octanoate ester is likely to be low and to be moderate for aquatic invertebrates (U.S. EPA, 1998d). Bromoxynil octanoate hydrolyzes rapidly in the environment. Half-lives of 1 to 14 days have been recorded in the field and 2 days in a laboratory study which are in general agreement with a previous study in three soil types (Ingram and Pullin, 1974). Both the octanoate and the resulting phenol are also subject to significant microbial and photolytic degradation. Nolte et a!. (1995) found that in the dark under simulated groundwater conditions, bomoxynil was very stable with less than 10% degradation in 1 month. The octanoate was rapidly hydrolyzed to the phenol under these conditions. However, exposure to sunlight degraded bromoxynil rapidly with significant loss in
57.7 Mitochondrial Uncouplers
4 hr. Millet et al. (1998) have calculated photolytic half-lives for bromoxynil in water that range from 0.12 days in summer to 4 days in winter at a 50° latitude. The photolytic half-life of the octanoate on soil was 2.6 days. Microbial metabolic reactions include hydrolysis of the ester and nitrile groups and debromination. The debromination of bromoxynil in simulated groundwater conditions occurs by halide exchange reactions in which chlorine replaces the bromine atoms (Grass et al., 2000). The phenol has the physicochemical characteristics that would allow leaching to ground water under favorable circumstances but its rapid degradation in the soil makes this less likely and this is borne out by the extreme rarity of detections of bromoxynil in surveys of well water in the United States (U.S. EPA, 1998d). It is detected more frequently in surface waters due to runoff (1.1 % of samples), generally at sub-ppb levels, but its persistence there is low with a half-life of <12 hr. The octanoate has characteristics [Kd = 190-300 mUg, Koc estimated as 10,000 (Wauchope et al., 1992)] that suggest strong binding to soils and low leaching capability. However, other estimates of these values [Kd = 7.0 and Koc = 1003 (U.S. EPA, 1998d)] indicate a greater degree of soil mobility. Ioxynil: General Properties and Uses 4-Hydroxy-3,5diiodobenzonitrile (Fig. 57.14) has CAS Reg. No. 1689-83-4. Like its close relative, bromoxynil, ioxynil is also used in several forms, as the free phenol, as the sodium salt (CAS Reg. No. 2961-62-8), or in esterified forms, particularly as the octanoate (CAS Reg. No. 3861-47-0). Water- and oil-soluble salts of the phenol are also utilized in combinations with amines such as dimethylamine and triethanolamine. loxinyl was developed by May & Baker Ltd. and Amchem Products Inc. and is described by Carpenter and Heywood (1963) and Wain (1963). The chemistry, acute toxicity, and bi010gical properties of ioxynil have been reviewed by Carpenter et al. (1964), and its development was described by Heywood (1966). Ioxynil has also been referred to under the common name bantrol (HSDB, 2000). The free phenol is a colorless solid, m.p. 212-213°C, v.p. < 1 X 10- 3 Pa (20°C) [estimated as l.9 x 10-5 Pa (25°C)], w.s. 50 ppm (20°C, pH unstated) and 130 ppm (25°C), log P3.43 (25°C, unionized phenol), 0.90 at pH 6.5, pKa 3.96, HLC 5.5 x 10- 10 atm m3 mole -1. The l-octanoUwater partiton coefficient and pKa value were determined by Chamberlain et al. (1996). The octanoate ester is a white solid, m.p. 59-60°C, v.p. 3.7 x 10- 3 Pa (l05°C), practically insoluble in water. It is easily hydrolyzed to release ioxynil under alkaline conditions. The alkali metal salts are readily soluble in water (e.g., the sodium salt has a solubility of 14% at 20-25°C). Ioxynil and its derivatives are selective contact postemergent herbicides. They are often combined with other herbicides, particularly chlorophenoxy compounds and sometimes with bromoxynil, to extend the herbicidal range. Trade names for the various forms of ioxynil include Actril, Actrilawn, Bentrol, Belgran, Certrol, Cipotril, Dantril, lotox, Iotril, Oxytril, Sanoxynil, Totril, Toxynil, and Trevespan.
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Toxicology Profile Published information on the toxicology of ioxynil is incomplete. The toxicology of ioxynil was previously reviewed in the first edition of this Handbook (Stevens and Sumner, 1991). The other primary sources of data are HSDB (2000) and Ahrens (1994). The biological properties of ioxynil including aspects of its toxicology and the occurrence of occupational poisoning have been reviewed (Anonymous, 1991 ). Acute Toxicity Ioxynil is a highly toxic compound to mammals with a potency broadly equivalent to that of bromoxynil (Table 57.5). Ioxynil is also highly toxic by acute oral exposure to cats, guinea pigs, and rabbits (LDso values 75, 76, and 180 mg/kg, respectively; Carpenter et al., 1964). The value for the inhalation toxicity to rats in Table 57.5 is from Ahrens (1994). It appears to be more in accord with the other data than that in Tomlin (2000) of >3 mg/l. The octanoate ester is somewhat less toxic than the phenol to rats. It has moderate toxicity by the dermal route. The sodium salt generally has similar toxicity to that of the free phenol, although a surprisingly high acute percutaneous LDso in rats for the formulated salt of 210 mg/kg has been cited (Tomlin, 2000). IrritationlSensitization Ioxynil is a mild eye and skin irritant, but it is not a skin sensitizer. Subchronic Toxicity In a 90-day feeding study in rats, ioxynil caused no adverse effects at dietary concentrations as high as 111 ppm (about 5.4 mg/kg/day) (Hayes, 1982). Dogs given ioxynil octanoate at 4.5 mg/kg/day for 3 months also showed no adverse effects (Ahrens, 1994). In a 30-week dietary study in dogs, weight loss and anemia were observed with a NOEL of 10 mg/kg/day. Chronic Toxicity In an 18-month dietary study in mice, liver and thyroid hypertrophy were observed at 30 and 100 ppm (approximately 5 and 15 mg/kg/day). A 24-month dietary study in rats also revealed thyroid toxicity at the same dietary concentrations. The NOEL was found at 10 ppm (0.5 mg/kg.day). Concern regarding the thyrotoxic and possible teratogenic effects of ioxynilled to regulatory review and the cancellation of many domestic uses of ioxynil in the UK (Flanagan et al., 1990; Ogilvie and Ramsden, 1988). Carcinogenicity
No data were found.
MutagenicitylGenotoxicity Ioxynil was nonmutagenic in an extensive battery of tests for mutations, chromosomal aberrations, and DNA damage and repair using both bacterial and mammalian systems. In additional studies, Carere et al. (1978) and Moriya et al. (1983) also obtained negative results with ioxynil in the Ames Salmonella assay. Reproductive Toxicity Ioxynil was negative for reproductive toxicity in a study in rats.
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CHAPTER 57 Pesticides Affecting Oxidative Phosphorylation
Developmental ToxicitylTeratogenicity According to WHO/ IPCSIILO (1993) "animal tests show that this substance possibly causes malformations in human babies." Both the phenol and the octanoate carry the European Community's Hazard Rating R63 "possible risk of harm to the unborn child" (Tomlin, 2000). No details of the studies underlying these designations were found. Studies in experimental animals were positive for teratogenicity. The NOEL was set at 5 mg/kg/day for rats and 15 mg/kg/day for rabbits (Ahrens, 1994). Hormonal Effects Because of the occurrence of thyroid hyperplasia in animal studies with ioxynil, and its structural similarities to 3,5,3'-triiodo-L-thyronine, Ogilvie and Ramsden (1988) examined its binding in vitro to human plasma proteins in order to evaluate its possible effects on the synthesis, transport, and metabolism of thyroid hormones. Ioxynil bound significantly to both hormone binding sites on the thyroxine binding prealbumin but not to the thyroxin binding globulin or albumin. It was concluded that ioxynil is unlikely to interfere with hormone transport by the globulin in vivo, and it is not a general mimic of thyroxine since many compounds with a variety of structures bind to the prealbumin. However, it could displace triiodothyronine from its binding to the prealbumin and accumulate in the cerebrospinal fluid where the prealbumin is concentrated. These results did not clearly explain the effects of ioxynil on the thyroid in vivo. Human Toxicology There are a number of examples of human poisoning by ioxynil either through exposure during its manufacture or use, or through suicide attempts. In a few cases, ingestion of ioxynil alone has been reported, but more commonly exposure is in conjunction with other herbicides, particularly chlorophenoxy compounds. As described below, the ingestion of 2 to 3 g of ioxynil by an adult can lead to death in less than an hour. The ingestion of 18 g by an adult led to death in 15 hr even with hospital care. Blood concentrations in fatal cases have ranged from 0.04 to 0.67 g/l (Flanagan et aI., 1990). After acute exposure, ioxynil, generally in conjunction with chlorophenoxy herbicides, may cause dizziness, hyperthermia, sweating, gastrointestinal disturbances and vomiting, tachycardia and possible cardiac arrest, pulmonary edema, weakness, increased respiration, acidosis and renal failure arising from dehydration, weight loss, myalgias of the legs, euphoria, agitation, syncope, coma, and death. No antidotes are known. Treatment after decontamination is supportive and symptomatic. Activated charcoal (25 to 100 g in adults) given as a slurry may be helpful. As with other uncouplers, salicylates are contraindicated. Further discussion of treatment and monitoring options is given in HSDB (2000). Smysl et al. (1977) report a fatal example of ioxynil intoxication in which a 54-year-old man who had previously undergone gastrectomy mistakenly swallowed an 11.3% solution of ioxynil. He died 45 minutes later. Autopsy revealed hyperemia of all organs, and edema of the lungs and brain.
The estimated amount of ioxynil consumed was 2-3 g (i.e., approximately 30-40 mg/kg). Ioxynil was present in the serum at 13.2 ppm. Alcohol was also present at a concentration of 0.135%. A report of nonfatal into xi cations in workers manufacturing both bromoxynil and ioxynil (Conso et aI., 1977) has already been described (Section 57.7.2.2). Dickey et al. (1988) report a case of poisoning in a 37 -yearold woman who ingested 190 ml of a mixture of two herbicides, ioxynil (35 g/l) and MCPP (105 g/l), a chlorophenoxy plant hormone mimic. Symptoms included metabolic acidosis, tachycardia, pupiUary constriction, and pyrexia. Seventeen hr after ingesting the herbicide, the patient developed fever, muscle rigidity, and cardiac asystole, and died I hour later. The authors considered it unusual that this patient suffered no loss of consciousness prior to death since coma is commonly reported as an early event in such poisonings. Moderate pulmonary and cerebral edema and early necrosis of the liver and renal tubules were recorded at autopsy. Most of the signs of poisoning appear to be typical of those caused by oxphos uncouplers which is reasonable since the blood level of MCPP 10 hr before death was 515 mg/l whereas that of ioxynil was 317 mg/l, and ioxynil has an acute toxicity to mammals that is several times higher than that of MCPP. In addition, chlorophenoxy herbicides themselves have weak uncoupling activity. The reported amount of ioxynil ingested (6.7 g) would represent a dose of about 50 mg/kg for a person weighing 130-140 pounds. In a study of occupational poisoning by a three-component herbicide mixture containing ioxynil, isoproturon (a urea herbicide), and MCPP, Gibaud (1983) concluded that only ioxynil was responsible for the toxic actions of the product, and the effects resembled those of other well-established phenolic uncouplers. Flanagan et al. (1990) review 11 similar cases of poisoning by ioxynil. Ten also involved co-exposure to chlorophenoxy herbicides. Seven of the 11 patients died. Signs and symptoms were typical of uncoupling, but chlorophenoxy herbicides also cause many of these effects. Of the 3 patients who died after admission to hospital, in each case fatal cardiac arrest occurred 5 to 8 hr after admission. Alkaline diuresis was found to enhance clearance of the chlorophenoxy compounds but was ineffective in speeding the elimination of ioxynil. Biochemical Mechanism of Action Ioxynil behaves as a typical phenolic uncoupler in studies with rat liver mitochondria. Respiration is increased at lower concentrations and inhibition occurs at higher concentrations. It is a reasonably potent uncoupler (UCso of I J.LM) making it about threefold more active than bromoxynil and 30-fold more active than 2,4-DNP (Parker, 1965). It also acted as a typical protonophoric uncoupler in mung bean mitochondria with activity at 4 J.LM compared to 0.5 J.LM for dinoseb (Moreland and Novitzky, 1988). Absorption, Metabolism, and Elimination loxynil is absorbed through the skin and eyes, causing local pain and redness.
57.7 Mitochondrial Uncouplers Environmental Fate and Toxicity The acute tOXICIty of ioxynil to birds is variable (Table 57.5). As with bromoxynil, pheasants are very susceptible whereas mallards appear to be remarkably resistant (Ahrens, 1994). The value for hens lies in between with an acute oral LDso of 200 mg/kg (Carpenter et aI., 1964). The octanoate has quite low toxicity to birds, even for pheasants which are very sensitive to the free phenol (Table 57.5). Both the free phenol and the octanoate have relatively low toxicities to aquatic species, although data are limited (Table 57.5). This stands in notable contrast to the often very high toxicity of its relative, bromoxynil, to aquatic forms. The estimated BCF value of 3 for ioxynil indicates a very low probability of bioconcentration from water by aquatic organisms. The estimated Koc of75 suggests ioxynil binds only weakly to soil colloids and will have a high mobility in the soil. It degrades rapidly in aerobic soil, primarily by microbial action, with a half-life of 9-10 days. Initial steps in degradation involve the hydrolysis of the cyano group to yield the benzamide and benzoic acid (Hsu and Camper, 1975). Deiodination also occurs, probably by halogen exchange with bromine and chlorine (Grass et aI., 2000). The half-life for atmospheric photodegradation, primarily through reaction with photochemically derived hydroxyl radicals, is about 74 days. The degradation of ioxynil in natural water samples in the dark, approximating the conditions of ground water, is slow with 10% degradation requiring over 1 month (Nolte et aI., 1995). However, photodegradation increases the rate of loss markedly. Millet et al. (1998) report calculated half-lives for photolysis in water ranging from 0.10 day in summer to 3 days in winter at a 50° latitude. Ioxynil is occasionally detected in surface water runoff sources in areas of use in Europe (HSDB, 2000) but only at extremely low concentrations (0.01 to 0.1 ppb). Ioxynil has a relatively low toxicity to aquatic species (Table 57.5) which presents a contrast to some results with bromoxynil in fish. Carpenter et al. (1964) note that the toxicity of ioxynil to fish, like that of bromoxynil, is very dependent on the hardness of the water, presumably because of decreased solubility of the calcium salt. In soft water the LCso to harlequin fish was 3300 ppb. In hard water the LCso increased to 74,000 ppb. Niclosamide: General Properties and Uses 2',5-Dichloro4'-nitrosalicylanilide (Fig. 57.14) has CAS Reg. No. 50-65-7. Niclosamide is utilized either as the free acid or as its salt with 2-aminoethanol (CAS Reg. No. 1420-04-8). The common name of the salt is clonitralid or niclosamide-olamine. It was discovered by Bayer AG and is described by Gonnert and Schraufstatter (1958). The development and properties of niclosamide are reviewed by Gonnert (1962) and Schraufstatter (1962). The chemistry, uses, and toxicology of niclosamide are described in detail by Andrews et al. (1983) and by Knowles (1991) in the previous edition of this handbook. There have been no important changes in our understanding of niclosamide's toxicology in the last decade, so coverage here provides a briefer outline and focuses predominantly on more recent or additional data. Niclosamide is just one of several currently available
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salicylanilides that are used clinically as anthelmintics and fasciolicides including closantel, oxyclozamide, rafoxamide, and resortantel. The chemistry, mode of action, pharmacokinetics, and toxicity of these compounds have recently been reviewed by Swan (1999). Niclosamide exists as yellowish grey crystals, m.p. 230°C, v.p. < 1 x 10-3 Pa (20°C), w.s. 1.6 ppm (pH 6.4, 20°C)/11 Oppm (pH 9.1, 20°C), log P4.7 (1.0 at pH 9.6). Clonitralid exists as yellow crystals, m.p. 216°C, v.p. 1.3 x 10-6 Pa (25°C), w.s. 230 ppm (20°C). Niclosamide is hydrolyzed only slowly at physiological pH (half-life about 7 days at pH 6.9) but it is susceptible to photolysis (e.g., a 60-80% loss was observed after 16 hr exposure of an aqueous solution to sunlight), but results seem to be very variable and formulated material may be much more stable (Andrewsetal.,1983). Niclosamide (BAY 25648) and its salt are used as a molluscicides in the control of snails that vector schistosomiasis, in human and veterinary practice as nematicides and as anthelmintics/cestocides in the treatment of internal parasites, particularly tapeworms, and as lampreycides (often in combination with 3-trifluoromethyl-4-nitrophenol). Trade names include Bayluscide, Cestocid, Fenasal, Mansonil, Niclocide, Phenesal, and Yomesan. Toxicology Profile The primary source of data is U.S. EPA (1999d). Older toxicology results have been extensively reviewed by Andrews et al. (1983) and Knowles (1991) and much of that work is not included here. The data presented below are in complete agreement with these previous studies regarding the unremarkable toxicology of niclosamide in experimental animals even at high doses. The level of genotoxicity of niclosamide seems to be a matter of some dispute, but since chronic exposure studies to assess its carcinogenicity have been conducted in both rats and mice with negative results, the issue does not seem to be a crucial one. One data gap is that no multigenerational study for reproductive toxicity was found. Acute Toxicity The acute toxicity of niclosamide to a variety of mammalian species is low by all normal routes of exposure. This is true both for free niclosamide (Table 57.5) and its ethanolamine salt (Andrews et aI., 1983). It is much more toxic by intravenous administration with an LDso in rats of 7.5 mg/kg (Vega et aI., 1988), which demonstrates the high intrinsic toxicity of the molecule. Signs of acute poisoning include vomiting, hypopnea, convulsions, and sedation (Andrews et aI., 1983). IrritationlSensitization Niclosamide is a strong eye irritant, but not a skin irritant. It is a moderate dermal sensitizer. Subchronic Toxicity In a 90-day dietary study in rats, no treatment-related effects were seen at doses up to 500 mg/kgl day. In a similar study in dogs with doses up to 25 mg/kg/day, no notable specific adverse effects were observed either. In
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CHAPTER 57 Pesticides Affecting Oxidative Phosphorylation
hamsters the only response noted with dietary concentrations up to 726 mg/kg/day was reduced weight gain. Chronic Toxicity A series of older chronic dietary studies in rats and dogs is summarized by Andrews et al. (1983). Rats appear to tolerate at least 1000 mg niclosamide/kg/day and dogs tolerate 100 mg/kg/day without evidence of adverse effects. Carcinogenicity In 1978 the National Cancer Institute conducted an l8-month dietary study in Osborne-Mendel rats and B6C3Fl mice using clonitralid at doses up to 1421 mg/kg/day (rat) or 78 mg/kg/day (mouse). This study was flawed because of inadequate survival among the male mice, but there was no evidence of carcinogenicity in female mice or in rats of either sex (U.S. EPA, 1999d). No evidence of carcinogenicity was seen in the other chronic studies reviewed by Andrews et al. (1983). MutagenicitylGenotoxicity Niclosamide was negative for forward mutations in a mouse lymphoma cell assay in vitro and for chromosomal effects in mouse bone marrow cells in vivo (U.S. EPA, 1999d). In contrast, in a similar study of the induction of chromosome aberrations and sister chromatid exchange in mouse bone marrow cells in vivo, Giri et al. (1996) reported that niclosamide gave positive responses. After an intraperitoneal dose of 25 mg/kg, both indices of genotoxicity were significantly increased compared to controls. A similar oral dose had no effects on sister chromatid exchange but did increase chromosomal aberrations. Niclosamide also gave some positive results in a study on the induction of chromosomal aberrations in human peripherallymphocytes both in vivo and in vitro (Ostrosky-Wegman et aI., 1986). Results differed considerably between individuals. Three of five patients showed an increase in chromosomal aberrations after niclosamide treatment but no evidence for sister chromatid exchange was observed. The in vitro results with human lymphocytes from several individuals also revealed substantial variability in responses for both clastogenicity and sister chromatid exchange. A metabolic activation (hepatic S9) system was necessary for most positive results in vitro. The results of additional studies on the genotoxicity of niclosamide are outlined by Espinosa-Aguirre et al. (1991). When clonitralid was tested in a dominant lethal assays in mice at levels that caused minor symptoms of intoxication, no evidence of mutagenicity was seen in the progeny (Andrews et al., 1983). A negative result was also obtained in a modified Ames assay (MacPhee and Podger, 1977) with or without activation, but a formulated sample of niclosamide was used in this study. Subsequent studies using versions of the Ames Salmonella mutation assay and technical niclosamide gave negative results when conducted without the inclusion of a metabolic activation (liver S9) system, but when the activation system was included, a weak positive result was obtained (Andrews et aI., 1983). The positive mutagenic outcome with activation was confirmed in a series of studies by Espinosa-Aguirre et al. (1989, 1991) and Cortinas de Nava et al. (1983). The role of metabolic activation in these positive resonses was studied
further by Espinosa-Aguirre et al. (1991). The hydrolysis products of niclosamide were either inactive (5-chlorosalicylic acid) or only weakly active (2-chloro-4-nitroaniline) in the Ames assay. Based on the use of Salmonella mutants with enhanced or diminished nitroreductase and N -acetylase activity, reduction of the nitro group and subsequent acetylation was considered to be the key reaction sequence leading to the formation of mutagens. Nitroreduction is an important metabolic pathway for niclosamide in mammals including humans (Andrews et aI., 1983; Knowles, 1991), and its N-acetylamino metabolite has been detected in human urine after oral dosing with niclosamide (Duhm et aI., 1961), which elevates the significance of its probable mutagenic activity. A direct test of the potency of this metabolite in mutagenesis assays would be valuable. Mice treated orally with five daily doses of niclosamide also produced Ames-active mutagens in the urine and the mutagenic activity was somewhat increased by the addition of tJ-glucuronidase (Cortinas de Nava et aI., 1983; Vega et aI., 1988). The amount of mutagen production was dose-dependant over the range of doses utilized from 30 to 100 mg/kg. Sperm malformations were also seen at doses of 60 mg/kg and higher but the relationship of this effect to the appearance of urinary mutagens is speCUlative. The evidence suggesting that niclosamide is converted to one or more genotoxic products by vertebrate tisssues is therefore strong. However, as already mentioned, there appears to be no evidence that niclosamide is carcinogenic in rats and mice in vivo. Reproductive Toxicity No information on reproductive toxicity seems to be available from typical regulatory guideline studies. Vega et al. (1988) observed an increase in abnormal sperm morphology in two strains of mice after five daily oral doses of niclosamide at 30-120 mg/kg. No effects on testis weight or sperm count were seen and the implications for reproductive success were not assessed. Developmental ToxicitylTeratogenicity A study of pregnant rabbits at oral doses of niclosamide up to 180 mg/kg/day revealed no maternal toxicity. Peritoneal hemorrhage was an effect seen in some fetuses but this was regarded as only equivocally related to the treatment. This study was considered to have several significant flaws by U.S. EPA (1999d). Awad (1995) conducted a developmental study with niclosamide in Wistar rats at a single dose of 80 mg/kg/day. An increase in the incidence of skeletal malformations in fetuses and evidence of fetotoxicity were claimed, but the results are marginal and it appears that the increase in malformations was not statistically significant. The dose administered was only 2% of the LD50 (4000 mg/kg) in these rats. Previous studies in rats and rabbits revealed no developmental toxicity (Andrews et al., 1983). Overall it appears very unlikely that niclosamide is a developmental toxicant in mammals. Human Toxicology Niclosamide is used as an oral medication for intestinal parasites, particularly tapeworms, in both
57.7 Mitochondrial Uncouplers
human and veterinary practice. A dose of 2 g per peron is typically given. Transient gastrointestinal upsets, abdominal discomfort, anorexia, diarrhea, drowsiness, and dizziness have been observed occasionally as adverse effects, but there appear to be no recorded incidents of serious poisoning, long lasting side-effects, or death due to the use of niclosamide either in its clinical or pesticidal uses. Severe dermatitis has been reported in a few cases in using a 25% EC formulation of niclosamide, but this was probably due to other chemicals in the formulation. Further information on human responses to niclosamide is provided by Andrews et al. (1983), Knowles (1991), and HSDB (2000).
Biochemical Mechanism of Toxicity Niclosamide and closely related salicylanilides are very strong uncouplers (Gonnert et al., 1963; Ilivicky and Casida, 1969; Williamson and Metcalf, 1967) with the most potent compounds showing activity on isolated mammalian and insect mitochondria at concentrations as low as 1 nM. Niclosamide is active as an uncoupler at 10-100 nM which almost WOO-fold more active than 2,4-DNP. Typically of uncouplers, niclosamide stimulates mitochondrial respiration at lower concentrations but inhibits respiration as the concentration is increased (Gonnert and Schraufstatter, 1958). Niclosamide stimulated succinate-based oxygen consumption in snail tissues with a maximal effect at the remarkably low concentration of 0.1 nM (Ishak et aI., 1970). At higher concentrations, respiration was again inhibited. Many other studies with tissues and mitochondria from target and nontarget species fully support the concept that uncoupling occurs at low concentrations followed by inhibition at higher ones and that effects on oxphos are likely to be a major factor underlying the toxicity of niclosamide (Andrews et aI., 1983). Absorption, Metabolism, and Excretion After an oral dose of niclosamide in rats, two-thirds of the dose passed out in the feces and one-third was absorbed and then excreted in the urine after nitroreduction. The amide bond does not appear to be hydrolyzed to any great extent. Studies in human volunteers also showed limited uptake of niclosamide from the gastrointestinal tract after oral dosing. Only 2-25% was eliminated in the urine over 4 days. Considerable variations in pharmacokinetics between individual were noted. It was concluded that uptake is limited and metabolism of the absorbed dose is rapid, primarily occurring by nitroreduction followed by N -acetylation (Andrews et aI., 1983; Duhm et aI., 1961). Niclosamide is poorly absorbed after dermal application also. A study in rats and minipigs found that only 2% and 16% respectively of a dermal application had been absorbed in 1 week (Brennan et aI., 1991). Environmental Fate and Toxicity This topic is reviewed by Andrews et al. (1983) and in outline by Tomlin (2000). The toxicity to birds is very varied with one species (unstated) having an acute oral LDso as low as 60 mg/kg while for others it is over 2000 mg/kg (U.S. EPA, 1999d; Table 57.5). In a short-term avian dietary study (species unstated), the LCso was
1239
over 5419 ppm, indicating a very low level of toxicity. The acute toxicity of niclosamide to fish is high to very high (Table 57.5) which necessitates care in its use in aquatic systems. The sea lamprey is highly sensitive with an LCso of 49 ppb which assures niclosamide's efficacy as a lampreycide. The toxicity to aquatic invertebrates is very varied. Crayfish and some aquatic insects are relatively tolerant of niclosamide, Daphnia are quite sensitive, and snails are highly sensitive (LCso values of 49-63 ppb) in keeping with its molIuscicidal uses. In practical use for snail control, a concentration of 600-1000 ppb for 8 hr or 333 ppb for 24 hr is recommended. Consequently significant fish toxicity is likely to occur during its use in antischistosomiasis programs (Andrews et aI., 1983). The uptake and metabolism of niclosamide has been studied in rainbow trout by Lech and his colleagues (e.g., see Statham and Lech, 1975). Limited bioconcentration was observed and depuration was essentially complete in 48 to 72 hr. Residues were strongly concentrated in the bile. A glucuronyl conjugate of niclosamide was identified as the major biliary metabolite. Niclosamide tends to be more toxic as the pH and water hardness decrease. It is only moderately bound to soil organic matter and soil microbial decomposition is slow (Gonnert and Schraufstatter, 1958). However, rapid degradation with a halflife of 0.3 days was reported in rice paddies in association with its use as a molluscide (Calumpang et aI., 1995). The behavior of niclosamide in aquatic systems has been studied in detail by Monkiedje et al. (1995). They found that hydrolysis occurred at pH values above 7, but little photolysis was seen in water. Under field conditions its half-life in ponds was 3.4 days.
57.7.2.3 Amine Uncouplers When connected to an appropriate combination of strong electron-withdrawing and lipophilic groups (R, R'), secondary amines (R-NH-R') act as lipophilic weak acids and thus fit the definition of uncouplers. The structural and physicochemical requirements for potent diarylamine uncouplers have been determined by Guo et al. (1991a, b). They concluded that like phenolic uncouplers, the best diarylamine uncouplers are lipophilic weak acids that can locate within the mitochondrial membrane and act as protonophores, although some differences between these two groups were also noted in shielding effects by bulky groups around the amine moiety. A number of these amine uncouplers are used as pesticides including a rodenticide (bromethalin), a fungicide (fluazinam), and an insecticide (chlorfenapyr). Sulfluramid, a minor insecticide, also fits within the definition of this amine uncoupler group, but is considered separately for other reasons.
Bromethalin: General Properties and Uses a, a, a- Trifluoro-N-methyl-4,6-dinitro-N -(2,4,6-tribromophenyl)-o-toluidine (Fig. 57.15), CAS Reg. No. 63333-65-7, was discovered by Eli Lilly & Co. and is described by Dreikom et al. (1979). Its development and properties are reviewed by Dreikom and O'Doherty (1984). It exists as white crystals, m.p. 150-151°C, v.p. 1.3 x 1O-s Pa (25°C), w.s. <10 ppb, log P4.26. Bromethalin is not
1240
CHAPTER 57
Br
Pesticides Affecting Oxidative Phosphorylation
CN
CFsO--ry AC 303.268
r l8Jc
I
FI uaz i nam
Figure 57.15
Uncouplers of oxidative phosphorylation-Secondary amines.
subject to hydrolysis, but it is very sensitive to photodegradation. Bromethalin (EL-614) is a rodenticide used in baits. Trade names include Assault, Trounce, and Vengeance. Toxicology Profile (l998f).
The primary source of data is U.S. EPA
Acute Toxicity As a commercial rodenticide, bromethalin necessarily shows a high toxicity to rodents (Table 57.5; values from U.S. EPA, 1998f). It is also highly toxic to other vertebrates such as dogs and cats with acute oral LDso values of 4.8 and 18 mg/kg, respectively. Van Lier and Cherry (1988) provide lower acute oral LDso values (i.e., 2.0 mg/kg for the rat, 1.8 mg/kg for the cat, and 5.3 mg/kg for the mouse). LDso values for the monkey (species unstated) and rabbit are 5.0 and 13.0 mg/kg, respectively. A notable exception to the high acute toxicity of bromethalin is the guinea pig for which the acute oral LDso is over 1000 mg/kg (van Lier and Cherry, 1988). Bromethalin's acute toxicity by dermal application is low, but its inhalation toxicity is very high (Table 57.5). At lower toxic doses, signs of acute poisoning in rodents (Dorman et aI., 1992a; van Lier and Cherry, 1988), dogs (Dorman et aI., 1990a, 1990b), and cats (Dorman et aI., 1992b) include slowly developing ataxia, hypothermia, hind1imb paralysis, loss of responses to sensory stimulation, labored respiration, depression, coma, and death generally due to respiratory arrest and often with terminal clonic convulsions. Focal motor seizures also occurred in cats. At higher doses in rodents (van Lier and Cherry, 1988) and dogs (Dorman et aI., 1990a) the clinical signs develop more rapidly over 4 to 24 hr and involve a stronger component of neuroexcitation including hyperexcitation, tremors, and generalized seizures.
Pathological examinations in the above studies gave very similar results across different mammalian species. They revealed generalized spongy degeneration and edema of the white matter of the brain, spinal cord, and optic nerve. This is caused by the separation of the myelin lamellae at the interperiod lines and the collection of fluid between the lamellae (edema) with progression to disruption of the lamellae and coalescence of the vacuoles into larger watery areas (spongy changes). Swelling of astrocytes and oligodendroglial cells also occurs. No inflammatory or macrophage response is seen, axonal degeneration is lacking, and the neural pathology is slowly reversible in surviving animals. Lipid peroxidation can be observed in the brain (Dorman et aI., 1992b). Many of the clinical signs and lethal consequences from bromethalin toxicosis are likely to arise from pressure effects on neuronal conduction due to extensive brain and spinal column swelling. The brain water content in rats may increase by as much as 15% during severe acute poisoning. The cerebrospinal fluid pressure is also increased by almost 400% (van Lier and Cherry, 1988). The corticosteroid dexamethasone was effective in reducing this increase in cerebrospinal fluid pressure in rats as was the intravenous infusion of urea or 25% mannitol as hyperosmotic diuretic agents. The electroencephlographic changes in dogs dosed with bromethalin at 6.25 mg/kg orally have been described by Dorman et al. (1991). Abnormal changes included spike and spike-wave patterns, high voltage slow wave activity, photoconvulsive or photoparoxysmal irritative responses, and voltage depression in all leads. The administration of a commercial extract of Gingko biloba to rats by gavage antagonized lipid peroxidation and the development of edema in the brains of rats treated with a toxic oral dose of bromethalin (1 mg/kg) and reduced the severity of the signs of poisoning. This preparation was chosen because of evidence that it ameliorated similar neurological effects caused by triethyltin (Dorman et al., 1992a). Other possible treatments for poisoning were examined by Dorman et al. (1990c). The repeated administration of a charcoal-sorbitol mixture immediately after dosing was effective but, contrary to the experience of van Lier and Cherry (1988) in rats, the administration of furosemide, mannitol, and dexamethasone in various combinations upon the appearance of clinical signs in dogs did not decrease the severity of poisoning. A number of case of poisoning by bromethalin in nontarget species including domestic cats have been reported (Martin and Johnson, 1989).
IrritationlSensitization Bromethalin is an eye irritant, but it is not a skin irritant or sensitizer. Subchronic Toxicity The LOEL for bromethalin was 0.125 mg/kg/day in 90-day gavage studies in both rats and dogs. The most sensitive endpoint was spongy degeneration of the white matter of the brain, optic nerve, and spinal column. No other effects were observed in rats. Severe neurological signs and deaths occurred in dogs given 0.2 mg/kg/day subchronically.
57.7 Mitochondrial Uncouplers Chronic Toxicity No chronic exposure studies appear to have been performed with bromethalin since they are not required for the registration of this compound for nonfood use as a rodenticide (U.S. EPA, 1998f). MutagenicitylGenotoxicity Bromethalin was not mutagenic in a battery of tests. No tests for clastogenicity were reported. Carcinogenicity
See Chronic Toxicity is observed.
Reproductive Toxicity
See Chronic Toxicity is observed.
Developmental ToxicitylTeratogenicity No developmental toxicity was observed in standard studies in rats and rabbits at doses up to 0.5 mg/kg/day. Severe maternal toxicity and some deaths occurred at the highest dose. Neurotoxicity The critical effects ofbromethalin on the white matter of the central nervous system have been described. No delayed neurotoxicity was observed in hens given two high oral doses of bromethalin. Human Toxicology One possible example of human poisoning by bromethalin has been described by Buller et al. (1996). In this case a 28-year-old male was found unconscious with evidence of consumption of rat poison. Two open packets, one containing bromethalin (0.01 %) and the other diphacinone (an anticoagulant) were nearby. On tactile stimulation, tremors and strong myoclonic contractions were seen. Clinical laboratory results were normal. After decontamination and 2 days of supportive and symptomatic therapy, recovery was near complete and the patient was discharged. No estimate was provided of the amount of rodenticide consumed. Bromethalin does cause neurological signs in experimental animals, but its specific role in this incident is uncertain. The very low concentration of active ingredient does not seem to be congruent with severe poisoning. A person of 70 kg would need to consume 700 g of a 0.01 % bait to achieve a dose of 1 mg/kg. Also, the signs of poisoning that are described are of limited comparability to those in experimental animals poisoned by bromethalin where spontaneous tremors and convulsions are seen primarily at high lethal doses. The nature of the treatment administered to resolve this poisoning was not described. Biochemical Mechanism of Action Bromethalin is a propesticide which must be converted by N -demethy1ation to the corresponding free diphenylamine to initiate toxicity (van Lier and Cherry, 1988; Fig. 57.15). Unlike bromethalin itself, the activation product is a strong uncoupler of brain and liver mitochondria from rats (van Lier and Cherry, 1988)) with similar UCso values of 600 and 870 nM, respectively. N -demethylbromethalin was even more active as an uncoupIer against mouse liver mitochondria. Based on comparative phamacokinetic studies in the rat, van Lier and Cherry concluded that bromethalin is converted rapidly and almost quantitatively to N-demethyl bromethalin. Inducers of microsomal
1241
cytochrome P-450 monooxygenases (3-methy1cholanthrene and Aroclor 1254) caused an increase in toxicity when given to mice 1 to 3 days prior to an oral challenge dose of bromethalin at the LDso. Phenobarbital, on the other hand, which induces a different suite of P-450 isoforms, antagonized its toxicity. A microsomal oxidation inhibitor, SKF-525A, given 1 hour before bromethalin antagonized its toxicity as expected if oxidative activation is occurring. On the other hand, SKF-525A increased the toxicity of N -demethy I bromethalin, suggesting that this is detoxified by microsomal oxygenase action. The delay in toxicity caused by this necessary metabolic activation probably contributes to the effective rodenticidal action of brometha1in since it tends to minimize bait-shyness. The characteristic swelling and vacuolization of the white matter of the brain and spinal cord has been attributed to the failure of the pumps responsible for ion regulation in the myelin sheath due to the uncoupling activity of the activation product. In turn, this causes ions to accumulate followed by osmotic entry of water, swelling, and splitting of the myelin layers to form vacuoles (Dorman et aI., 1992b; van Lier and Cherry, 1988). Both in its structure and in its toxicological effects, bromethalin closely resembles fentrifanil, an experimental diarylamine acaricide. This compound is a potent direct-acting uncoupler of oxphos (Nizamani and Hollingworth, 1980) and causes vacuolization of myelin and edema of the white matter of the central nervous system (Lock et aI., 1981) which appears to be identical to that caused by bromethalin. These and other chemically induced myelinopathies are reviewed in detail by van Gemert and Killeen (1998). Absorption, Metabolism, and Elimination After oral dosing in rats, bromoethalin was taken up rapidly with a half-life for absorption into the plasma of 2.7 hr. Clearance from the plasma after dosing ceased was slower with a half-life of about 5.6 days (van Lier and Cherry, 1988) N -demethylbromethalin was the major metabolite. Environmental Fate and Toxicity Birds are highly sensitive to the acute toxic effects of bromethalin (Table 57.5). In 8-day feeding studies with birds the LCsos were 210 ppm for bobwhite quail and 620 ppm for mallard ducks, indicating a moderate to high toxicity. Bromethalin is also extremely toxic to aquatic vertebrates and invertebrates (Table 57.5), but because of its use patterns and formulations it is unlikely to present a significant risk in practice. It is rather persistent with an average half-life in aerobic soils of 178 days. Chlorfenapyr: General Properties and Uses 4-Bromo-2(4-chlorophenyl)-1-ethoxymethy1-5 -trifluoromethylpyrrole- 3carbonitrile (Fig. 57.15), CAS Reg. No. 122453-73-0, was discovered by American Cyanamid and is described by Lovell et al. (1990). Its development and properties ar reviewed by Addor et al. (1992). It exists as light yellow crystals, m.p. WO-101°C, v.p. <1.3 x lO-s Pa (25°C), w.s. 0.14 ppm (25°C), log P4.83.
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CHAPTER 57
Pesticides Affecting Oxidative Phosphorylation
Chlorfenapyr (AC 303,630, MK-242) is a new broad spectrum insecticide and acaricide. It was developed by structural modification of a microbial natural product, dioxapyrrolomycin, as described by Addor et at. (1992). A general review of chlorfenapyr's properties, mechanism of action, and structureactivity relations is provided by Hunt and Treacy (1997). Although it is now widely used as an insecticide in many countries, concerns over its relatively long environmental persistence and potential for acute lethal and reproductive toxicity in birds and possible aquatic toxicity have made it a controversial compound in the United States (Williams, 1999) and so far havc limited its registration to nonagricultural uses. Trade names include Alert, Citrex, Intrepid, Phantom, Pirate, Stalker, and Sunfire. Toxicology Profile (1997d).
incidence of malignant histiocytic sarcoma and testicular interstitial cell tumors was discovered in males fed 30.8 mg chlorfenapyr/kg/day for 2 years. Some increase in liver adenomas and/or carcinomas was observed in females but the results were inconsistent. A significant increase in endometrial stromal polyps was also observed. As a result of the rat study, the carcinogenicity of chlorfenapyr was classified by the U.S. Environmental Protection Agency as "cannot be determined, suggestive." MutagenicitylGenotoxicity Chlorfenapyr is not mutagenic or clastogenic in a standard battery of assays including mutations in bacterial and mammalian cells, a micronucleus test in mice in vivo, and chromosomal aberration and DNA repair assays in mammalian cells in vitro.
The primary data source is U.S. EPA
Acute Toxicity The acute toxicity to rats is remarkably low for a compound that ultimately acts as a powerful uncoupler but the toxicity to mice and other vertebrates is considerably higher (Table 57.5). IrritationlSensitization Chlorfenapyr is a moderate but reversible eye irritant, but not a skin irritant. It does not cause dermal sensitization in guinea pigs. Subchronic Toxicity In a 90-day dietary toxicity study in rats, the LOEL was found to be 48 mg/kg/day based on reduced body weight and liver weight increases. At higher doses (97.5 mg/kg/day) males exhibited decreased activity, ataxia, anorexia, and chromodacryorrhea, and multiple alterations were seen in blood chemistry. In a similar study in mice, an LOEL for chlorfenapyr was found at 14.8 mg/kg/day in males and 40 mg/kg/day in females with an endpoint of hepatocellular hypertrophy. Spongiform encephalopathy was noted in the brain and the myelin of the spinal cord in both sexes at 63-76 mg/kg/day, a dose that was lethal to some animals. In dogs, no remarkable adverse effects other than lowered weight gains were seen in either a 90-day or a I-year dietary study with doses up to 10.1 mg/kg/day. A 28-day dermal study in rabbits showed first adverse efects at 400 mg/kg involving liver alterations (particularly vacuolization which increased in severity at 1000 mg/kg) and blood chemistry changes. Chronic Toxicity A I-year dietary study in dogs revealed only decreased body weigh at the highest dose (8.7 mg/kg/day). In an 18-month dietary study in mice, neuropathology in the form of extensive vacuolization of the white matter of the brain and spinal cord was again observed along with skin ulceration at 16.6 (M) to 21.9 (F) mg/kg/day. A 2-year study in rats revealed liver toxicity in the form of non-neoplastic hepatocellular enlargement at 15 (M) and 18.6 (F) mg/kg/day in the diet. Carcinogenicity No carcinogenic response was seen in the mouse chronic study, but in the study in rats, an increased
Reproductive Toxicity In a two-generation study in rats, the only reproductive adverse effect observed with chlorfenapyr was reduced weight gain in the pups which occurred at 22 mg/kg/day. This dose also caused lower weight gains in the parents. Developmental ToxicitylTeratogenicity No evidence for developmental toxicity was found in studies with pregnant rats at doses up to 225 mg/kg/day and rabbits at doses up to 30 mg/kg/day. Neurotoxicity The neurotoxic effects of chlorfenapyr have also given cause for concern. In an acute test, rats given 180 mg/kg showed no notable neurobehavioral response beyond lethargy and no neuropathology. In a I-year dietary study in rats, myelin sheath swelling was recorded after 13 weeks at 28-37 mg/kg/day. At 52 weeks, the myelinopathic process was more generalized and severe and was seen at a dose of 13.6 mg/kg/day. Only males were affected. The effect had disappeared after a recovery period of 16 weeks. As already described, similar effects were observed in dietary studies with mice.
Biochemical Mechanism of Action Chlorfenapyr shows virtually no uncoupling activity but removal of the N -ethoxymethyl group through microsomal oxidation (Fig. 57.15) releases the corresponding free pyrrole (AC 303,268) which is a lipophilic weak acid with very strong uncoupling activity (Black et aI., 1994). Inhibition of this oxidative activation reaction in insects in vivo with piperonyl butoxide antagonizes the toxicity of chlorfenapyr (Treacy et aI., 1994). AC 302,268 has an acute toxicity higher than that of chlorfenapyr in both vertebrates and insects (e.g., the acute oral LDso for the rat is 29 mg/kg). Upon injection into insects it cause an almost immediate and massive stimulation of respiration whereas the parent chlorfenapyr does so only after a significant delay (Black et aI., 1994). All these observations support the idea that chlorfenapyr is a proinsecticide and requires metabolic conversion to the active uncoupler, AC 303,268, before it exerts a toxic action. AC 303,268, and its further degradation products, is a
57.7 Mitochondrial Uncouplers
common metabolites of chlorfenapyr found in studies with vertebrates (below). Absorption, Metabolism, and Elimination An oral dose of chlorfenapyr was poorly absorbed in rats and over 80% was eliminated in the feces. The absorbed dose was rapidly cleared in the urine. Urinary metabolites included products of N -alkoxy side chain oxidation and removal (e.g., AC 303,268) and ring hydroxylation products and their conjugates. Similar results were obtained with goats and hens. In fish and soil, debromination is observed as an additional metabolic route (U.S. EPA, 1997d). Environmental Fate and Toxicity The acute tOXICIty of chlorfenapyr to birds is high to very high (Table 57.5). The subchronic toxicity to birds is also very high (e.g., in mallard ducks, chlorfenapyr at 2.5 ppm in the diet reduced adult body weights and decreased egg-laying and hatch rates by about 4060%). Risk assessments by the U.S. Environmental Protection Agency suggested a substantial reproductive risk to birds from some uses of chlorfenapyr and indicated the need for the development of mitigation strategies (U.S. EPA, 1997d). Although these results suggest that there is a considerable potential for the use of chlorfenapyr to cause acute and reproductive toxicity in wild birds, there appear to be no reports of detectable impacts on birds or other wildlife when these have been closely monitored in areas where chlorfenapyr has been used (e.g., see Williams, 1999). Chlorfenapyr is also very highly toxic to fish and aquatic invertebrates (Table 57.5) which raises additional questions regarding its potential environmental impact and risk mitigation strategies. Chlorfenapyr binds firmly to soil with a Koc value about 11,500 mUg. This and its low water solubility indicate a low leaching potential but also imply possible environmental movement by surface runoff attached to soil colloids. Chlorfenapyr has a rather long half-life of 1.4 years in aerobic soils which suggests a capacity for some environmental accumulation with repeated use, possibly to toxic levels (U.S. EPA, 1997d). Fluazinam: General Properties and Uses 3-Chloro- N(3-chloro-5-trifluoromethyl-2-pyridyl)-a, a, a-trifluoro-2,6-dinitro-p-toluidine (Fig. 57.15), CAS Reg. No. 79622-59-6, was discovered by Ishihara Sangyo Kaisha, Ltd. and is described by Anema et al. (1992). It exists as yellow crystals, m.p. 115-1 17°C, v.p. 1.5 x 10-3 Pa (25°C), w.s. 0.157 ppm (pH 7, 25°C), log P3.56. Fluazinam (ASC-66825, IKF-1216) is a broad spectrum fungicide with additional acaricidal properties which is used on fruits, vegetables, and turfgrass. Trade names include Altima, Frowncide, Legacy, Shirlan, and Shogun. Toxicology Profile (2000b).
The general source of data is U.S. EPA
1243
Acute Toxicity Fluazinam has a very low acute toxicity to terrestrial vertebrates by the oral and dermal routes but is somewhat more toxic by inhalation (Table 57.5). IrritationlSensitization Fluazinam is irritating to the skin and very irritating to the eye. It is a weak dermal sensitizer. Subchronic Toxicity At the LOEL (41 mg/kg/day) in a 13week study, male rats showed hepatocellular hypertrophy and chronic sinusodial inflammation. In a subchronic dietary study in dogs, the LOEL was 100 mg/kg/day based on ocular changes and bile duct hyperplasia sometimes with cholangiofibrosis. Chronic Toxicity In a chronic feeding study in dogs the LOEL was 10 mg/kg/ day based on nonspecific toxicity. No ocular effects were seen. At the highest doses (100 and 1000 ppm) in a chronic feeding study in mice, macroscopic and microscopic lesions were produced in a number of tissues, particularly the liver and testes. In a chronic dietary study in rats, a LOEL was obtained at 100 ppm (about 4.3 mg/kg/day) based on a number of effects including macroscopic and microscopic lesions in multiple organs. A second study in rats also revealed lesions in the liver and testes as notable toxic endpoints. In a parallel study in mice, using dietary concentrations from 1000 to 7000 ppm, it was noted that an impurity in the technical material caused vacuolization in the brain and spinal cord, an effect seen with some other potent uncouplers (Section 57.7.2.3). Carcinogenicity No evidence of carcinogenicity was found in a 2-year rat dietary study. In mice, some increases in hepatocellular adenomas were seen at 3000 ppm, but this was not dose-dependent and no carcinomas were found. It was concluded that fluazinam is not a carcinogen in mice (U.S. EPA, 2000b). MutagenicitylGenotoxicity No genotoxic effects were observed in a battery of tests including two mutation tests in bacteria, a chromosome aberrations test in mammalian cells, a mouse micronucleus test, and a DNA repair test in bacteria. Reproductive Toxicity In a two-generation reproduction study in rats, the dietary NOEL for reproductive effects was 100 ppm (10.1 mg/kg/day) which was well above the NOEL for parental toxicity (20 ppm). Developmental ToxicitylTeratogenicity In both rats and rabbits, developmental effects were seen only at doses higher than those that were maternally toxic. Neurotoxicity No neurotoxic effects were seen in rats with an acute oral doses up to 2000 mg/kg, or in a subchronic study with doses up to 233 mg/kg/day. Hormonal Actions No evidence for adverse effects related to the disruption of estrogen actions was seen in any mammalian or avian study.
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CHAPTER 57 Pesticides Affecting Oxidative Phosphorylation
Human Toxicology There has been a history of skin irritation and sensitization associated with the repeated application of fluazinam in agricultural uses (Adams, 1997; Bruynzeel et al., 1995; van Ginkel and Sabapathy, 1995). Biochemical Mechanism of Action Fluazinam is a lipophilic weak acid with strong uncoupling activity on mitochondria in vitro (Brandt et al., 1992; Guo et al., 1991c; Hollingworth and Gadelhak, 1998). A detailed analysis of the relationship between physicochemical characteristics, structure, and uncoupling activity for fluazinam and its analogs is provided by Guo et al. (1991a, b). Fluazinam bears a structural resemblance to the active metabolite of bromethalin, but it differs greatly from this compound in its very low acute toxicity, primarily because it reacts readily with cellular thiols such as glutathione through the chlorine substituent in the phenyl ring (Clarke et al., 1998; Guo et al., 1991a; Hollingworth and Gadelhak, 1998). Analyses of the quantitative structure-activity relationships for the toxicity of fluazinam analogs to several fungi have led to the conclusion that, in addition to uncoupling effects, reactivity with sulfhydryl groups may contribute to the toxicity of fluazinam in some cases (Akagi et al., 1995, 1996b). Sulfhydryl (or other nucleophilic) reactions are probably also involved in the skin irritation and sensitization seen with fluazinam which resembles that caused by l-chloro-2,4-dinitrobenzene, a wellknown sulfhydryl-reactive compound. Absorption, Metabolism, and Excretion After an oral dose in rats, rapid and complete elimination occurred, primarily in the feces. Metabolites included products of nitro reduction, and glutathione and glucuronide conjugation (U.S. EPA, 2000b). Environmental Fate and Toxicity Fluazinam has a low acute toxicity to birds, but it is much more toxic to fish and aquatic invertebrates (Table 57.5). It binds strongly to soil with Kd values from 143 to 820. In the soil, it has a half-life of 33-62 days (Tomlin,2000). 57.7.2.4 Perfluorooctanesulfonic Acid and Derivatives The insecticidal properties of this chemical class were discovered serendipitously. Perfluororalkanecarboxylic acid and perfluoroalkanesulfonic acid derivatives have numerous industrial uses as surfactants (e.g., in conditioning paper products, fabric stain-proofing, and antifoaming agents). While investigating slow-acting insecticides for the control of fire ants, Vander Meer et al. (1985, 1986) included perfluoroctanesulfonamide surfactants as formulation aids. Control formulations made in this way, and lacking an apparent active ingredient, also showed excellent delayed toxicity to the ants. Subsequent studies of a variety of related fluorinated alkylsulfonamides showed that optimum insecticidal activity is found with an eight carbon chain leading to the development of sulfluramid as a slowacting insecticide. The free perfluorooctanesulfonic acid was also found to have good insecticidal properties (Vander Meer
Su I flu rami d
N-Deethyl sulfluramid (NDES)
Lithium perfluorooctanesulfonate (LPOS) Figure 57.16
Pefluorooctanesulfonyl pesticides.
et al., 1985) and the lithium salt of this acid was recently registered for limited use in the United States (U.S. EPA, 199ge).
Sulfluramid: General Properties and Uses N-ethylperfluorooctane-l-sulfonamide(Fig. 57.16), CAS Reg. No. 4151-50-2, was discovered by Vander Meer et al. (1985), developed by Griffin Chemical Co., and described and reviewed by Vander Meer et al. (1987). Structure-activity relationships for this group of compounds as insecticides are presented by Vander Meer et al. (1985, 1987). It exists as colorless crystals, m.p. 96°C, v.p. 5.7 x 10-5 Pa (25°C), practically insoluble in water, log P3.1, >6.8 (unionized form), pKa 9.5 (methanol). The technical material is a mixture of isomers, with approximately 70-90% linear and 1030% branched perfluorooctane chains. Sulfluramid (GX 071) is primarily a stomach poison and is used in bait formulations to control cockroaches, ants, termites, and wasps under the trade names of Alstar, Finitron, Mirex-S, Transflur, and Volcano. Sulfluramid-impregnated cardboard is used for termite control under the name FirstLine. Toxicology Profile (1989d).
The primary source of data is U.S. EPA
Acute Toxicity Sulfluramid has a low acute toxicity to vertebrates by all routes of exposure (Table 57.5), although there is some confusion regarding the exact LD50. Values in the literature range from 500 to > 5000 mg/kg. Signs of poisoning in rats (Vander Meer et aI., 1986) include decreased food consumption, weight loss, and emaciation. Transient lethargy, depression, and dyspnea were observed in sheep given sulfluramid orally at 15 mg/kg (Vitayavirasuk and Bowen, 1999). IrritationlSensitization Sulfluramid is not an eye irritant but it does cause mild skin irritation.
57.7 Mitochondrial Uncouplers
Subchronic Toxicity In dogs, pale mucous membranes, tachypnea, weight loss, diarrhea, and lethargy were observed after high doses for 2 weeks (Schnellmann, 1990). In 90-day dietary exposure studies, NOELs ranged from 10 ppm for rats to 100 ppm for dogs (F). Chronic Toxicity No results of chronic toxicity testing, including carcinogenicity studies, were found, probably because sulfiuramid is not registered for food uses. Carcinogenicity
See Chronic Toxicity.
MutagenicitylGenotoxicity Sulfiuramid gave negative results in limited genotoxicity testing (Ames mutagenicity and Chinese hamster ovary sister chromatid exchange assays). Reproductive Toxicity As a result of the observation that the oral administration of sulfiuramid caused transient sterility in dogs, Stump et al. (1997) studied the effect of oral doses (0.3 to 3.0 mg/kg/day) in pregnant New Zealand White rabbits. Neonatal mortality was seen at each dose, but no treatment-related effects were seen on sexual maturation or testicular function in the offspring and no histopathological findings in the reproductive organs were noted. Developmental ToxicitylTeratogenicity found in this area.
No
results
1245
drial uncoupling has been confirmed by Starkov et al. (2001) and Gadelhak and Hollingworth (unpublished). Studies with cockroaches show that both sulfiuramid and NDES cause a strong increase in respiratory rates in vivo, but sulfiuramid does so only after a 30-60 min delay, whereas NDES has an almost immediate effect (Hollingworth and Gadelhak, 1998). The respiratory rate increase caused by sulfiuramid, but not by NDES, is inhibited by pretreatment of the insects with the microsomal monooxygenase inhibitor piperonyl butoxide. This fully supports the idea that sulfiuramid is a propesticde that must be converted to NDES which then acts as an uncoupler. In addition, perfiuorooctanesulfonic acid and its derivatives are extremely powerful surfactum (e.g., see Kissa, 2001; Shinoda et aI., 1972) and the ability to disrupt membrane integrity must also be considered as a possible contributory mechanism to their effects on mitochondria. Human Toxicology In a study of workers in a pesticide packaging plant in Brazil, Machado-Neto et al. (1999) measured dermal exposure to sulfiuramid. It was concluded that this level of exposure was well within safety limits based on a NOEL (rat, 90-day, oral) of 10 mg/kg/day and a lO-fold safety factor. Even with a lOO-fold safety factor, these exposure would have been regarded as acceptable.
were
Biochemical Mechanism of Action Sulfiuramid is readily metabolized to its N -deethylated analog, perfiuorooctanesulfonamide (NDES). Studies in vitro with kidney proximal tubules (Schnellmann, 1990) showed that sulfiuramid caused an increase in respiratory rate over the range of 10-100 I-lM and cellular necrosis at higher concentrations. NDES was severalfold more potent than sulfiuramid in these actions. Further investigation with renal mitochondria indicated that NDES is a moderately active mitochondrial uncoupler causing a 50% increase in state 4 respiration at 5 I-lM through its action as an ionophore (Schnellmann and Manning, 1990). Although the parent sulfiuramid also stimulated state 4 respiration in mitochondria, the authors concluded that this was because of its conversion to its N -deethyl metabo1ite by the mitochondria and that it did not itself act as a protonophore. An uncoupling activity is possible since in these compounds the sulfonamide group is weakly acidic. Sulfiuramid has a pKa of 9.5 in methanol whereas NDES has a pKa of 9 these values should be about 2 units lower in water. They have high log P values. In addition, a special mechanism of uncoupling may be occurring with the desethyl analog. It has been shown that fatty acids can act as protonophoric uncouplers since the anionic form can be transported back across the mitochondrial inner membrane by a specific transporter system (Wallace and Starkov, 2000). As a fatty acid analog, NDES may bind to this transporter, thus facilitating its transfer across the membrane and promoting its uncoupling potency. The specific protonophoric action of sulfiuramid, and, more actively, NDES, leading to mitochon-
Absorption, Metabolism, and Elimination Sulfiuramid is converted to its N-deethyl analog in vitro (Schnellmann and Manning, 1990) and in vivo in rats and dogs (Arrendale et aI., 1989; Manning et aI., 1991) and sheep (Vitayavirasuk and Bowen, 1999). Manning et al. (1991) found that sulfiuramid is absorbed slowly and incompletely from the gastrointestinal tract in rats but the absorbed dose is rapidly N -deethylated. The half-life for clearing sulfiuramid from the blood was16 to 20 hr, depending on the vehicle used for administration. NDES had a much longer half-life for clearance of 103 to 107 hr. Grossman et al. (1992) also studied the distribution and elimination of sulfiuramid in rats dosed for 56 days in their diet at 75 mg/kg. No mortality was observed. NDES was present at all times in the blood and tissues, but sulfiuramid itself could not be detected. The clearance half-life from the blood for NDES at the end of the exposure was 10.8 days. The rate and extent of any further conversion of either sulfiuramid or NDES to perfiuorooctanesulfonic acid were not assessed in these studies and are unclear, but the slow clearance of NDES suggests that such a conversion cannot be rapid if it occurs at all. The free acid also has insecticidal properties and significant mammalian toxicity (Section 57.7.2.4). Environmental Fate and Toxicity Limited data suggest that sulfiuramid has moderate oral toxicity to birds (Table 57.5). The data from 8-day dietary studies in birds tend to confirm this with LCso values of 460 ppm for bobwhite quail and 165 ppm for mallard ducks. Its toxicity to fish is very low with LCso values for several species greater than 8000 to 10,000 ppm, but it is considerably more toxic to Daphnia (Table 57.5).
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Pesticides Affecting Oxidative Phosphorylation
The fate of sulfluramid in the environment is unclear, but, despite its high chemical stability, eventually it may degrade to perfluorooctanesulfonic acid. The environmental properties of this compound are outlined in below. Lithium Perfluorooctanesulfonate (LPOS): General Properties and Uses The structure of LPOS is shown in Fig. 57.16. It has CAS Reg. No. 29457-72-5. It was discovered by Vander Meer et al. (1985) and developed by S. C. 10hnson & Sons. It readily dissociates in water to yield perfluorooctanesulfonic acid (PFOS). Since the toxicological properties of the two compounds should be very similar they are treated jointly in this discussion. Perfluorooctanesulfonic acid is an extremely strong acid and under virtually all biological and environmental conditions it will exist solely in the form of the dissociated sulfonate anion. For simplicity, both the acid and its anion are here referred to as PFOS. LPOS is an off-white powder, decomp. 308°. PFOS has a v.p. of 3.3 x 10-4 Pa (20°C). Its W.s. varies from 519 ppm in pure water to 5 ppm in unfilltered sea water, HLC 7.2 x 10-9 atm m3 mole -1 in pure water. LPOS is registered in the United States for outdoor residential nonfood use only, particularly in bait stations for wasps and hornets. The trade name is Sulfotine. Derivatives of PFOS have also been widely used as stable surfactants in surface coatings to repel soil, oil, and water from paper and fabrics [e.g., in Scotchgard®, in paints and adhesive, in fire-retardant foams, and in a variety of other industrial and domestic applications (U.S. EPA, 2000c)]. Almost 6.5 million pounds were produced in 2000. The amount of this material used for insect control is minor. Toxicology Profile The primary sources of data for LPOS are U.S. EPA (199ge) and Cal EPA (2000c), and for PFOS it is U.S. EPA (2000c). Acute Toxicity LPOS has a moderate acute toxic for mammals by the oral and inhalation routes and a low dermal toxicity. The acute oral LDso of PFOS to rats is 251 mg/kg. IrritationlSensitization LPOS is a severe but reversible eye irritant and a slight skin irritant. Subchronic Toxicity LPOS administered for 90 days in the drinking water decreased body weight, suppressed hematopoiesis, decreased serum triglycerides, and caused hepatic vacuolization in females with LOELs in the range of 0.3 to 3.0 mg/kg/day. Anorexia and convulsions were reported after lethal subchronic doses of PFOS in rats. In parallel with the LPOS study, these signs were accompanied by hepatotoxicity (increased liver enzymes, hepatic vacuolization, and hypertrophy), gastrointestinal effects, and hematologic abnormalities at doses of 2 mg/kg/day or more. In Rhesus monkeys after repeat oral dosing with PFOS, signs of poisoning included emesis, diarrhea, anorexia, hypoactivity, prostration, convulsions, and death. Atrophy of the salivary
glands and pancreas, a marked decrease in serum cholesterol, and lipid depletion in the adrenals were also reported. No monkeys survived beyond 3 weeks at 10 mg/kg/day or beyond 7 weeks at 4.5 mg/kg/day. Many of these responses were also seen in Cynomolgus monkeys given doses of the potassium salt of PFOS for 26 weeks. At the LOEL, 0.75 mg/kg/day, morbidity, decreased body weight, and lowered cholesterol were noted (Seacat et aI., 2001). Hormonal effects included lowered triiodothyronine levels, increased thyroid-stimulating hormone, and lowered estradiollevels in males. Hepatocellular hypertrophy and lipid vacuolization, but no evidence of peroxisomal proliferation, were seen. The microscopic changes in the liver completely reversed after 27 weeks of recovery. It was concluded that adverse effects in this study occurred only at serum levels greater than 100 ppm and were reversible. Chronic Toxicity No data were found regarding chronic toxicity and carcinogenicity since LPOS is not registered for food uses (U.S. EPA, 199ge). Carcinogenicity
See Chronic Toxicity.
MutagenicitylGenotoxicity Neither LPOS nor PFOS is genotoxic in standard tests for mutation (Ames and Chinese hamster ovary cell assay), for chromosomal effects (mouse micronucleus and Chinese hamster ovary cell assays), or for DNA damage and repair (unscheduled DNA synthesis in rat primary hepatocytes) . Reproductive Toxicity In a two-generation reproductive study in rats, dietary PFOS had an LOEL of 0.4 mg/kg/day for second generation offspring based on reductions in body weight. Reversible delays in reflex and physical development raised concerns regarding the potential for developmental neurotoxicity. Considerable mortality was seen in first generation offspring at higher dietary PFOS doses (1.6 and 3.2 mg/kg/day). Developmental ToxicitylTeratogenicity No increased fetal sensitivity compared to the maternal parent was observed in developmental studies in pregnant rats and rabbits. LPOS was not teratogenic in rabbits but at oral doses of 4 mg/kg/day during days 7 through 19 of gestation it did cause increased abortions, postimplantation losses, premature deliveries, and growth retardation (Hen wood et aI., 1994a; U.S. EPA, 199ge). In rats, teratogenicity (cleft palate, edema, decreased bone ossification) and embryolethality were observed, but only at a dose (12 mg/kg/day) that also caused severe maternal toxicity and some maternal mortality (Henwood et aI., 1994b; U.S. EPA, 199ge). Parallel studies with PFOS (U.S. EPA, 2000c) showed similar effects in rats exposed at 5 mg/kg./day. In rabbits exposed to PFOS during gestation, reductions in fetal weight and increases in delayed ossification were seen at doses of 2.5 mg/kg/day and higher. Absorption, Metabolism, and Elimination PFOS is firmly bound to liver proteins and resists clearance in male rats and
57.8 Abbreviations mice. Clearance is also delayed by enterohepatic circulation of the acid, perhaps conjugated as the glucuronide (Vitayavirasuk and Bowen, 1999). The administration of cholestyramine has been shown to speed the elimination of both perfluorooctanecarboxylic acid and PFOS by binding these compounds in the gut leading to enhanced fecal elimination (John son et aI., 1984). In Cynomolgus monkeys the serum half-life of PFOS was 193-224 days (Seacat et aI., 2001). Biochemical Mechanism of Action The biochemical mechanism underlying the acute toxicity of LPOS is unknown. It is not a conventional uncoupler because it is a very strong acid and could not act as a transmembrane protonophore in the typical uncoupler model. Despite this, it caused a rapid and strong increase in respiration in cockroaches immediately after injection (Hollingworth and Gadelhak, 1998) although it has a rather weak ability to increase respiratory rates in mitochondria in vitro. The unfluorinated analog of PFOS, octanesulfonic acid, was completely inactive in causing these effects. The strong surfactant properties of PFOS may cause general membrane disruption. The lack of strong uncoupling activity and the production of a nonspecific ion permeability in mitochondrial membranes have also been reported by Starkov et al. (2001). Like its more extensively studied analog, perfluorooctanoic acid, PFOS is a peroxisome proliferator in rodents (Haughom and Spydevold, 1992; Hosokawa and Satoh, 1993; Sohlenius et aI., 1993). This probably occurs because these compounds resemble long chain fatty acids and induce the peroxisomal enzymes responsible for fatty acid catabolism, but these perfluorinated analogs are not metabolizable. In rats, PFOS caused a pronounced reduction in cholesterol and triacylglycerols in serum coupled with a rise in the liver triacylglycerols and an increase in palmitate oxidation and carboxylesterase activity. The hypolipemic effect was attributed to reduced production of cholesteryl esters and enhanced oxidation of fatty acids in the liver which led to impaired production of lipoprotein particles. In the mouse, a typical induction of the enzyme involved in fatty acid oxidation, peroxisomal catalase, and the w-hydroxylation of lauric acid were observed. PFOS was found to act as a typical peroxisome proliferator in subchronic feeding studies in Cynomolgus monkeys (Seacat et al., 2001). Also like perfluorooctanoic acid, PFOS has been shown to inhibit intercellular communication via gap junctions in rat liver epithelial cells in vitro (Upham et aI., 1998). This action correlates well with the ability to act as a tumor promoter. Perfluorooctanoic acid is a known hepatocarcinogen and, because of its parallel actions as a peroxisome proliferator and inhibitor of gap junctions, it is reasonable to speculate that PFOS may act similarly. The role of peroxisome proliferation in carcinogenesis remains controversial, particularly in regard to the extrapolation of results in rodents to humans who appear to be much less sensitive to this effect. Comprehensive reviews of this topic are provided in an IARC Technical Report (IARC, 1995) and by Cattley et al. (1998) and Roberts (1999).
1247
Human Toxicology In routine screening, PFOS was detected in the serum of fluorochemical manufacturing workers with a mean value of approximately 2 ppm (Olsen et al., 1999). PFOS levels less than 6 ppm did not cause any changes in serum hepatic enzymes, cholesterol, or lipoproteins and no evidence of health effects were observed. In tests of blood bank samples from the United States, PFOS was identified at concentrations in the general range of 10-100 ppb in serum with mean levels of 30-54 ppb (Hansen et aI., 2001; U.S. EPA, 2000c) indicating general human exposure to PFOS and/or its derivatives. Because of these observations, coupled with the widespread industrial and domestic use of perfluorooctanesulfonyl surfactants, their long environmental persistence, and indications of the resultant bioaccumulation of PFOS in humans and the environment, their manufacturer, the 3M Company, has recently decided to phase out their production by the end of 2002 (U.S. EPA,2000c). Absorption, Metabolism, and Excretion PFOS is well absorbed after an oral dose and distributes primarily in the serum and liver. It is not metabolized. Elimination is slow and occurs through both the urine and feces. A limited study with retired manufacturing workers indicated an elimination half-life of approximately 4 years (U.S. EPA, 2000c). Environmental Fate and Toxicity LPOS is highly toxic to birds on acute exposure, but it is only slightly to moderately toxic to fish and the toxicity to Daphnia is low. PFOS is an extremely stable compound in the environment with little propensity for degradation chemically, photochemically, or microbiologically (U.S. EPA, 2000c). Because of the extensive industrial use of PFOS derivatives, it can be detected in low amounts in wildlife worldwide including the Arctic and remote Pacific Ocean locations (Giesy and Kannan, 2001). The levels found were typically in the range of 10-500 ng/g tissue, but up to 3680 ng/g was detected in the livers of mink in the midwestem United States. The highest levels were found in more industrialized areas. These authors present evidence that PFOS can bioaccumulate within food chains (e.g., mink fed fish containing PFOS at 120 ng/g wet weight accumulated about 2600 ng/g in their liver). So far, no toxic effects have been ascribed to the widespread occurrence of this compound in fish, birds, and marine mammals, and the concentrations detected appear to be less than those known to cause adverse effects. The mechanisms of degradation of the various perfluoroctanesulfonyl compounds to PFOS, and the routes of global transport and biological accumulation, remain to be established. Because of its low Henry's Law constant value, PFOS is unlikely to volatilize from water to air.
57.8 ABBREVIATIONS 2,4-DNP
2,4-dinitrophenol
ADI
Acceptable daily intake
1248
CHAPTER 57
Pesticides Affecting Oxidative Phosphorylation
BCF
Biological concentration (bioconcentration) factor
ACKNOWLEDGEMENTS
CAS Reg. No. DC CD
Chemical Abstracts Service Registry Number Dicyclohexy lcarbodiimide
EC50
Concentration that causes an effect at 50% of the maximum level Female Henry's law constant Concentration needed to cause 50% inhibition of a process Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and the WHO Core Assessment Group on Pesticide Residues
I would like to recognize the assistance of Dr. Jerry Baron and those colleagues from agrochemical companies who provided information on individual pesticides, Dr. Satoru Miyazaki for his assistance in the translation of documents from Japanese, and Dr. Kurunthachalam Kannan for helphul discussions on the environmental effects of organotins and perfluoroctanesulfonic acid and its derivatives.
F
HLC 150 JMPR
Equilibrium constant for binding to soil
GABA LC50 LD50 LOEC LOEL logP
LPOS M
Equilibrium constant for binding to soil, adjusted for organic matter content Gamma-aminobutyric acid Concentration lethal to 50% of the population Dose lethal to 50% of the popualtion Lowest concentration causing an observable effect Lowest observable (adverse) effect level Logarithm of the l-octanol: water partition coefficient Lithium perfluorooctanesulfonate Male
ROS
Milligrams of compound per kilogram of body weight Melting point N -deethyl sulfluramid (perfluorooctanesulfonamide) No observable (adverse) effect level Ornithine decarboxylase Oxidative phosphorylation Perfluorooctanesulfonic acid Negative logarithm of the dissociation constant (Ka) Parts per billion Part per million Parts per trillion Coenzyme Q (ubiquinone) Unit cancer potency factor from the USEPA cancer risk model Reactive oxygen species
SmP
submitochondrial particle
UC50
Concentration causing 50% uncoupling of respiration
v.p.
Vapor pressure Water solubility
mglkg
m.p. NDES NOEL ODC Oxphos PFOS pKa ppb ppm ppt Q
Qr
W.s.
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Pesticides Affecting Oxidative Phosphorylation
Takeshige, K., and Minakami, S., (1979). NADH- and NADPH-dependent formation of superoxide anions by bovine heart submitochondrial particles and NADH-ubiquinone reductase preparation. Biochem. J. 180, 129-135. Taketa, E, Siebenlist, K., Kasten-Jolly, J., and Palosaari, N. (1980). Interaction oftriethyltin with cat hemoglobin: identification of binding sites and effects on hemoglobin function. Arch. Biochem. Biophys. 203, 466-472. Tanabe, S. (1999). Butyltin contamination in marine mammals-A review. Marine Poll. Bull. 39,62-72. Terada, H. (1981). The interaction of highly active uncouplers with mitochondria. Biochim. Biophys. Acta 639, 225-242. Thomson, A. M., Meyer, D. J., and Hayes, J. D. (1998). Sequence, catalytic properties and expression of chicken glutathione-dependent prostglandin D2 synthase, a novel class Sigma glutathione-S -transferase. Biochem. J. 333, 317-325. Tipton, K. E, and Singer, T. P. (1993). Advances in our understanding of the mechanism of the neurotoxicity of MPTP and related compounds. 1. Neurochem.61, 1191-1206. Tomlin, C. D. S. (2000). "The Pesticide Manual," 12th ed. British Crop Protection Council, Farnham, Surrey, UK. Treacy, M., Miller, T., Black, B., Gard, I., Hunt, D., and Hollingworth, R M. (1994). Uncoupling activity and pesticidal properties of pyrroles. Biochem. Soc. Trans. 22, 244-247. Tsao, R, and Eto, M. (1991). Photolysis of flutolanil fungicide and the effects of some photosensitizers. Agric. Bio!. Chem. 55, 763-768. Tsuda, T., Aoki, S., Kojima, M., and Harada, H. (1990). The influence of pH on the accumulation of tri-n-butyltin chloride and triphenyltin chloride in carp. Comp. Biochem. Physio!. C 95, 151-153. Tsuda, T., Aoki, S., Kojima, M., and Fujita, T. (1992). Accumulation and excretion of pesticides used in golf courses by carp (Cyprinus carpio) and willow shiner (Gnathopogon caerulescens). Comp. Biochem. Physio!. IOIe, 6366. Tsukihara, T., Aoyama, H., Yamashita, K., Tomizaki, T., Yamaguchi, H., Shinzawa-Itoh, K., Nakashima, R, Yaono, R., Yoshikawa, S., Yamashita, E., and Shinazawa-Itoh, K. (1996). The whole structure of the 13-subunit oxidized cytochrome oxidase at 2.8 angstroms. Science 272, 1136-1144. Turner, J., and Joseph, R (1998). "In Vitro Reactivity of Azoxstrobin with Glutathione and Glutathione Transferase." Absl. 9th Internal. Congr. Pestic. Chem., No. 5C-002. Turrens, J. E, and Boveris, A. (1980). Generation of superoxide anion by the NADH dehydrogenase of bovine heart mitochondria. Biochem. J. 191, 421-427. Tyler, D. (1992). "The Mitochondrion in Health and Disease." VCH, New York. Udeani, G. 0., Gerhauser, c., Thomas, C. E, Moon, R. c., Kosmeder, J. W., Kinghorn, A. D., Moriarty, R M., and Pezzuto, J. M. (1997). Cancer chemopreventive activity mediated by deguelin, a naturally occurring rotenoid. Cancer Res. 57, 3424-3428. Ueno, H., Miyoshi, H., Inoue, M., Niidome, Y., and Iwamura, H. (1996). Structural factors of rotenone required for inhibition of various NADHUbiquinone oxidoreductases. Biochim. Biophys. Acta 1276, 195-202. Unai, T., Cheng, H.-M., Yamamoto, I., and Casida, J. E. (1973). Chemical and biological O-demethylation of rotenone derivatives. Agr. Bio!. Chem. 37, 1937-1944 Upham, B. L., Deocampo, N. D., Wurl, B., and Trosko, J. E. (1998). Inhibition of gap junctional intercellular communication by perfluorinated fatty acids is dependent on the chain length of the fluorinated tail. Int. J. Cancer 78, 491-495. US. EPA (1989a). Carboxin. In "Drinking Water Health Advisory. Pesticides," pp. 151-164. Lewis, Brighton, MI. U.S. EPA (1989b). Dinocap: Notice of intent to cancel registrations; conclusion of Special Review. Fed. Reg. 54(23), 5908-5920. U.S. EPA (1989c). Dinoseb. In "Drinking Water Health Advisory. Pesticides," pp. 323-334. Lewis, Chelsea, MI. U.S. EPA (1989d). "EPA Pesticide Fact Sheet 3/89: Sulfluramid (GX-071)." U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1994). "Reregistration Eligibility Decision (RED): FenbutatinOxide." EPA 738-R-94-024, U.S. Environmental Protection Agency, Washington, DC.
U.S. EPA (1997a). BASF Corporation; Pesticide tolerance petition filing. Fed. Reg.62, 11450-11453. U.S. EPA (1997b). "Pesticide Fact Sheet: Azoxystrobin." Office of Prevention, Pesticide and Toxic Substances, U.S. Environmental Protection Agency, Publication 7501C, Washington, DC. U.S. EPA (1997c). Zeneca Ag products; Pesticide tolerance petition filing. Fed. Reg.62,11441-1I447. US. EPA (1997d). "Chlorfenapyr-129093: Health Effects Division Risk Characterization for Use of the New Chemical Chlorfenapyr inion Cotton (5F4456)." U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1998a). Pyridaben (Sanmite) pesticide tolerances for emergency exemptions 9/98. Fed. Reg. 63, 53294-53301. U.S. EPA (l998b). "Pesticide Fact Sheet: Kresoxim-Methyl." Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1998c). "Reregistration Eligibility Decision (RED): Hydramethylnon." Document 738-R-98-023, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1998d). "Reregistration Eligibility Decision (RED). Bromoxynil." EPA738-R-98-013, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (I 998e). "Reregistration Eligibility Decision (RED). Paranitophenol." EPA 738-R-97-016, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (l998f). "Reregistration Eligibility Decision: Rodenticide Cluster:" 738-R-98-007, United States Environmental Protection Agency, Washington, DC. U.S. EPA (l999a). Notice of filing of pesticide petitions. Fed. Reg. 64, 80908102. U.S. EPA (1999b). "Pesticide Fact Sheet: Trifloxystrobin." Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1999c). "Reregistration Eligibility Decision (RED): Triphenyltin Hydroxide (TPTH)." EPA 738-R-99-01O, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1999d). "Reregistration Eligibility Decision (RED): 3-TrifluoroMethyl-4-Nitro-Phenol and NicIosamide." EPA 738-R-99-007, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (199ge). "New Pesticide Fact Sheet: Lithium Perfluorooctane Sulfonate (LPOS). EPA-730-F-99-009, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (2000a). Pyridaben; Pesticide tolerance. Fed. Reg. 65,43704-43713. U.S. EPA (2000b). Notice of filing pesticide petitions to establish and to extend tolerances for certain pesticide chemicals in or on food. Fed. Reg. 65, 76253-76258. U.S. EPA (2000c). Perfluorooctyl sulfonates; proposed significant new use rule. Fed. Reg. 65,62319-62333. van den Berg, K. J., van Raaij, J. A. G. M., Bragt, P. c., and Notten, W. R. E (1991). Interactions of halogenated industrial chemicals with transthyretin and effects on thyroid-hormone levels in vivo. Arch. Toxicol. 65, 15-19. van der Kerk, G. J. M., and Luijten, J. G. A. (1954). Investigations on organotin compounds. HI. The biocidal properties of organotin compounds. J. App!. Chem. 4, 314-322. Vander Meer, R K., Lofgren, C. S., and Williams, D. E (1985). Fluoroaliphatic sulfones: A new class of delayed-action insecticides for control of Solenopsis invicata (Hymenoptera: Formicidae). J. Econ. Entomol. 78, 1190-1197. Vander Meer, R. K., Lofgren, C. S., and Williams, D. E (1986). Control of Solenopsis invicta with delayed-action fluorinated toxicants. Pestic. Sci. 17, 449-455. Vander Meer, R K., Lofgren, C. S., and Williams, D. E (1987). Fluorinated sulfonamides-A new class of delayed-action toxic ants for fire ant control. ACS Sympos. Ser. 355, 226-240. van Gemert, M., and Killeen, J. (1998). Chemically induced myelinopathies. Internat. J. Toxico!. 17,231-275. van Ginkel, C. J., and Sabapathy, N. N. (1995). Allergic contact dermatitis from the newly introduced fungicide fluazinam. Contact Dermat. 32, 160-162. van Lier, R .B., and Cherry, L. D. (1988). The toxicity and mechanism of action of bromethalin: A new single-feeding rodenticide. Fund App!. Tox. 11, 664672.
References
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White, G. A., and Thorn, G. D. (1975). Structure-activity relationships of carboxamide fungicides and the succinic dehydrogenase complex of Cryptococcus laurentii and Ustilago maydis. Pestic. Biochem. Physiol. 5, 380-395. WHOJIPCSJILO (1993). "International Chemical Safety Cards: Ioxynil." ICSC No. 0900, World Health Organization, Geneva. Wiesner, P., Popp, B., Schmid, A., Benz, R., and Kayser, H. (1996). Isolation of a mitochondrial porin of the fly Protophormia: Porin modification by the pesticide CGA 140'408 studied in lipid bilayer membranes. Biochem. Biophys. Acta 1282, 216-224. Wiggins, T. E., and Jager, B. J. (1994). Mode of action of the new methoxyacrylate antifungal agent ICIA5504. Biochem. Soc. Trans. 22, 68S. Williams, W. (1999). Pirate fear. Sci. Amer. 281, 26-30. Williamson, R. L., and Metcalf, R. L. (1967). Salicylanilides: A new group of active uncoup1ers of oxidative phosphorylation. Science 158, 1694-1695. Wishkovsky, A., Mathews, E. S., and Weeks, B. A. (1989). Effect of tributyltin on chemiluminescent response of phagocytes from three species of estuarine fish. Arch. Environ. Contam. Toxicol. 18, 825-831. Wood, E., Latli, B., and Casida, J. E. (1996). Fenazaquin acaricide specific binding sites in NADH: Ubiquinone oxidoreductase and apparently the ATP synthase stalk. Pestic. Biochem. Physiol. 54, 135-145. Wu, R. M., Chang, Y. C., and Chiu, H. C. (1990). Acute triphenyltin intoxication: A case report. J. Neural. Neurosurg. Psychiatry 53, 356-357. Wulf, R. G., and Byington, K H. (1975). On the structure-activity relationships and mechanism of organotin induced, non energy dependent swelling of liver mitochondria. Arch. Biochem. Biophys. 167, 176-185. Xia, D., Yu, C.-A., Kim, H., Xia, J.-Z ., Kachurin, A. M., Zhang, L., Yu, L., Deisenhofer, J., Yu, C. A., and Xian, J. Z. (1997). Crystal structure of the cytochrome bCj complex from bovine heart mitochondria. Science 277, 6066. Yamabe, Y., Hoshino, A., Imura, N., Suzuki, T., and Himeno, S. (2000). Enhancement of androgen-dependent transcription and cell proliferation by tributyltin and triphenyltin in human prostate cancer cells. Toxicol. Appl. Pharmacol. 169, 177-184. Yamada, H., and Sasaki, Y. E (1993). Organotins are co-clastogens in a whole mammalian system. Mutat. Res. 301, 195-200. Yamamoto, 1. (1969). Mode of action of natural insecticides. Resid. Rev. 25, 161-174. Yamamoto, 1., Unai, T., Ohkawa, H., and Casida, J. E. (1971). Stereochemical considerations in the formation and biological activity of the rotenone metabolites. Pestic. Biochem. Physiol. 1, 143-150. Yamano, T., and Morita, S. (1993). Effects of pesticides on isolated rat hepatocytes, mitochondria, and microsomes. Arch. Environ. Contam. Toxicol. 25, 271-278. Yamano, T., and Morita, S. (1995). Effects of pesticides on isolated rat hepatocytes, mitochondria, and microsomes. H. Arch. Enviran. Contam. Toxicol. 28, 1-7. Yumita, T., and Yamamoto, 1. (1982). Photodegradation of mepronil. J. Pestic. Sci. 7, 125-132. Zazueta, c., Reyes-Vivas, H., Bravo, c., Pichardo, J., Corona, N., and Chavez, E. (1994). Triphenyltin as inductor of mitochondrial membrane permeability transition. J. Bioenerg. Biomembr. 26, 457-462. Zhang, Y., Marcillat, 0., Giulivi, c., Enster, L., and Davies, K. J. (1990). The oxidative inactivation of mitochondrial electron transport chain coponents and ATPase. J. BioI. Chem. 265, 1633-1636. Zychlinski, L., and Zolnierowicz, S. (1990). Comparison of uncoupling activities of chlorophenoxy herbicides in rat liver mitochondria. Toxicol. Lett. 52, 25-34.
CHAPTER
58 Pyrethroid Chemistry and Metabolism Hideo Kaneko and Junshi Miyamoto Sumitomo Chemical Co., Ltd.
Natural pyrethrins, the insecticidal ingredient occurring in the flowers of Tanacetum cinerariaefolium (also known as Chrysanthemum cinerariaefolium or Pyrethrum cinerariaefolium), have been used widely for human and animal health protection by controlling indoor pest insects such as cockroaches, houseflies, and mosquitoes. Natural pyrethrins consist of six compounds (pyrethrin I and II; jasmolin I and II; and cinerin I and II). The investigation of the chemical structures of natural pyrethrins was started in 1920s and their absolute stereochemistry was completed, and elucidated in the early 1970s (Chamberlain et aI., 1998). Along with the investigations, extensive efforts on modification of chemical structures have been made in many laboratories to improve chemical properties in terms of stability in the environment (air, light, and heat) as well as better biological performance (higher selective toxicity). From these investigations, many synthetic pyrethroids have been elaborated and worldwide used for agricultural and sanitary purposes. They can be classified into the so-called first- and second-generation pyrethroids. The characteristic feature of the first-generation pyrethroids, which are esters of chrysanthemic acid and alcohols having furan ring and terminal side chain moieties, is to be highly sensitive to light, air, and temperature. Therefore, these pyrethroids have been used mainly for control of indoor pests. On the other hand, the second-generation pyrethroids, which commonly have 3phenoxy benzyl alcohol derivatives in the alcohol moiety, have excellent insecticidal activity as well as sufficient stability in outdoor conditions by replacement of photolabile moieties with dichlorovinyl, dibromovinyl substituent and aromatic rings. Thus the second-generation pyrethroids are used worldwide for agricultural pests. Many in vivo and in vitro metabolism studies of synthetic pyrethroid insecticides including their chiral and geometrical isomers have been carried out in mammals for safety assessment. However, all detailed metabolism data have not been published. In these cases, the reports of joint World Health Organization-Food and Agricultural Organization (WHOIFAO) expert meetings on pesticide residues and of the International Handbook of Pesticide Toxicology Volume 2. Agents
Programme on Chemical Safety (IPCS), Environmental Health Criteria (WHO), were referred to. This chapter deals with metabolism of more than 20 pyrethroid insecticides in laboratory animals and human beings in alphabetical order, being focused mainly on in vivo metabolism. The positions labeled with 3H or 14C are stated as the acid or alcohol moiety unless specified, because the acid or alcohol moiety does not undergo further degradation to a smaller pieces. Data on nomenclature and physical chemistry are mainly cited from A World Compendium, The Pesticide Manual, 11th edition (Tomlin, 1997), and Metabolic Pathways of Agrochemicals, Part 2, Insecticides and Fungicides (Roberts and Hutson, 1999). In addition, there are several excellent reviews about mammalian metabolism of pyrethroids; however, the book The Pyrethroid Insecticides (Leahey, 1985) was mainly referred to.
58.1 ALLETHRIN (BIOALLETHRIN, d-ALLETHRIN, S-BIOALLETHRIN) Chemical Name (RS)-3-allyl-2-methyl-4-oxocyclopent-2eny1 ( 1RS)-cis-trans-2,2-dimethy1-3-(2-methy 1prop-1-enyI) cyclopropanecarboxylate. Synonyms Allethrin (BSI, ISO, JMAF, ESA) is the common name in use. The trade name is Pynamin. d-Allethrin (trade name Pynamin Forte) is an ester of (1R)-cis-transchrysanthemic acid and (RS)-allethrolone. Bioallethrin is an ester of (1 R)-trans-chrysanthemic acid and (RS)-allethrolone. S-Bioallethrin is an ester of (1 R)-trans-chrysanthemic acid and (S)-allethrolone. The CAS registry number are 584-79-2 (allethrin, bioallethrin) and 28434-00-6 (S-bioallethrin). Physical and Chemical Properties (d-AlIethrin) The empirical formula is C19H2603; molecular weight is 302.4. Its form is a yellow to amber viscous liquid; its specific gravity is 1.01 at 20 D C; log Kow = 4.96. It is practically insoluble in
1263
Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
1264
CHAPTER 58
Pyrethroid Chemistry and Metabolism
HOH2C~
~oo~ AIlelhrin
COO~ ~
0
, 0"
HOOC~
0
/
COO~~
COCA
/ HO~
A o,id
0
COO~
\
AM B
COO~
AMC
AMA Figure 58.1
[HOOC~
OH
0
: VCH20H HOOC~
o
coo~
0
HOOC~
AIlethrolone
HOOC~
COO~
A o'd
HOOC~
COOH
OHC~
<E---
coo~
0
0
1
0
Metabolic pathways of allethrin in animals.
water, but is soluble in most organic solvents. It is unstable to DV light and is hydrolyzed in alkaline media.
58.2 BIFENTHRIN
Metabolism When allethrin labeled with 14C in the acid moiety or with 3H or 14C in the alcohol moiety was orally administered to rats at 1-5 mg/kg, the 14C and 3H derived from the acid and alcohol moieties were excreted into the urine (47-51 %) and feces (27-29%) within 48 hr after administration. Most of the metabolites excreted into the urine are ester linkage-cleaved products [chrysanthemic dicarboxylic acid (CDCA) and allethrolone] and ester linkage-retaining products. However, the fecal metabolites are not adequately characterized (Elliott et aI., 1972; IPCS, 1989). The major metabolic reactions (Fig. 58.1) of allethrin are as follows: (1) hydrolysis of the ester linkage, (2) formation of the 2,3-diol from the allyl moiety, (3) hydroxylation at the methylene position of the allyl moiety, (4) hydroxylation at one of the gem-dimethyl groups, and (5) oxidation at the trans-methyl group of the isobutenyl moiety. In addition, epoxidation of the double bond of the acid moiety takes place in vitro in mouse liver microsomes (Class et aI., 1990). There are some species differences in in vitro microsome oxidation sites of allethrin between rats and mice: rat microsomes appear to preferentially oxidize the trans-methyl group of the isobutenyl moiety. On the other hand, the major oxidation sites by mouse microsomes are the trans-methyl group of the isobutenyl group, the methylene position of the allyl group and the 7,8 double bond of the acid moiety (Class et aI., 1990).
cis- 3-(2-chloro-3,3,3-trifiuoroprop-1-eny1)-2,2-dimethylcyclo-
Chemical Name
2-Methylbiphenyl-3-ylmethyl (Z)-(1RS)-
propanecarboxylate. Synonyms Bifenthrin (BSI, ANSI, ISO) is the common name in use. The trade name is Talster. Code designations include FMC 54800. The CAS registry number is 82657-04-3. Physical and Chemical Properties The empirical formula is C23H22CIF302; molecular weight is 422.9. Its form is a viscous liquid or a crystalline, or waxy solid; its specific gravity is 1.21 at 25°C; log Kow ~ 6. It is less soluble (0.1 mg/l) in water and is soluble in most organic solvents. It is rather stable in natural daylight and water (pH 5-9) . Metabolism Male and female rats were treated with 14C_ bifenthrin labeled in the acid or alcohol moiety at single oral doses of 4 and 35 mg/kg. 14C was rapidly excreted into feces and urine, and the excretion rates of the 14C to feces and urine were 66-83% and 13-25%, respectively. Highest residues were found in the fat, with values of slightly more than 1 ppm after low-dose administration and 8 and 16 ppm in males and females, respectively, after application of the high dose. Residue levels in other organs were in most cases <0.2 ppm after lowdose administration and <1 ppm after high-dose administration (FAOIWHO, 1992).
58.3 Cycloprothrin
1265
HOHi
COOII )t
CF 3
-E--
)tCOOCH'
Cl HP acid
CF 3
CH 3
\
Bilenthrin
HOOC~ CH 3
Cl
t -E-- HOH 2C
CH 3
,..;
",..;
BP alcohol
BP acid
HOH
n
c
~
~3 O(OCH 3l2
2 -
)fc
CF 3
CF 3
'ACOOCH'
(0'Cl ~ CF
0 OCH,
3
CH 3
Cl
OH
)lo,L
20H
-E--
"cOOH
CF 3
ct-hydroxymethy I
n
3
4'--D H-B P acid
~
OH
OH 3
X )(A'
HOH
·COOH
CF 3
Cl
TFP acid
I'" -1 U
CH
3'- or 4'--D H-hydroxy methyl -bifenthrin
4'--D H-bilenthrin
HOOe
Figure 58.2
t
)
{t
dimethoxy-B P acid
Cl hy droxy methy I-bile nth rin
HOH 2C
dimethoxy BP alcohol
HOOC
)lo, ~COOCH2
)
Cl
tmns-hyd roxymethy I TFP acid
HOOC ,..; OCH CH 3 4'-methoxy BP acid 3
Metabolic pathways of bifenthrin in animals.
The major fecal metabolites possessed intact ester linkage hydroxylated in the acid or alcohol moiety such as hydroxymethyl-bifenthrin, 4'-OH-bifenthrin, and 3'- or 4'-OHhydroxy methyl bifenthrin. Ester-cleaved products derived from mono- and dihydroxylated parent compounds were also detected. On the other hand, the majority of urinary metabolites were ester-cleaved products such as 4'-OH-BPacid (4'-hydroxy2-methyl-3-phenylbenzoic acid), BPacid (2-methyl-3-phenylbenzoic acid), 4'-OH-BPalcohol (4'-hydroxy-2-methyl-3phenylbenzyl alcohol), dimethoxy-BPacid, 4'-methoxy BPacid, dimethoxy BPalcohol, BPalcohol, TFPacid [3-(2-chloro-3,3,3triftuoro-l-propeny 1)-2,2-dimethyl-cyclopropanecarboxylic acid], cis- and trans-hydroxymethyl TFPacid. The major metabolic pathways (see Fig. 58.2) are considered to be hydrolysis of ester linkage, oxidation at the methyl group of the acid moiety and at the 3'- and 4' -positions of the phenyl group, and O-methylation. The conjugation reactions are considered to take place; however, detailed information is not available (FAOIWHO, 1992). The tissue residues were examined after oral administration of l4C-bifenthrin at 0.5 mg/kg/day for 70 days. The peak 14C concentrations on an average were 9.6 ppm in fat, 1.7 ppm in skin, 0.4 ppm in liver, 0.3 ppm in kidney, 1. 7 ppm in ovaries, 3.2 ppm in sciatic nerve, 0.06 ppm in whole blood, and 0.06 ppm in
plasma. Analyses were extended for an additional 85 days following cessation of dosing (depuration phase). Half-lives of 51 days (fat), 50 days (skin), 19 days (liver), 28 days (kidney), and 40 days (ovaries and sciatic nerve) were estimated from 14C_ depuration. Analysis of the fat revealed that the parent chemical accounted for a majority (65-85%) of the 14C-residues in fat.
58.3 CYCLOPROTHRIN Chemical Name (RS)-a-Cyano-3-phenoxybenzyl (RS)-2,2dichloro-l-( 4-ethoxyphenyl) cyclopropanecarboxylate. Synonyms Cycloprothrin (BSI, ISO) is the common name in use. The trade name is Cyclosaal. Code designations include GH-414 and NK-8116. The CAS registry number is 63935-38-6. Physical and Chemical Properties The empirical formula is C26H21 ChN04; molecular weight is 482.4. Its form is a yellow to brown viscous liquid; its specific gravity is 1.256 at 25°C; log Kow = 4.19. It is less soluble (0.091 mg/l) in water at 25°C, but is soluble in most organic solvents.
1266
CHAPTER 58
d
Pyrethroid Chemistry and Metabolism
~"
:
HOCHO 2 4
I
COO~H~°Y)~ CN
V
V
~X~:OO\H~OY) V V
C2H50L:JJ
.
CN
Cycloprothrin
\
~~:OOIH~OY) V V
HOL:JJ
.
CN
HO --{;ycloprothrin
1
~X~:OOH
HOL:JJ
'
HO -Dcid Figure 58.3
Metabolic pathways of cyc1oprotbrin in animals.
Metabolism On single or consecutive (once a day for 7 days) oral administration of cycloprothrin labeled with 14C in the acid moiety to male rats at 50 mg/kg/day, the 14C was rapidly and almost completely eliminated into urine (36%) and feces (63%) within 7 days after administration. 14C tissue residue levels reached maximum 3 hr after single oral administration and thereafter decreased with time. 14C tissue residue levels after repeated administration were about 3.6 times higher compared with those of a single oral dose. 14C residue levels were relatively high in the fat and skin and 14C depletion from these tissues was slower than from other tissues (Seguchi et ai., 1991). HO-acid was a predominant metabolite in urine and feces, accounting for 39% of the dose. In addition, HO-cycloprothrin, C2HsO-acid, and HOC2H40-acid were also found in the feces and urine as minor metabolites. The major metabolic pathways of cycloprothrin (Fig. 58.3) are cleavage of the ester linkage and oxidation at the ethoxy position of the acid moiety (Seguchi et ai., 1991). Although metabolism of the alcohol moiety is not available, it can be predicted on the basis of metabolism of pyrethroids having the same alcohol moiety such as fenvalerate and fenpropathrin.
58.4 CYFLUTHRIN Chemical Name (RS)-a-Cyano-4-ftuoro-3-phenoxybenzyl (IRS)-cis- trans- 3-(2,2-dichloroviny1)-2,2-dimethylcyclopro-
panecarboxy late. Synonyms Cyftuthrin (BSI, ISO, BAN) is the common name in use. Trade names are Baythroid, Baygon aerosol, and Solfac. Code designations include Bay FCR 1272. The CAS registry number is 68359-37-5. Physical and Chemical Properties The empirical formula is C22HlSC12FN03; molecular weight is 434.3. Its form is a colorless crystal; its specific gravity is 1.28 at 20°C; log Kow = 6. It is less soluble (0.002-0.003mg/l) in water at 20°C, but is soluble in most organic solvents. It is rather stable at room temperature and in acidic water, but is unstable in alkaline water. Metabolism The acid moiety of cyftuthrin is the same as those of permethrin and cypermethrin; accordingly, metabolism of the acid moiety was not investigated, because the acid moiety should undergo the same metabolic fate after ester hydrolysis. 14C derived from the alcohol moiety was rapidly and com-
58.5 Cyhalothrin Cl ~I
X ~ ~ ~COOCH~OA/
V
~COOH
/
Cl
DCVA
CF 3
eN Cyhalothrin
CHO ---7
a'"
I O I o F
/-
glucuronide
FPS aid
C fluthrin
r?'y0~CH20H
\
~~
C~COOH
r?'y0l(V00H
CF
VF~
VFV
1267
3
FPS acid
FPS alc
~l HOV
canjugates
CH 20H
r?'y0~COOH
r?'y0yY0NHCH,COOH
VF~
j Cl
FV
~COOH CF 3
4'-oH-fPS acid
/
X'
7 HOOC~O~ !
~
~
,I",o,"old,
r?'y0~CONHCH,COOH
HO
V
FV
glycine conjugate
58.5 CYHALOTHRIN (l.-CYHALOTHRIN) Chemical Name (RS)-a-cyano-3-phenoxybenzyl (Z)-(lRS)cis-3-(2-chloro-3,3,3-trifiuoropropenyl)-2,2-dimethylcyclopropanecarboxylate; A-cyhalothrin, (RS)-a-cyano-3-phenoxy-
PBacid
OH
4'-{) H-PBacid
Figure 58.4 Metabolic pathways of cyfluthrin in animals.
pletely excreted into urine and feces after single oral administration of 14C-alcohol-Iabeled preparation to rats at 0.5 and 10 mg/kg, 55-70 and 25-35% ofthe dose being excreted into urine and feces, respectively. Excretion of 14C into bile was about 34%. The fat and sciatic nerve showed relatively higher 14C tissues residues (FAOIWHO, 1986). Major metabolic pathways (Fig. 58.4) are ester cleavage and oxidation at the 4' position of the alcohol moiety. Major metabolites were 4'-OH-FPBacid [4-fiuoro-3-(4'hydroxyphenoxy)benzoic acid] and its conjugates (glucuronide or sulfate), accounting for about 40-50% of recovered urinary l4C from rats given the labeled preparation at 0.5 mg/kg. Glycine conjugates of FPBacid (4-fiuoro-3-phenoxybenzoic acid) and 4'-OH-FPBacid were also found as minor metabolites. The hydroxylation at the 4'-position of the alcohol moiety is major in rats for cyfiuthrin as with pyrethroids having the 3-phenoxybenzyl alcohol or a-cyano-3-phenoxybenzyl alcohol, although cyfiuthrin has the fiuoro atom in 4-position of the benzyl ring (FAOIWHO, 1986). The metabolism of cyfiuthrin was examined in humans after exposure of nine male volunteers to aerosol (unlabeled cyfiuthrin). The cis- and trans-acid metabolites and FPBacid were detected, indicating that the ester hydrolysis occurs in humans (Leng et aI., 1997).
~ ~ HOOC~OA/
1
sulfate Figure 58.5
Metabolic pathways of cyhalothrin in animals.
benzyl (Z)-( 1R)-cis- 3-(2-chloro-3,3,3-trifiuoropropenyl)-2, 2-dimethylcyclopropanecarboxylate. Synonyms Cyhalothrin (BSI, ISO, BAN) and A-cyhalothrin are the common names in use. Trade names are Cyhalon and Grenade for cyhalothrin and Karate, Warrior, and Icon for A-cyhalothrin. Code designations include PP563 and ICI146814 for cyhalothrin and PP321 and ICIA0321 for A-cyhalothrin. The CAS registry numbers are 68085-85-8 for cyhalothrin and 91465-08-6 for A-cyhalothrin. Physical and Chemical Properties (Cyhalothrin) The empirical formula is C23H19ClF3N03; molecular weight is 449.9. Its form is a yellow to brown viscous oil; its specific gravity is 1.25 at 25°C; log Kow = 6.8. It is less soluble (0.004 mg/l) in water at 20°C, but is soluble in most organic solvents. It is stable to light and unstable in alkaline medium. Metabolism Cyhalothrin was rapidly excreted into urine and feces after oral administration of 14C-Iabeled acid or alcohol preparation to rats at 1 or 25 mg/kg, and 14C was excreted into feces (40-65%) and into urine (20-40%) for 7 days. The fat showed the highest residue compared with other tissues. Major metabolic reactions (Fig. 58.5) are ester hydrolysis and hydroxylation at the alcohol moiety. The metabolic fates of the alcohol moiety, a-cyano-3-phenoxybenzyl alcohol, was the
1268
CHAPTER 58
Pyrethroid Chemistry and Metabolism
same as those of pyrethroid insecticides having the same alcohol moiety such as fenvalerate, cypermethrin, and deltamethrin. The cyano group of the alcohol moiety of cyhalothrin is expected to undergo conversion to SCN ion. The major metabolites of the acid moiety is cyclopropylcarboxylic acid and its glucuronide and those from the alcohol moiety is PBacid, 4'-OH-PBacid and sulfate of 4'-OH-PBacid (FAOIWHO, 1984; IPCS, 1990a). A-Cyhalothrin is manufactured by crystallization of the more active pair of enantiomers from cyhalothrin. The comparative metabolism of A-cyhalothrin with or without enantiomer pair A and cyhalothrin revealed that enantiomer pair A had little or no effect on the absorption, distribution tissue retention, or metabolic profiles, implying that enantiomers of cyhalothrin behave independently (IPCS, 1990a).
58.6 CYPERMETHRIN (a-, ~ -CYPERMETHRIN)
p-, 0,
Chemical Name (RS)-a-Cyano-3-phenoxybenzyl (lRS)-eistrans-3-(2,2-dichlorovinyl)-2,2-dimethylcyclopropanecarboxylate. a-Cypermethrin is a racemate comprising «S)-(l R)-eis) and «R)-(1S)-cis). ,B-Cypermethrin is a mixture comprising two enantiomeric pairs in the ratio of about 2: 3. e-Cypermethrin is a mixture of enantiomers «S)-(lR)-trans) and «R)-(l S)-trans) in the ratio of 1 : 1. ~ -Cypermethrin is a mixture comprising «S)-(lRS)-cis-trans). Synonyms Cypermethrin (BSI, ISO, ANSI, BAN) is the common name in use. Trade names are Agrothrin, Arrivo, Cymbush, Cymperator, Cynoff, Ripcord, and several other names. Code designations include NRDCI49, PP383, FMC30980, WL43467, and LE79-600. The CAS registry number is 5231507-8. Physical and Chemical Properties The empirical formula is C22H19ChN03; molecular weight is 416.3. Its form is a yellow-brown viscous semisolid; its specific gravity is 1.23 at 20°C; log Kow = 6.6. It is less soluble (0.004 mg/l) in water at 20°C, but is soluble in most organic solvents. It is relatively stable to light in weakly acidic water, but is unstable in alkaline medium. Metabolism On single oral administration of each of 14C_ (lRS)- trans- and (lRS)- cis-cypermethrin labeled in the benzyl ring, the cyclopropane ring, or the CN group to male and female rats at 1-5 mg/kg, 14C from the acid and alcohol moieties was rapidly and almost completely excreted into the urine and feces. The 14C from the CN group was relatively slowly excreted in the urine and feces, the total recovery being 50-67%. The tissue residues of rats treated with the acid- or alcohol-labeled preparations were generally very low except for the fat (ca. 1 ppm). In contrast, the CN-labeled preparation showed relatively high residue levels, especially in the stomach (contents), intestines, and skin (Crawford et aI., 1981a).
The major metabolic reactions (Fig. 58.6) of trans- and ciscypermethrin were cleavage of ester linkage, oxidation at the trans- and eis-methyl cyclopropane ring and at 4'-position of the phenoxy group, and conversion of the CN group to SCN ion. The following minor species differences were observed: (1) oxidation at 5- and 6-positions of the alcohol moiety was observed in mice but not in rats; (2) ester metabolites such as 2'-OH-, 5-0H-, and trans-OH,4'-OH-cypermethrin were detected in feces of mice but not of rats. The remarkable species difference in metabolites was the PBacid-taurine conjugate, which was the predominant metabolite in mice, but it was not detected in rats. The ester linkage of cis-cypermethrin seems to be more stable than that of the corresponding trans isomer, based on the nature of urinary and fecal metabolites and excretion rate (Crawford et aI., 1981b; Edwards et aI., 1990; Hutson and Casida, 1978; Hutson et aI., 1981). There are, additionally, species differences of conjugation reactions of the alcohol moiety in other species; PBacidglycine is predominant in sheep, cat, gerbil, and ferret; PBacidtaurine in ferret; PBacid-glycylvaline in mallard duck; and PBacid-glucuronide and/or 4' -OH-PBacid-glucuronide in hamster, guinea-pig, marmoset, and rabbit. The rat was unique in utilizing sulfuric acid for conjugation of 3-phenoxybenzyl moiety among animal species tested (Huckle et aI., 1981). Metabolism of cypermethrin (cis : trans = 1 : 1) in humans was investigated after oral administration to six male volunteers at 3.3 mg per person. The four metabolites from the acid and alcohol moieties were analyzed in urine. The amount of eis- and trans-ClzCA was approximately equal to that of PBacid and 4'-OH-PBacid. The ratio of trans- to cis-ChCA was on average 2 : 1, implying that ester hydrolysis is the major metabolic pathway and that the trans isomer was more rapidly hydrolyzed than the eis isomer, as is the case with rats. On the other hand, dermal application of cypermethrin (cis : trans = 56 : 44) led to formation of the different ratio of metabolites (the ratio of transto cis-ClzCA is 1 : 1.2) from oral administration (Woollen et aI., 1992).
58.7 CYPHENOTHRIN Chemical Name (RS)-a-Cyano-3-phenoxybenzyl (lR)-cistrans- 2,2-dimethyl-3-(2-methy lprop-l-enyl)cyclopropanecarboxylate. Synonyms Cyphenothrin (BSI, ISO) is the common name in use. The trade name is Gokilaht. Code designations include S-2703 Forte. The CAS registry number is 39515-40-7. Physical and Chemical Properties The empirical formula is C24H2SN03; molecular weight is 375.5. Its form is a viscous yellow liquid; its specific gravity is 1.08 at 25°C; log Kow = 6.2. It is less soluble «0.01 mg/l) in water at 25°C, but is soluble in most organic solvents. It is stable under normal storage conditions.
58.7 Cyphenothrin
Cl,
V
o
C~COOCH-cr°'O I I I CN
-'"
-'"
1
glucuronide ~ CI,_
V
glycine conjugate~ C~COOH
.
---"7 glucuronide
PBalc
Cypermethrin
I",,"" co'l'gol, "-------
0
HOH2C~O~
1269
Cl
i
~KcvoD]
CH 20H
C~COOCH-cr0 '0 I
CN
I
-'"
I
-'"
t
t-OH-qper
?glycine conjugate
~glucuronide HOOC~O~ ..
CI2CA
0_ .0
---"7taunne conjugate
PBacid
glucuronide
sulfate
CI~20H Cl
HOOC~O~
HO~
eOOH
V
6-{) H-f'Bacid
t-C H20 H-C 12eA
HOOC~O~
V
glucuronide sulfate
~OH
4'-{)H-f'Bacid
HOOC?,0'O
glucuronide
5-{) H-f'Bacid Figure 58.6 Metabolic pathways of cypennethrin in animals.
Metabolism Single oral or subcutaneous administration of 14C-trans- or cis-cyphenothrin labeled in the acid or alcohol moiety to rats at 2-4 mg/kg resulted in almost complete elimination of the 14C from the animal body. Major excretion routes with the acid- or alcohol- (except for the CN group) labeled preparation were the urine and feces. The total recovery of the 14C within 7 days after administration of these labeled preparations was more than 93% in urine and feces. On the other hand, the 14C derived from the CN group was more slowly excreted. In addition, 4-6% of the 14C was expired as 14C02. The total 14C recovery was 60-80% for the trans and cis isomers. The three labeled preparations of the trans and cis isomers showed more urinary excretion of the 14C with subcutaneous than with oral administration (Kaneko et al., 1984c). 14C tissue residue levels 7 days after single oral or subcutaneous administration of each of the 14C-Iabeled preparations of trans- and cis-cyphenothrin were measured. With the acid- and alcohol-Iabeled preparations of the trans and cis isomers, the tisssue residue levels were generally very low. On the other hand, the CN-Iabeled preparation showed relatively higher tissue residues than other labeled preparations. Both the trans and cis isomers underwent the following major metabolic reactions (Fig. 58.7): (1) oxidation at the 2/- and
4/ -phenoxy positions of the alcohol moiety; (2) oxidation at
the isobutenyl and the gem-dimethyl groups of the acid moiety; (3) cleavage of ester linkage; (4) conversion of the CN ion to SCN ion and C02; and (5) conjugation of the resulting carboxylic acids and phenols with glucuronic acid, sulfuric acid, and glycine. In vivo and in vitro comparative metabolism studies of phenothrin and cyphenothrin showed the following results: (1) The trans isomers of cyphenothrin and phenothrin were hydrolyzed more rapidly in vitro (liver homogenates) and in vivo than the corresponding cis isomers, and cis-cyphenothrin was hydrolyzed to a larger extent than cis-phenothrin. (2) Plasma esterases showed a different substrate specificity from the liver esterases and hydrolyzed the trans and cis isomers of cyphenothrin and phenothrin to nearly the same extents. From the results of the in vivo and in vitro studies, the CN group introduced into the molecule did not affect the biodegradability of trans-cyphenothrin, but rather made cis-cyphenothrin more biodegradable than cis-phenothrin. These in vivo metabolic profiles (ester hydrolysis rate, excretion pattern into urine and feces) may be mainly determined by activity and/or substrate specificity of the liver esterases (Kaneko et al., 1984c).
1270
CHAPTER 58
Pyrethroid Chemistry and Metabolism
0COOH H000
HOO~
COOH wc-acid-c-CA
HOOC wt-acid-c-CA
t
. COOH wc-acld-t-CA
i
r!:; c-CA
COOH
COOH
COOH
wt-Dlc-t-CA
~ g'","m";d, ~~! t-CA
eOOH
~glucuronide
~
0COO\H()o' O / [H01:a°'O]
,--.L.L_----''-------,
i
wc-alc-t--{;A
wc-alc-c-CA
CO ° H
HOH2~
HOH2~
COOH
~
wt-Dcid-t-CA
i
0COOH [OH'YS i J
HOH 2C wt-Dlc-c-CA
HOOC~
COOCH'U0'O
I
I
. CN 0.. I 0.. I ""lc1>l.;is,---C",;vu'ID"'lh""en-'-"o'-"th.! . r! !.ln_ _ _ _ _ _ _.....J
trans-C henothrinCN
Ester Metabolites
0..
I
0..
[HOI:U"~H ~
t'~--7)[CWJ<E-E---~~ Rj
R2
R3
1) COOH CH 3
H COOH H 2) CH3 3) CH 20H CH 3 H 4) CH 3 CH20 H H COOH H 5) CH3 COOH OH 6) CH3 Figure 58.7
R, CH3 CH3 CH 3 CH3 CH20 H CH3
OHCa°'O
~
si! ~2
~
PBald
HOOC'U0'O glycine conjugate J
[OHC'U°'O-OH]
PBacid
\9lucuronide
)
HOOC~O~OH
sulfate/ 2'-or
4'--DH--PBaci~ glucuronide
Metabolic pathways of cyphenothrin in animals.
58.8 DELTAMETHRIN Chemical Name (S)-a-Cyano-3-phenoxybenzyl (lR)-cis-3(2,2-dibromovinyl)-2,2-dimethylcyclopropanecarboxylate. Synonyms Deltamethrin (BSI, ISO) is the common name in use. Trade names are Decis, Butox, K-Othrine, Kordon, and Sadethrin. Code designations include NRDC 161, AEF 032640, and RU 22974. The CAS registry number is 52918-63-5. Physical and Chemical Properties The empirical formula is C22H19Br2N03; molecular weight is 505.2. Its form is a colorless crystal; its specific gravity (bulk density) 0.55 at 25°C; log Kow = 4.6. It is less soluble « 0.0002 mg/l) in water at 25°C, but is soluble in most organic solvents. It is stable to air and in acidic conditions, but rather unstable in alkaline medium. Metabolism On oral administration to rats at 0.60-1.64 mg/kg, the acid and alcohol moieties of deltamethrin were almost completely eliminated from the body within 2-4 days. On the other hand, the CN group was eliminated more slowly than
the acid and alcohol moieties, the total recovery during 8 days being 79% of the dosed radiocarbon (43 and 36% in the urine and feces, respectively). 14C tissue residues with deltamethrin preparation labeled in the acid moiety or in the alcohol moiety were generally very low whereas the fat showed somewhat higher residue levels (0.1-0.2 ppm). The 14C derived from the CN group showed relatively high residue levels, especially in skin and stomach. Essentially all the 14C in the stomach was SCN ion (Ruzo et al., 1978). The major metabolic reactions (Fig. 58.8) of deltamethrin in rats are oxidation at the trans-methyl relative to the carbonyl group of the acid moiety and 2'-, 4'-, and 5-positions of the alcohol moiety, cleavage of ester linkage, conversion of the CN group to SCN ion and 2-iminothiazolidine-4-carboxylic acid (!TCA), and conjugation of these carboxylic acid and phenol derivatives with sulfuric acid, glycine, and/or glucuronic acid. Although the major metabolic pathways in mice are similar to those in rats, there are the following species differences in metabolism: (1) amino acid conjugation reactions of the alcohol moiety (rat, glycine; mouse, taurine); (2) rats produced more phenolic metabolites than mice; and (3) mice produce trans-
58.9 Empenthrin
y,
8~OO\f?O~ V
, HOOC~O~ glucuronide ~
V
Br
5-{)H-PBacid HOOC
sulfale~ glucuronide
sulfale
t//
~ [
scw
8;~OOCIH,;yO~
4'-{)H-PBacid
Br
0
~ OH
0
0
'6
I
NH::/l ~
/.
-
CDOH
~
9lUCU,ronide glycme conjugate
c-B r2CA
~ OH
t
Br~cOOCH
~ [cw] ~ rOHCo0'O~ II I CN
V
CN 4'-{)H-iJec
OH]
2'-{)H-PBacid
Br
H
~ ~
V
HOOC
5-{) H-iJec
O
Br
CN I /.
/
1271
Br
Dellamethrin
~NO /.
0
'0 /.
/.
t
OHC~O~
COOH
V
ITCA
V
~sulfate
~ glucuronide
PBald
t
taurine conjugate glucuronide
HOOCO°'O
glyc,ine conjugate
PBacid
glucuronide
~
HOH2C~O~
0 0 PBalc
Br glucuronide
~
HOH2C~O~
0
~
4'-{) H-P Bale Figure 58.8
X
OH
~COOqH~O~
Br
0H
CNV V
2'-{) H-iJec
Metabolic pathways of deltamethrin in animals,
hydroxy methyl cyclopropanecarboxylic acid to a larger extent than rats (Ruzo et aI., 1978, 1979). A human metabolism study of deltamethrin was carried out in three volunteers after a single oral dose of 3 mg of 14C_ deltamethrin per person. The 14C was more rapidly excreted into urine (51-59%) than into feces (10-26%), total excretion of 14C being 64-77% of the dose for 96 hr (IPCS, 1990b).
58.9 EMPENTHRIN Chemical Name (E)-(RS)-1-Ethynyl-2-methylpent-2-enyl (lRS)-cis-trans- 2,2-dimethyl-3-(2-methylprop-1-enyl)cyclopropanecarboxylate. Synonyms Empenthrin (BSI, ISO) is the common name in use. The trade name is Vaporthrin. Code designations include S-2852 Forte. The CAS registry number is 54406-48-3.
Physical and Chemical Properties The empirical formula is ClSH2602; molecular weight is 274.4. Its form is a yellow liquid; its specific gravity is 0.927 at 20°C; log Kow = 5(est). It is less soluble (0.111 mg/l) in water at 25°C, but is soluble in most organic solvents. It is stable under normal conditions. Metabolism When single oral administration of (lR)-cis- or (lR)-trans-empenthrin labeled with 14C in the alcohol moiety was given to rats at 3-600 mg/kg (female), 97-106% of the dosed 14C was rapidly eliminated into urine and feces within 7 days after administration. Monitoring of the expired air indicated that less than 1.1 % of the dose was excreted as 14C02. Urinary, fecal, and exhaled 14C excretion accounted for 22-41, 60-74, and 1-2% of the dose, respectively (Isobe et aI., 1992). 14C tissue levels reached maxima at 1-8 hr after administration of the cis or trans isomer at 3 mg/kg and decreased thereafter. Liver and kidney tissues showed higher 14C concentrations than other tissues. No notable sex-related difference
1272
CHAPTER 58
Pyrethroid Chemistry and Metabolism
HO~ OH
2-oxo-EM PA
1
glucuronide
6-DH-EMPA
1
glucuronide
1
glucuronide
Figure 58.9 Metabolic pathways of empenthrin in animals.
was observed in distribution or excretion of the radioactivity. 14C tissue residues were lower in rats receiving the trans isomer than in those receiving the cis isomer. The parent compound accounted for 7-13 and 17-26% of the dose in the feces of rats receiving the cis and (rans isomers, respectively. The major metabolites were 1-ethynyl-2methylpent-2-enol (EMPA), 6-0H-EMPA, and 2-oxo-EMPA and their glucuronides. An ester-retaining metabolite by Emethyl hydroxylation at the isobutenyl group in the acid moiety was also found. Major metabolic reactions (Fig. 58.9) in rats were cleavage of the ester linkage and glucuronide formation of the resulting alcohol derivatives; as a minor pathway, oxidation of the methylene group in the alcohol moiety, and hydration of the triple bond in the alcohol moiety were found. In addition, the oxidation at the methyl group of the isobuteny1group occurred (lsobe et a!., 1992).
58.10 ETOFENPROX Chemical Name 3-phenoxybenzyl ether.
2-(4-Ethoxyphenyl)-2-methylpropyl
Synonyms Etofenprox (ISO, BSI, INN) is the common name in use. Code designations include MTI-500. The trade name is Trebon. The CAS registry number is 80844-07-01. Physical and Chemical Properties The empirical formula is C2SH2S03; molecular weight is 376.5. Its form is a white crystal; its specific gravity is 1.157 at 23°C; log Kow = 7.05. It is less soluble «0.001 mg/l ) in water at 25°C, but is soluble in most organic solvents. It is stable to light and in acidic and alkaline medium. Metabolism The metabolism of etofenprox has been studied in rats and dogs. On single oral administration of l4C-etofenprox (a 1 : 1 mixture of [1_ l4 C-propyl]-etofenprox and [a- 14 C-benzyl]-etofenprox to both sexes of rats at 30 or 180 mg/kg, l4C excretion
rates in feces and urine were 87-90 and 7-9%, respectively, of the dosed 14C 5 days after administration. No 14C was found in expired air. Plasma 14C reached peak levels after 3-5 hr. 14C excreted into bile was found to account for 10-30% and unchanged etofenprox was not found in the bile. l4C tissue concentration was the highest in fat, as unchanged parent compound. When 14C-etofenprox was administered to pregnant rats (gestation day 10 to day 16) at 30 mg/kg/day, the l4C was transferred to the fetus through the placenta; however, their levels were low compared to other tissues of mother animals. Etofenprox was secreted in milk as the unchanged compound (FAOIWHO,1993).
The major fecal metabolites were deethylated etofenprox (DE) and 4'-OH-etofenprox (4'OH). The major biotransformation routes (Fig. 58.10) involve O-deethylation of the ethoxyphenyl moiety and hydroxylation of the phenoxybenzyl moiety followed by conjugation with glucuronide or sulfate. Oxidation of the a-CH2 group followed by hydrolysis represents an additional route. On single oral administration of 14C-etofenprox (a 1 : 1 mixture of [1_ l4 C-propyl]-etofenprox and [a- 14 C-benzyl]-etofenprox to beagle dogs of each sex at 30 mg/kg, total excretion in feces was 90 and 6% of the dosed 14C in urine over 5 days after administration. Based on the results of metabolites in excreta, the total estimated gastrointestinal absorption was 14-51 %. Tissue concentrations were highest in liver. The results indicate a lower gastrointestinal absorption rate in dogs than in rats. The major biotransformation routes were the same as in rats (FAOIWHO, 1993).
58.11 FENPROPATHRIN Chemical Name (RS)-a-Cyano-3-phenoxybenzyl 2,2,3,3tetramethylcyclopropanecarboxylate. Synonyms Fenpropathrin (BSI, ISO, ANSI) is the common name in use. Trade names are Rody, Danitol, Meothrin, Ortho, and Danitol. Code designations include S-3206. The CAS registry number is 64257-84-7. Physical and Chemical Properties The empirical formula is C22H23N03; molecular weight is 349.4. Its form is a yellowbrown solid; its specific gravity is 1.15 at 25°C; log Kow = 6. It is less soluble (0.0141 mg/l) in water at 25°C, but is soluble in most organic solvents. It is stable to light, but is unstable in alkaline medium. Metabolism Single oral administration of 14C-acid- and 14C_ alcohol-labeled-fenpropathrin preparations to rats at 2.4-26.8 mg/kg resulted in almost complete elimination of the 14C from the animal body within 7 days. Major excretion routes of 14C_ acid and 14C-alcohol preparations were the urine and feces. The 14C recoveries with the acid and alcohol preparations were 96102% (urine, 27-44%; feces, 58-70%) and 96-98% (urine, 2643%; feces, 54-71%), respectively. The 14C excretion pattern
58.12 Fenvalerate (Esfenva1erate)
1273
a-CO
~H'OO'tIl~COOO~J
L
DE 4'OH
1 o
HOOC~ ~
V
~
~OH
glucuronide or sulfate Figure 58.10
Metabolic pathways of etofenprox in animals.
into the urine and feces was very similar between both labeled preparations (Kaneko et ai., 1987). With 14C-acid and -alcohol preparations, the tissue residue levels 7 days after single oral administration of each of the 14C-Iabeled fenpropathrin preparations were generally very low. However, the fat showed slightly higher residue level for both labeled preparations compared with other tissues (Kaneko et al., 1987). Fenpropathrin was rapidly metabolized in rats via cleavage of the ester linkage, oxidation at both or one methyl group of the acid moiety and at the 4' -position of the alcohol moiety, and conjugation with sulfuric acid, glucuronic acid, and amino acid (Fig. 58.11). The CN group is presumed to undergo the same metabolic reaction, conversion of the CN group to SCN ion, as those observed in pyrethroid insecticides having a-cyano3-phenoxybenzyl alcohol such as cyphenothrin, deltamethrin, and fenvalerate. The major urinary metabolites were sulfate of 4' -OH-PBacid from the alcohol moiety and free and glucuronides of 2,2,3,3-tetramethylcyclopropanecarboxylic acid (TMPA) and its hydroxymethyl derivatives from the acid moiety. The major fecal metabolites retained ester linkage with oxidation at the 4' -position of the alcohol and the trans-methyl group (Crawford and Hutson, 1977; Kaneko et al., 1987).
58.12 FENVALERATE (ESFENVALERATE) Chemical Name (RS)-a-cyano-3-phenoxybenzyl (RS)-2-( 4chlorophenyl)-3-methylbutyrate; esfenvalerate is (S)-a-cyano3-phenoxybenzy (S)-2-(4-chloropheny1)-3-methylbutyrate.
Synonyms Fenvalerate (BSI, ISO, ESA) is the common name in use. Trade names are Sumicidin, Pydrin, and several other names. Code designations include S-5602 and WL43775. Esfenvalerate (BSI, ISO) is an insecticidally active isomer of four isomers of fenvalerate and is the common name in use. Trade names are Sumi-alpha and Asana. Code designations include S5602Aa, DPX-YB656, and S-1844. The CAS registry numbers are 51630-58-1 for fenvalerate and 66230-04-4 for esfenvalerate. Physical and Chemical Properties (Fenvalerate) The empirical formula is C2sH22CIN03; molecular weight is 419.9. Its form is a viscous yellow or brown liquid and sometimes partly crystalline at room temperature; its specific gravity is 1.17 at 25°C; log Kow = 6.2. It is less soluble « 0.01 mg/l) in water at 25°C, but is readily soluble in most organic solvents. It is relatively stable in acidic media, but unstable in alkaline medium. Metabolism Single oral administration of the 14C preparations of fenvalerate and its (2S) isomer labeled in the acid or alcohol moiety to both sexes of rats and mice at 6.7-8.4 mg/kg resulted in almost complete elimination of the 14C from the animal body. Major excretion routes for the 14C in both animal species were the urine and feces. The total recovery of the 14C 6 or 7 days after administration was 93-102% in rats and mice. In contrast, the 14C from the 14CN-preparation of fenvalerate and its (2S) isomer was more slowly excreted than other 14C preparations, and mainly into the urine and feces. Additionally approximately 6-14% of the 14C was expired as 14C02 in the two species. The total recovery of the 14C was 75-81 % in
1274
CHAPTER 58
glucuronide
Pyrethroid Chemistry and Metabolism
<E:--A
A
E
COOqH~O~ CN I I
COOH TM PA
IloA
J
HOH 2C
::-....
::-....
~
qlucuronide <E:--~
-
Fenpropathrin
t
_
)A
COOqH~ O~ CN I
HOA
::-....
~
HA
HA
V
OH
0Y'i1
V
OH
~ I
I
::-....
::-....
COOH-Fenp.
«------
p.
4'-DH,CH OH-fenp. 2i
CH20 H-F enp.
CN
~
l C;~~ HOCH I CN
:::-....
::-....
COO~H~O~
COOH TMPA---COOH
4'r-fec COOqH~ CN I
COOGH~O~ CN I I
COOH TMPA ---CH 20 H
::-....
~
::-....
0
I
~
::-....
l
HO\~C I CN
::-....
~
::-....
I
OHCa°X) PBald
qlycine coniugate Figure 58.11
<E:--
HOOC~O~ 0~..0 PB acid
Metabolic pathways of fenpropathrin in animals.
rats and 88-89% in mice (Kaneko et aI., 1981a; Ohkawa et aI., 1979). Single oral administration of 14C-acid- and 14C-alcohol (ring)-fenvalerate to beagle dogs resulted in rapid 14C elimination from the animal bodies. Major l4C excretion routes were the urine and feces. The l4C recovery was 87% (55.5 and 31.6% in the feces and urine) and 79% (42.3 and 36.8% in the feces and urine) 3 days after oral administration of the acidand alcohol-labeled preparations, respectively (Kaneko et aI., 1984a). 14C tissue residue levels 6 or 7 days after administration of 14C-labeled preparations to both sexes of rats and mice were determined. With the preparations of fenvalerate and its (2S) isomer labeled in the acid and alcohol moieties except for the CN group, the residue level in the fat was relatively higher in rats and mice, whereas the residue levels in other tissues, including blood, hair, liver, kidney, and skin, were low. However, administration of the CN-labeled preparations resulted in somewhat higher tissue residues, in general, compared with other labeled preparations. Higher residues were especially found in the hair, skin, stomach, blood, and fat, and it was found
that most of these residues was due to retention of SCN ion. The 14C levels in these tissues were lower in mice than in rats. Fenvalerate underwent the following major metabolic reactions (Fig. 61.12); hydroxylation at 4'-phenoxy position of the alcohol moiety and C-2 and C-3 positions of the acid moiety, cleavage of the ester linkage, conversion of the CN group to SCN ion and C02, and conjugation of the resulting carboxylic acids, phenols, and alcohols with glucuronic acid, sulfuric acid, and/or glycine. With the alcohol moiety, the following apparent species differences were observed between dogs and rodents such as rats and mice: (1) hydroxylation at both the 2'- and 4' -positions of the alcohol moiety occurred in rats and mice, but only at the 4'-position in dogs; (2) PBalc and 4'-OH-PBalc from the alcohol moiety were obtained from dogs to a considerable extent, but were not detected in rats or mice; (3) PBacid-glycine was found to a larger extent in dogs than in rats or mice. There were also the following remarkable species differences, particularly in the major conjugates of the alcohol moiety: PBacid-glycine was predominant in dogs, 4' -OH-PBacid-sulfate in rats, and
58.13 Flucythrinate
Cl
-O( -
~ /,
cl
CH 2
'/ \
,0
CO CI--Bacid-lactone
\
coO Cl ~ /, H 2,3-{)H-CP A-lactane
9
Cl ~ /, coO 3-{)H-CPIA-lactone
t
t
Cl
~CH2 \
CH2
-
t
-0(
CH OH clCH20H ~CH OH I 2 .;--- Cl ';---CI 2.;--- Cl ~ /, COOH OH COOH ~ /, COOH ~ /, CI--Bacid 2,3-{)H-CPIA 3-{)H-CPIA ~ CPIA
-O(
~-/,
conJgate
Cl
1275
Y
Cl
------i>glucuronide
COOH
Y~
~COOCH-a0'O ~COOCH-a°'00H I I I I t=\
t=\
I~
CN
'"
'"
CN
'"
'"
2'- or 4'-{) H-F envalerate
.j,
Cl
-O( -
~ /,
CO
[
! \
,0
CO CI--B Dacid-()nhydride
HO~~O°'O
]
[CN-J~'----...L./---;j;d;------~/
sc~t
10 2
OHCa°J:]
j PBalc
.j,
[OHCO°'00Hl~ 'j
PBald
HOOC-a0'O
taurine conjugate
t.
fYO~0H V 4'-{)H-f'Balc ~OH HOOC
0 0
2'- or 4'-{) H-f'Bacid
PB acid
~
.j,
HOCHn'?[rO~
.j,
.j,
HOCH2-a°'O
Figure 58.12
[HO~~O°'O°HJ
glYCine conjugate
~
glucuronide
~
sulfate
~
glucuronide
Metabolic pathways of fenvalerate in animals.
PBacid-taurine in mice (Kaneko et aI., 1981a; Ohkawa et aI., 1979). A comparative metabolism study of the four optical isomers of fenvalerate was carried out. The 14C-labeled preparations of the four isomers labeled in the acid moiety were administered to rats and mice, out of the four isomers, only the (2R, as) isomer produced cholesterol ester conjugate, which is an ester of the acid moiety 2-(4-chlorophenyl)isovaleric acid (CPIA) of the (2R, as) isomer and cholesterol. This metabolite was found in relatively larger amounts in spleen, lymph node, adrenal, and liver tissues than other tissues. This conjugate was demonstrated to be formed by transesterification reaction, not by any of three known pathways of cholesterol ester biosynthesis (acylcoA: cholesterol o-acyltransferase (ACAT), lecithin: cholesterol o-acyltransferase (LCAT), and cholesterol esterase), and to be a causative agent of granulomatous changes which were caused by long-term or subacute administration of fenvalerate, but not by esfenvalerate (Kaneko et aI., 1986a, 1988; Miyamoto et aI., 1986; Okuno et aI., 1986). In addition, a comparative metabolism study of fenvalerate and esfenvalerate was carried out and the results showed that there was no significant differences in metabolism between fenvalerate and esfenvalerate except for formation of a cholesterol ester conjugate from fenvalerate and that the other three isomers of fenvalerate did not seem to affect the absorption, excretion, distribution (including placentral transfer), and biotransformation of esfenvalerate (Isobe et aI., 1990; Shiba et aI., 1990).
58.13 FLUCYTHRINATE Chemical Name (RS)-a-Cyano-3-phenoxybenzyl (S)-2-(4diftuoromethoxyphenyl)-3-methylbutyrate. Synonyms Flucythrinate (BSI, ISO, ANSI) is the common name in use. Trade names are Cybolt, Cythrin, Fluent, and PayOff. Code designations include AC 222705 and CL222705. The CAS registry number is 70124-77-5. Physical and Chemical Properties The empirical formula is C26H23F2N04; molecular weight is 451.4. Its form is a dark amber viscous liquid; its specific gravity is 1.189 at 22°C; log Kow = 6.8. It is less soluble (0.5 mg/l) in water at 21°C, but is soluble in most organic solvents. It is unstable in alkaline conditions. Metabolism Flucythrinate is the same as fenvalerate in terms of chemical structure except for substitution of the benzene ring of the acid moiety: chlorine atom for fenvalerate and the diftuoromethoxy group for ftucythrinate. When 14C-preparations of ftucythrinate labeled in the acid or alcohol moiety were administered to rats at 19.7 mg/kg, 14C was excreted into urine (20-30%) and feces (70-73%). The major metabolic reactions (Fig. 58.13) are cleavage of ester linkage and oxidation at the gem-dimethyl groups of the acid moiety and the 4' -position of the alcohol moi-
1276
CHAPTER 58
Pyrethroid Chemistry and Metabolism
CFHO 2
~Y
~COOSH-al ~
CN
0
'0 ~
1
~Y
CF2HO~COOCH:b)0'O -----?>
Fluc thrinate
~
1
metabolites from the alcohol moiety ( please refer to fenvalerate)
CF2HO~COOH
~Y
1
2
~Y
HO~COOH
I
CON ~
/1~ r ~yCHzOHJ [F2HO~COOH
nY
HO CF2 ~C 0 NHC H2C 00 H
1
CF2H 0 Figure 58.13
-ox 'I
-
'\
CH2
co
::::0
Metabolic pathways of f1ucythrinate in animals.
ety. In addition, de-difluoromethylation takes place. The major metabolites from the alcohol moiety are the same as those from fenvalerate and the major metabolite from the acid moiety is 2-(4-difluoromethoxyphenyl)-3-methyIbutyric acid (FAOIWHO, 1985).
4-fluoro-3-phenoxybenzoie acid (FPBacid), 4' -OH-FPBacid, and their glycine conjugates from the alcohol moiety. The major fecal metabolites were unchanged flumethrin and flumethrin acid. The major metabolic reactions (Fig. 61.14) of flumethrin are ester hydrolysis and oxidation of the 4' -position of the alcohol moiety (FAOIWHO, 1996).
58.14 FLUMETHRIN Chemical
Name
(RS)-a-Cyano-4-fluoro-3-phenoxybenzyl
58.15 r-FLUVALINATE (FLUVALINATE)
3-(fJ ,4-dichlorostyry1)-2,2-dimethylcyclopropanecarboxylate.
Synonyms Flumethrin (BAN) is the common name in use. Trade names are Bayticol and Bayvarol. Code designations include BAY VI6045, and BAY Vq1950. The CAS registry number is 69770-45-2. Physical and Chemical Properties The empirical formula is C28H22ChFN03; molecular weight is 510.4. Its form is a yellowish and highly viscous oil; log Kow = 6.2. It is less soluble (0.001 mg/l) in water at 20°C, but is soluble in most organic solvents. Metabolism When l4C-flumethrin labeled in the acid moiety was administered to rats at 1-5 mg/kg, absorption was rapid, but incomplete. The maximum 14C levels in plasma were obtained in about 8 hr after administration. The major urinary metabo1ites are 3-(2-chloro-2-(4-chloropheny I)etheny1)-2,2-dimethylcyclopropanecarboxy lie acid (flumethrin acid) and its glucuronide from the acid moiety, and
Chemical Name (RS)-a-Cyano-3-phenoxybenzyl N-(2chloro-a ,a ,a-trifluoro- p-toly 1)-D-valine; fluvalinate is (RS)-acyano-3-phenoxybenzy N -(2-chloro-a,a ,a-trifluoro-p-tolyl)DL-valine. Synonyms r-Fluvalinate (BSI, ISO) is the common name in use. Trade names are Mavrik and Klartan. Code designations include SAN5271. The CAS registry numbers are 102851-06-9 for r-fluvalinate and 69409-94-5 for fluvalinate. Physical and Chemical Properties The empirical formula is C26H22CIF3N203; molecular weight is 502.93. Its form is a viscous amber oil; its specific gravity is 1.26 at 25°C; log Kow = 4.26 or 6.4. It is less soluble «0.001 mg/l) in water at 25°C, but is soluble in most organic solvents. It is unstable to light and in alkaline medium. Metabolism The metabolism of fluvalinate has been extensively studied with the 14C-preparation labeled in the acid moi-
58.16 Imiprothrin
1277
Flumethrin
/ CI-Q-\=CHACOOf Cl Flumethrin acid
glucuronide
HOOC~°'O FPBacid
HOOC~O~ F 4'-DH-FPBacid
Figure 58.14
:>
F
:> OH
Metabolic pathways of ftumethrin in animals.
ety. When fluvalinate labeled with 14C in the acid moiety was administered to rats at 1 mglkg, 14C was rapidly excreted into urine (9-19%) and feces (75-88%) within 4 days after administration. The major 14C component (45% of the fecal 14 C) was the parent compound. Liver showed relatively higher 14C tissue residue than other tissues, indicating that the 14C tissue residues from fluvalinate are somewhat different from those of other pyrethroids having the same alcohol moiety. The major metabolic pathways (Fig. 58.15) are cleavage of ester linkage and oxidation at the acid and alcohol moieties. Ester hydrolysis leads to formation of anilino acid. The anilino acid is further conjugated with amino acids (glycine, serine, threonine, and valine), bile acids (cholic, taurochoric, and taurochenodeoxycholic), and glycerols (oleoyl- and linoleoylglycerol). In addition, an amide derivative of anilino acid was found. These conjugation reactions of anilino acid with bile acid (cholic acid, taurocholic acid, and taurochenodeoxycholic acid) and with glycerol and monoglycerides have rarely been reported as conjugates with xenobiotics (Quistad et aI., 1982, 1983). The major urinary metabolites are ani lino acid, its hydroxymethyl derivative, its glycine conjugate, ha10aniline, and sulfate conjugate of hydroxyhaloaniline. On the other hand, the major fecal metabolites are anilino acid, its amide derivative, and several conjugates of anilino acid with several endogenous components. An unexpected difference between fluvalinate and other pyrethroids is the minimal amount of hydroxylation at the 4'-position of the alcohol moiety. Thus 4' -hydroxy-
fluvalinate is a very minor fecal metabolite, whereas the amount of 4' -hydroxylated parent compound is significant for other pyrethroids, that is, 4'-hydroxy-deltamethrin and 4' -hydroxycypermethrin (Quistad et aI., 1982, 1983). When 14C-fluvalinate labeled in the alcohol moiety was orally administered to rats, the alcohol moiety showed metabolic fates similar to those of pyrethroids having a-cyano-3phenoxybenzyl alcohol (Staiger and Quistad, 1984). When 14C-acid-fluvalinate was administered to rhesus monkeys at 1 mglkg, the 14C was excreted into urine (37%) and feces (55%) within 5 days after administration. The major metabolites found were anilino acid as its hydroxymethyl derivatives and glucuronide. There are several species differences between rats and monkeys: (1) glucuronide conjugation of anilino acid is a major metabolic pathway in rhesus monkeys, whereas little or no glucuronides are detected in rats; (2) conjugation with bile acids is a significant reaction in rats, but it is only a very minor process in rhesus monkeys; and (3) conjugation with glycerol and monoglycerides occurs in rats, but is not detected in rhesus monkeys (Quistad and Selim, 1983).
58.16 IMIPROTHRIN Chemical Name 2,5-Dioxo-3-(2-prop-2-ynyl)irnidazolidin1-ylmethyl (1 R)-cis,trans-2,2-dimethyl-3-(2-methy lprop-leny I)cyclopropanecarboxy late.
1278
CHAPTER 58 Pyrethroid Chemistry and Metabolism
CF~HN~OOCH-V°D l"=/. Cl
CN
I1
1
~
~
Fluvalinate
/~
[Ho~~-V°DJ 1
refer to fenvalerate metabolism
CFG-HN:X:OOII Cl
--------,----3>
Anilino acid
1
1
amide derivative amino acid conjugate bile acid conjugate glyceroles glucuronide
CF~NH 3"=/. 2
Cl 1OH
CF~NH ~2
--~)
sulfate
Cl
Figure 58.15
Metabolic pathways of fluvalinate in animals.
Synonyms Imiprothrin (BSI, ISO) is the common name in use. Code designations include S-4056F and S-41311. The trade name is Pralle. The CAS registry number is 72963-72-5. Physical and Chemical Properties The empirical formula is C17H22N204; molecular weight is 318.4. Its form is a viscous liquid; its specific gravity is 1.1 at 20°C; log Kow = 2.9. It is less soluble (93.5 mg/l) in water at 25°C, but is soluble in most organic solvents. Metabolism When l4C-(IR)-trans- or (IR)-cis-imiprothrin labeled in the alcohol moiety was administered orally to rats at 1 or 200 mg/kg, the 14C was rapidly and almost completely eliminated from rats within 7 days after administration (98-103% of the dosed 14C). The urinary 14C-excretion was 83-97%, whereas the fecal excretion was 16% or less. The urinary excretion of l4C for the trans isomer was slightly larger (89-97%) than that for the cis isomer (83-91%). 14C_ excretion into the expired air was less than 3%. 14C tissue residues on the 7th day after administration were generally low in all of the dosed group. There were no marked sexrelated differences in the rate of 14C-excretion and the 14C_ tissue residues between either treatment group (Saito et aI., 1996). The major metabolic reactions (Fig. 58.16) of trans- and cis-imiprothrin in rats are (1) cleavage of the ester linkage,
(2) cleavage of the imidomethylene linkage, (3) hydroxylation of the imidazolidine ring, (4) dealkylation of the 2-propynyl group, and (5) oxidation at the trans-methyl group in the isobutenyl side chain (Saito et aI., 1995). The major urinary metabolites were 2,4-dioxo-l-(2-propynyl)-imidazolidine (PGH), PGH-OH, and hydantoin (HYD) from the trans isomer and these ester-cleaved metabolites and metabolites with intact ester linkage from the cis isomer.
58.17 KADETHRIN (RUI5525) Chemical Name 5-benzyl-3-furylmethyl (E)-(IR)-cis-2,2dimethyl-3-(2-oxothiolan-3-ylidenemethyl) cyclopropanecarboxylate. Synonyms Kadethrin (kadethrine) is the common name in use and is also the trade name. Code designations include RU 15525. The CAS registry number is 58769-20-3. Physical and Chemical Properties The empirical formula is C23H2404S; molecular weight is 396.5. Its form is a yellowbrown viscous oil; log Kow = 5.4. It is practically insoluble in water, but is soluble in most organic solvents. It is unstable to light and in alkaline medium.
58.18 Pennethrin
~COOCH~~"'4] ----7lHOCH2~~] o
l
0
o
Hp~ o
/
o
OH
HOOC~COOCH2P~ o
OH
PGH-DH
W
i
H~~
OH
tilcid-ct-PGH-O H
o
HOOC,
o
V COOCH
f=\.fY
2
NAN~ );-1
o
o PGH
~c OOCH'~N "'4
[~COOCH2~~1 / o
OH
OH
1279
W
-+---'»
t-acid-ci;-PGH
rOCH:~ "'4]
<E---+--
o trons -Imiprothrin
ci;-Imiprothrin
HYD
~ Figure 58.16
/
Metabolic pathways of imiprothrin in animals.
Metabolism The acid moiety of kadethrin is unique in terms of chemical structure and the alcohol moiety is the same as that of resmethrin. The metabolism of kadethrin has been studied after oral administration of 14C-labeled preparations in the acid or alcohol moiety to rats. The 14C from the acid and alcohol moieties was rapidly and completely excreted into urine and feces (Ohsawa and Casida, 1980). The major metabolic reactions (Fig. 58.17) are cleavage of the ester linkage and oxidation at the acid and alcohol moieties. The alcohol moiety after ester hydrolysis is further metabolized as shown in the studies with resmethrin (Miyamoto et aI., 1971). The acid moiety initially undergoes hydrolysis of the thiolactone ring, and the resulting mercaptan derivative is oxidized directly to a sulfonic acid or is first methylated before oxidation to a methyl sulfoxide and then a methyl sulfone. Hydroxylation reactions and thiolactone hydrolysis also occur on the intact molecule (Ohsawa and Casida, 1980).
58.18 PERMETHRIN Chemical Name 3-Phenoxybenzyl (IRS)-cis-trans-3-(2,2dichloroviny1)-2,2-dimethylcyclopropanecarboxy late. Synonyms Permethrin (BSI, ISO, ANSI, ESA, BAN) is the common name in use. There are many trade names, such as Adion, Ambush, Assithrin, Cliper, Coopex, Corsair, Dragnet, Dragon, and Eksmin. Code designations include S-3l5l, pp 557, FMC 33297, NRDC143, WL43479, and LE79-519. The CAS registry number is 52645-53-1. Physical and Chemical Properties The empirical formula is C21H20Ch03; molecular weight is 391.3. Its form is a pale yellow-brown liquid; its specific gravity is 1.19-1.27 at 20 D C; log Kow = 6.1. It is less soluble (ca. 0.2 mg/l) in water at 20 D C, but is soluble in most organic solvents. It is more stable in acid medium than in alkaline medium.
1280
CHAPTER 58
Pyrethroid Chemistry and Metabolism
"OY0«-HOHt)'-.2
1V
§TO" \
X s--\o l-~ ~
I
I
~COOCH,
~
0 /-~" Y. i-r---7 S--\
l../
iJIO"
-COOCH,
\ I
I
~ OH
Kadethrin (refer to resmethrin metabolism)
1 qlucuronide ~ glycine conjugate
«--
o)i
LJ='~ 'to 0 H
S--\
1
lHcFL~coojl IHOOC
o HOOC
CH";;~
l
HOOC
L
'COOH~ CH;;~ 'toOH
Figure 58.17
1 o HOOC
6.
Ho-S~
l---?
'toOH
Metabolic pathways of kadethrin in animals.
Metabolism The metabolism of permethrin has been studied in detail in a wide variety of animals in vivo and in vitro (rats, cows, hens, and goats). However, the in vivo metabolism data relating to mammalian toxicology are limited to rats. When the four 14C-preparations of (lRS)-trans-, (lR)-trans-, (lRS)-cis, and (lR)-cis-permethrin labeled in the alcohol and acid moieties were administered orally to male rats at 1.64.S mg/kg, the compounds were rapidly metabolized and the 14C from the acid and alcohol moiety was almost completely eliminated from the body within a few days. The l4C from the cis isomer was excreted into the urine and feces almost equally, whereas more than SO% of the dosed 14C from the trans isomer appeared in the urine. The 14C tissue residues were very low although the fat with the cis isomer showed relatively higher residue levels (Elliott et a!., 1976; Gaughan et a!., 1977). The major metabolic reactions (Fig. 5S.1S) of both permethrin isomers were oxidation at the trans and cis positions of the gem-dimethyl group of the acid moiety and at the 2'- and 4'-positions of the alcohol moiety, ester cleavage, and the conjugation of the resulting carboxylic acids, alcohols, and phenols with glucuronic acid, glycine, and sulfuric acid. The cis isomer is more stable than the trans isomer, and the cis isomer yielded four fecal ester metabolites which resulted from hydroxylation at the 2'- and 4'-positions of the phenoxy group, at the trans-methyl group, and at both of the two latter sites. The ester-cleaved metabolites were extensively excreted into the urine, whereas the metabolites retaining ester linkage were found only in the feces. There were no significant differences in metabolism between the (lRS)-isomers and (lR)-isomers (EIliott et a!., 1976; Gaughan et a!., 1977).
Rat and mouse liver microsomes hydrolyzed the trans isomers more rapidly than the cis isomers, however, mouse microsomes oxidized the cis isomers more extensively than rat microsomes. The preferential oxidation sites of the alcohol moiety are the 4' -position for rats and the 2'-, 4'-, and 6positions for mice (Shono et a!., 1979; Soderlund and Casida, 1977).
58.19 PHENOTHRIN (d-PHENOTHRIN) Chemical Name 3-phenoxybenzyl (lRS)-cis-trans-2,2-dimethy 1-3-(2-methyIprop-l-enyl)cyclopropanecarboxylate. d-Phenothrin is a mixture of two isomers [«IR)-cis) and «IR)-trans)]. Synonyms Phenothrin (BSI, ISO, BAN) is the common name in use. The trade name is Sumithrin for d-phenothrin. Code designations include S-2539 for phenothrin and S-2539 Forte for d-phenothrin. The CAS registry number is 26002-S0-2. Physical and Chemical Properties The empirical formula is C23H2603; molecular weight is 350.5. Its form is a pale yellow to yellow-brown liquid; its specific gravity is 1.06 at 20°C; log Kow = 7.4. It is less soluble (0.01 mg/l) in water at 25°C, but is soluble in most organic solvents. It is unstable to light and in alkaline medium. Metabolism The metabolism of phenothrin was studied in rats after single oral or dermal application of trans- and cisphenothrin labeled with 14C in the alcohol moiety at several
58.19 Phenothrin
1281
HOH,C
C~COOCH''10Y0'0 ~ 4'-{)H, t-OH-c-jler
Cl
t CI>----,
J,
'I.
HOOC~O~
cr-~OOH
.V 0
;! 4'-{)H-fBacid
t-CI,CA
glucuronide
OH
ct-Permethrin
sulfate
J,
HOH,C
J,
ClOH,C
c~COOCH'~O~ V
HOH,C
V-7
0
tCr D PBalc
C~COOCH''10Y0~ l!...;::)
Cl
t-CH,OH-c-jler
t-CH,OH-t-per
CI'>=.~
c1- y
O
o
c-CH,OH-t-CI,CA~actone
I
CI>=i.H,OH
1-------+CI~H'OH
Cl -
~H
COOCH'tCr~ , 0D~ ,
PBald
X
~
H~H'Ci CVCOOH Cl t-C H,0 H-c-C I,C A
glucuronide
OH
CVCOOCH''10Y0~ Cl ~ ~ 2'-{)H-c-jler
c-CH,OH-t-per
C OOH c-CH,OH-t-CI,CA
0
Cl
H'OH
,~
c~o
COOH ---3> Cl 0 Cl c-CH OH-c-CI CA ,c-CH,OH-c-CI,CA2 2 lactone
HOOCCH,HNO~O~ sulfote
V~ PBocid-glycine
Figure 58.18 Metabolic pathways of pennethrin in animals.
doses (Kaneko et aI., 1981b; Miyamoto et aI., 1974; Suzuki et aI., 1976). The major excretion route of 14C was urine for the trans isomers, whereas it was feces for the cis isomers. The residue levels of 14C were generally very low except in fat, which showed slightly higher residue levels. Dermal absorption rate of 14C was estimated to be 3-17% in rats, depending on the formulation used. In addition, comparative metabolism of the six isomers [(1R)-trans, (1 RS)-trans, (IS)-trans, (1R)-cis, (1 RS)-cis, and (lS)-cis] of phenothrin was investigated after single oral administration of 14C-Iabeled preparation of the six phenothrin isomers to rats and mice at 10 mg/kg (Kaneko et aI., 1984c). There were no significant differences in the amount of 14C urinary and fecal excretion between (IR)-trans and (IRS)-trans isomers or between (IR)-cis and (lRS)-cis isomers. On the other hand, the (1S)-trans and (lS)-cis isomers showed slightly larger 14C urinary excretion than the corresponding (IR)- and (IRS)-trans and cis isomers, respectively (Izumi et aI., 1984). The major metabolic reactions (Fig. 58.19) of the cis isomer in both animals were oxidation at the cis- and trans-methyl of the isobutenyl group, at the trans-methyl of the gem-dimethyl group attached to the cyclopropane ring and the 4'-position of the alcohol moiety. Hydroxylation at the 2'-position of the alcohol moiety occurred to a smaller extent. On the other hand, with the trans isomer, the main metabolic reaction was cleavage of ester linkage. The hydroxylation at the 4'-position of the alcohol moiety occurred with the trans isomer to the same degree as that with the cis isomer. Moreover, small amounts of the metabolites hydroxylated at the trans-methyl of the isobutenyl
group and at the 2' -position of the alcohol moiety were obtained from the trans isomer. The species differences observed in phenothrin were as follows: (1) 4'-OH-PBacid-sulfate and PBacid-taurine were characteristic to rats and mice, respectively, as is the case with fenvalerate, deltamethrin, and cypermethrin. (2) All six phenothrin isomers underwent cleavage of the ester linkage and hydroxylation at the 4'-position of the alcohol moiety more rapidly in rats than in mice, and these facts may contribute to the higher urinary 14C excretion in rats than in mice (Izumi et aI., 1984; Kaneko et aI., 1984b; Miyamoto et aI., 1974;Suzukietal., 1976). With respect to metabolic reactions, the (1S)-trans and (IS)cis isomers underwent ester cleavage to a larger extent than the other chiral isomers in both rats and mice. There were some differences in the extent of oxidation at the 4' -position of the alcohol moiety of phenothrin isomers among three trans isomers and among three cis isomers. With the trans isomers, there was no apparent difference in the extent of hydroxylation at the 4'-position, whereas the (IS)-cis isomers underwent hydroxylation to a slightly larger extent compared with the other cis isomers. Based on the results, it may be concluded that the (IR) and (1RS) isomers behaved in similar metabolic manners, and that the (1 S) isomers received cleavage of the ester linkage slightly more readily than the corresponding (1R) and (1RS) isomers, although the (IS) isomers also received the same metabolic reactions as the (IR) and (IRS) isomers. However, purified carboxyesterase showed virtually no difference in ester hydrolysis between (1R)-trans and (IS)-trans isomers or between (IR)-cis and (IS)-cis isomers, although it hydrolyzed
1282
CHAPTER 58 Pyrethroid Chemistry and Metabolism
~
COOCH2tr°'O
ct-Phenothrin
trans-P henothrin
\Ester Metabolites
\Ester Metabolites
X
:~
3
R~COOCHUOn R2 I I :::-...
:::-...
R3 ~: CH 3 C H 0 H CH3 CH3 2 CH3 COOH CH3 OHH CH20 H CH3 CH3 CH3 CH20 H CH3 ~~ COOH CH3 CH3 CH3 CH3 COOH °HH CH3 COOH CH20 H CH3 COOH CH20 H OH R1
1 R4
HOOC~O~ ..-----'?taurine conjugate .0~ glucuronide :::-.. PBacid:::-'" ~glycine conjugate
0 __
I
R2
1) CH20 H CH3 2)
3) 4)
5) 6) 7) 8) 9)
Figure 58.19
2
r
~
~
HOOCU°'O HOOCUO
~ sulfate
4'-DH-PBacid
~
sulfate
~
OH
glucuronide
COOCHUOn R3
1) COOH
R1
Rz CH 3
R3 H
2) CH,OH
CH,
H
0
2'-{)H-PBacid
~
sulfate
Metabolic pathways of phenothrin in animals.
the (lR)-trans and (1S)-trans isomers more rapidly than the (1R)-cis and (lS)-cis isomers, respectively (Izumi et aI., 1984; Suzuki and Miyamoto, 1978).
58.20 PRALLETHRIN Chemical Name (S)-2-Methyl-4-oxo-3-prop-2-ynylcyclopent-2-eny1 (1 R)-cis-trans- 2,2-dimethy1-3-(2-methylprop-1eny1)cyclopropanecarboxylate. Synonyms Prallethrin (BSI, ISO) is the common name in use. The trade name is Etoc. Code designations include S-4068 SF. The CAS registry number is 23031-36-9. Physical and Chemical Properties The empirical formula is C19H2403; molecular weight is 300.4. Its form is a yellow to yellow-brown liquid; its specific gravity is 1.03 at 20°C; log Kow = 4.49. It is less soluble (8 mg/l) in water at 25°C, but is soluble in most organic solvents. Metabolism On single oral or subcutaneous administration of (4S, 1R)-trans- or (4S, 1R)-cis-prallethrin labeled with 14C in the alcohol moiety to rats at 2 mg/kg, 96-104% of the dosed 14C was eliminated into urine and feces within 7 days after administration. Monitoring of the expired air indicated that less than 0.1 % of the dose was excreted as 14C02 (Shiba et aI., 1988).
Urinary excretion of 14C with the trans isomer was larger (60-78%) than the cis isomer (17-32%), whereas 14C fecal excretion with the trans isomer was smaller than that with the cis isomer. 14C levels in blood and other tissues reached maximum within 3 hr after oral administration and thereafter decreased rapidly. 14C tissue residues were generally very low on the 7th day after administration (Shiba et aI., 1988). Nineteen metabo1ites were identified in urine and feces (Shiba et aI., 1988). Two new types of S-linked conjugation (sulfonic acid and mercapturic acid types) were additionally identified (Tomigahara et al., 1994c). The major biotransformation reactions (Fig. 58.20) of prallethrin are summarized as follows: (1) oxidation at the methyl group of the isobutenyl group in the acid moiety; (2) hydroxylation at the C-1 or the C-2 position of the propynyl group in alcohol moiety; (3) cleavage of the ester linkage; and (4) conjugation of resulting metabolites with glucuronic acid, sulfuric acid, sulfonic acid, or mercapturic acid.
58.21 PYRETHRINS Chemical Name The six known insecticidally active compounds in pyrethrum are esters of two acids and three alcohols. Specifically, pyrethrin I is the pyrethrolone ester of chrysanthemic acid, pyrethrin 11 is the pyrethrolone ester of pyrethric
58.22 Resmethrin
-
'!=4coo~
0coo~
0
ct-Prallethrin
mercapturic acid conjugate ~
COO~ o Rl
1) CH 20 H 2) CO OH 3) CH 3 4) CH 20 H 5) CH 3
Rz CH 3 CH 3 CH3 CH 3 CH 3
J
sulfonic acid
~conjugate
HO~~~--.~ o
Rl~~ R2
glucuronide
~ HO W
~
1
R2
R3
~
COO~
W· o
HO-(yy
o
o
OHOH
glucuronide
I Ester Metabolltes I R
/\
R3
11 11 12 12 13
r--
0
trons - Prallethrin
I Ester Metaboiltes I
1283
~ glucuronide
sulfate
1) 2) 3) 4) 5)
Rl CH 3 CH 3 CH 3 CH 3 CH 3
R, CH 20H CO OH CH3 CH 20 H CH3
R3
11 11 12 13 13
o Figure 58.20 Metabolic pathways of prallethrin in animals.
acid, cinerin I is the cinerolone ester of chrysanthemic acid, cinerin 11 is the cinerolone ester of pyrethric acid, jasmolin I is the jasmolone ester of chrysanthemic acid, and jasmolin 11 is the jasmolone ester of pyrethric acid.
Synonyms Pyrethrins (BSI, ISO, JMAF, ESA) is the common name in use. There are several trade names such as Alfadex, Evergreen, and ExciteR. The CAS registry numbers are 121-21I (pyrethrin I), 4466-14-2 (jasmolin I), 25402-06-6 (cinerin I), 121-29-9 (pyrethrin 11), 1172-63-0 (jasmolin 11), and 121-20-0 (cinerin 11). Physical and Chemical Properties molecular weights: Pyrethrin I Pyrethrin 11 Cinerin I Cinerine 11 Jasmolin I Jasmolin 11
C21 H2S 0 3 C22H2S0S C20 H 2S 0 3 C21 H 2S 0 S C21 H30 0 3 C22 H 30 0 S
Empirical formulas and 328.4 372.4 316.4 360.4 330.4 374.5
Their forms may be viscous oils or tan dusts; log Kow = 5.9 (pyrethrin I), 4.3 (pyrethrin 11). They are practically insoluble in water, but are readily soluble in most organic solvents. They are unstable to light and air.
Metabolism On single oral administration of pyrethrin I and pyrethrin 11 labeled with 3H or 14C in the acid or alcohol moiety
to rats and mice at 1-5 mg/kg, the 3H and 14C from both compounds were excreted into urine and feces almost equally. The trans-methyl group of the isobuteny1 side group of pyrethrin I is readily oxidized to hydroxymethyl derivatives, and further to the corresponding carboxylic acid derivatives through aldehydes. Hydrolyis of the methyl ester of pyrethrin 11 yields the same acid. This acid is further oxidized at the pentadienyl group of the alcohol moiety, which is probably initiated by epoxidation of the terminal double bond. Hydrolysis of this epoxide resulted in formation of isomeric diols, one of which is conjugated with an unidentified aromatic acid (Elliott et al., 1972). The major metabolic reactions (Fig. 58.21) of both compounds were hydrolysis of methoxycarbonyl group, oxidation at the trans-methyl of the isobutenyl group of the acid moiety, and epoxidation at the pentadienyl side chain to yield initially a 4,5-epoxide from which the two diol derivatives were derived (Casida et aI., 1971; Elliott et aI., 1972).
58.22 RESMETHRIN (BIORESMETHRIN, CISMETHRIN) Chemical Name 5-benzyl-3-furylmethyl (IRS)-cis-trans-3(2-methylprop-l-enyl)-cyclopropanecarboxylate. Synonyms Resmethrin (BSI, ISO, ANSI, JMAF, ESA) is the common name in use. Trade names are Chrysron and Synthrin. Code designations include SBP-1382, NRDC104, and FMC17370. Bioresmethrin and cismethrin are (lR)-trans and
1284
CHAPTER 58 Pyrethroid Chemistry and Metabolism
FX,oo~ Pyrethrin I
CH300C~
1. ~
C00
0
Pyrethrin II
4
i
HOOC
'
A
COOH
~
OHC
Wt;- or wc-<Jcid -c- or t;-cA
0
~
cis~trans
interconversion --;. initiated by COOH proton abstraction at the C-3 carbon
\
1
wt;- or W.r;Dlcc-or t-CA
HOH?FX,oo~ Palc
i
0
~COOCHp;o --<
1 HOOC~
~HOOH
COO
PMA
1
glucuronide
~
1
HOOC~ ~OH coo
OH PMB
Figure 58.21
Resmethrin
loocFX,oo~
0
conjugote
HOH'C~COOH
D(lC) OH
/-
o (;r'0 HOH,C~FA V
o
i HOOC
t;-CA
~v~ HOOC~FCA V
1
o
o
0
~COOH c-or
HOOC
sG/Cl 4'-0H-BFCA
OH
sGi'O
HOOC u-DH-BFCA
1
Metabolic pathways of pyrethrin in animals.
OH
1o
glucuronide, sulfote
if'O
HOOC u-Keto-BFCA
(1R)-eis isomers of resmethrin, respectively. The CAS registry number is 10453-86-8. Physical and Chemical Properties The empirical formula is C22H2603; molecular weight is 338.4. Its form is a colorless crystal; its specific gravity is 0.958-0.968 at 20°C; log Kow = 5.43. It is less soluble (0.038 mg/l) in water at 25°C, but is soluble in most organic solvents. It is unstable to air and light. Metabolism The metabolism of resmethrin was studied in rats after single oral administration of (1RS,trans)-, (1 R,trans)-, or (1 R,eis)-resmethrin labeled with 14C in the alcohol or acid moiety (Miyamoto et ai., 1986; Ueda et ai., 1975a). In addition, in vitro studies with mouse and rat liver microsomes have also been carried out to elucidate metabolism of resmethrin in detail (Ueda et al., 1975b). Resmethrin was rapidly metabolized and the 14C from the acid moiety was more rapidly eliminated from the body than that from the alcohol moiety. The major metabolites were 5-benzyl-3-furoic acid (BFCA), 4'-OH-BFCA, and a-OH-BFCA from the alcohol moiety and hydroxymethyl and dicarboxylic acid derivatives of chrysanthemic acid from the acid moiety. Although the trans-methyl group of the isobutenyl group was predominantly oxidized in bioresmethrin and cismethrin, the eis-methyl group of cismethrin was also oxidized. In addition, epimerization at C-3 of the cyclopropane ring was found, leading to isomerized forms of dicarboxylic acid derivatives of chrysanthemic acid. Bioresmethrin was hydrolyzed
Figure 58.22
Metabolic pathways of resmethrin in animals.
more rapidly in rat liver microsomes than cismethrin, whereas the plasma did not show this eis-trans specificity (White et ai., 1976). The major metabolic reactions (Fig. 58.22) of resmethrin are cleavage of the ester linkage, oxidation at the trans- and the eis-methyl group of the isobutenyl group, at the 4' -position of the benzene ring, and at the benzylic position of the alcohol moiety, and conjugation of the resulting metabolites with glucuronic acid and sulfuric acid.
58.23 TETRAMETHRIN (d- TERAMETHRIN) Chemical Name 3,4,5,6-Tetrahydrophthalimidomethyl (1 RS)-cis-trans-chrysanthemate; d-tetramethrin is composed of the (1 R) isomers. Synonyms Tetramethrin (BSI, ISO, ANSI) is the common name in use. Trade names are Neo-Pynamin and Duracide for tetamethrin and Neo-Pynamin Forte, Tedion V-18, Duracide, and Tedone for d-tetramethrin. Code designations include SP 1103 and FMC 9260. The CAS registry numbers are 7696-12-0; 51348-90-4 for (1 R)-cis isomer and 1166-46-7 for (1 R)-trans isomer.
58.23 Tetramethrin (d-Teramethrin)
HOOC~ -
0
COOCH2N~
~ o S03 H
""",0\ c;d-J-
~O
0
--------7)
HN~ ~ ~ 0 S03 H
X
HOCH2N~ o S03H
~c0COOCH'N~
/
HOOCl('l
HOOC
HOOC
c'6-A cid-3"'() H-N PY -SA
~~
~ /
°MTI
HP!
;/
o
HOOCil
HOOC~O
OH 1"'() H-5-{)xo-11 PA
~
HN;=O
"Tolmm'Ih'~
o S03 H
Figure 58.23
H2NOC~
~
0
~COOCH2N~
3"'() H-MT!-SA
I
HOCH'N~
",,,-T,leom,lhde
OH
HN
1285
1
OH
HN~
rr-V o
HOOCX) HOOC
3"'()H-11PI-1 and 2
TCDA
HOOC il HOOCt-/ OH 1"'()H-11PA
Metabolic pathways of tetramethrin in animals.
Physical and Chemical Properties (Tetramethrin) The empirical formula is C19H2SN04; molecular weight is 331.4. Its form is a colorless crystal; its specific gravity is 1.1 at 20°C; its vapor pressure is 0.944 mPa at 30°C; log Kow = 4.6. It is less soluble (1.83 mg/l) in water at 25°C, but is soluble in most organic solvents. It is unstable in strong acid and alkaline medium. Metabolism In vivo and in vitro metabolism of this pyrethroid in rats had been reported several times in the 1960s, 1970s, 1980s, and 1990s. The metabolic fates of the alcohol moiety, the tetrahydrophthalimido group, appeared to be completed in the 1990s. On the other hand, the metabolic fate of the acid moiety was the same as that of pyrethroids having chrysanthemic acid such as resmethrin (Kaneko et aI., 1981c; Ueda et aI., 1975a). When 14C-(lRS)-trans- or 14C-(1RS)-eis-tetramethrin labeled in the alcohol moiety was administered orally to rats at 2 or 250 mglkg, the 14C was almost completely eliminated from rats within 7 days after administration. 14C recovery in feces and urine was 38-56 and 42-58%, respectively, with the trans isomer, and 66-91 and 9-31%, respectively, in feces and urine with the eis isomer. Fourteen metabolites were found in excreta. For both isomers the main metabolites were sulfonate derivatives in feces and alcohol derivatives and dicarboxylic acid derivatives derived from the 3,4,5,6-tetrahydrophthalimide moiety in urine. The sulfonic acid conjugates have a sulfonic acid group incorporated into the double bond of the 3,4,5,6tetrahydrophthalimide moiety. Two of five sulfonic acid conjugates (trans-Acid-3-0H-NPY-SA and TPI-SA) were detected in the urine; however, their amounts were smaller than in the feces. In addition, the sulfonic acid conjugates were not detected in the bile or urine of the bile-duet-cannulated rat given 14C_
alcohol-trans- or eis-tetramethrin. Therefore, it is likely that the sulfonic acid conjugates were produced in the intestinal tract (Tomigahara et aI., 1994b). The major metabolic reactions (Fig. 58.23) of trans- and eistetramethrin in rats were as follows: (I) cleavage of ester linkage; (2) cleavage of the imide likage; (3) hydroxylation of cyclohexene or cyclohexane ring of the 3,4,5,6tetrahydrophthalimide moiety; (4) oxidation at the methyl group of the isobutenyl moiety of the acid moiety; (5) reduction at the 1,2-double bond of the 3,4,5,6-tetrahydrophthalimide moiety; and (6) incorporation of a sulfonic acid group into the 1,2-double bond of the 3,4,5-6-tetrahydrophthalimide moiety (Kaneko et aI., 1981c; Miyamoto et aI., 1968; Tomigahara et aI., 1994a, b, 1996). The 14C was rapidly and almost completely excreted into the urine and feces after a single oral dose of each ofthe 14C-(1 R)trans, (lRS)-trans, (lR)-eis, and (1 RS)-cis isomers labeled in the acid moiety to pregnant rats at 300 mglkg. The 14C from both the trans isomers was excreted into the urine to a larger extent than the feces. There were no statistical differences in the 14C excretion between both the trans isomers. The 14C derived from both the eis isomers was almost completely eliminated from rat body for 7 days as with the trans isomers. However, significant differences were observed in the excretion of the 14C into the urine and feces between the (lR)- and (lRS)-cis isomers. There were no appreciable differences in 14C tissue residues and the nature of metabolites between the (lR)- and (1 R S)-cis isomers, although there were significant differences in absorption in the gastrointestinal tissues between these isomers. From these findings, once entered into the bloodstream, the (lR)- and (lRS)-eis isomers apparently undergo the same metabolic reactions. Overall there seems to be no significant
1286
CHAPTER 58
Pyrethroid Chemistry and Metabolism
~{ JVCOOqH~O ~ Br7l}\ Br Br
CN
V
V
Br
>="{
8r
~N V
V
Dellamethrin
Tralomethrin
B{ JVCOOqH~O~ CITI}\ CN Cl Br Tralocythrin
V
COOCH-
V
Cl, JVCOOCH-
V
---7 refer to
deltamethrin metabolism
~
V
---7 refer to
cypermethrin metabolism
Figure 58.24
differences in the metabolic reactions between the trans isomers and between the cis isomers. The nature and amounts of the metabolites detected in the fetus with Neo-Pynamin [a mixture of (lRS)-isomers] and Neo-Pynamin Forte [a mixture of (lS)-isomers] were nearly the same, indicating that these two compounds behaved in the same way to give the same placental transfer regardless of their optical differences (Kaneko et aI., 1984b).
58.24 TRALOMETHTIN Chemical Name (S)-a-Cyano-3-phenoxybenzyl (l R)-cis-2, 2-dimethy1-3-[( R S)-l ,2,2,2-tetrabromoethyl] cyclopropanecarboxylate. Synonyms Tralomethrin (BSI, ANSI, ISO) is the common name in use. Trade names are Saga, Scout, Tralox, Tracker, and Tralate. Code designations include RU 25474, NU83l, and HAG 107. The CAS registry number is 66841-25-6. Physical and Chemical Properties The empirical formula is C22H19Br4N03; molecular weight is 665.0. Its form is an orange to yellow resinous solid; its specific gravity is 1.70 at 20°C; log Kow is about 5. It is less soluble (0.080 mg/l) in water at 25°C, but is soluble in most organic solvents. Metabolism A comparative metabolism study has been carried out on the fate of tralomethrin and deltamethrin in male rats after single oral administration of 14C-tralomethrin and 14C_ deltamethrin labeled in the acid moiety, the alcohol moiety, and the CN group at 0.3-0.5 mg/kg (Cole et aI., 1982). Tralomethrin was not normally detected in the treated animals or their excreta because it undergoes rapid and substantially complete debromination to form deltamethrin. After formation of deltamethrin, this compound appears to undergo the same metabolic fate of deltamethrin (Fig. 58.24). Similarly, tralocythrin was rapidly converted to (1 R, a S)-cis-cypermethrin by debromination as is the case with tralomethrin (Cole et aI., 1982). The debromination was mediated by tissue thiols such as glutathione (Kaneko et aI., 1986b;Ruzoetal., 1981).
58.25 SUMMARY All the pyrethroid insecticides investigated so far are rapidly metabolized in mammals and their metabolites are almost completely excreted in the urine and feces within several days of a single oral or subcutaneous administration except their cyano moiety. Tissue residues are generally very low, showing that they are biodegradable and nonbioaccumulative (Miyamoto et aI., 1995). This is due partly to rapid metabolism. The major metabolic reactions of pyrethroids are commonly oxidation of methyl groups and aromatic rings in the molecules, hydrolysis of the ester linkage, and several types of conjugation reactions. With respect to cleavage of the ester linkage, there are significant differences between geometrical trans and cis isomers. The trans isomers of pyrethroids having chrysanthemic acid derivatives in the acid moiety such as phenothrin and permethrin are more rapidly hydrolyzed than the corresponding cis isomers, and the cis isomers yield more metabolites retaining intact ester linkage than the trans isomers (Casida and Ruzo, 1980; Leahey, 1985; Miyamoto, 1976, 1981; Ruzo and Casida, 1977). With respect to oxidation, the trans-methyl of the isobutenyl group in chrysanthemates is oxidized preferentially to the cismethyl group, and the 4'-position of the phenoxy ring in the phenoxybenzyl derivatives is more readily oxidized than all other positions. Conjugation reactions of pyrethroids include hydrophilic conjugations with glucuronic acid, sulfuric acid, and amino acids and lipophilic conjugations with cholesterol, bile acid, and triglyceride. The optical isomers of pyrethroids do not show significant differences in metabolism reactions, except that only one isomer of the four optical isomers of fenvalerate produces the cholesterol ester conjugate. Although human data are very limited, there seems to be substantially no remarkable species difference in metabolism between laboratory animals and human beings. Ester hydrolysis and oxidation reaction seem to be major metabolic reactions in human beings for pyrethroids.
REFERENCES Casida, J. E., and Ruzo, L. O. (1980). Metabolic chemistry of pyrethroid insecticides. Pestic. Sci. 11,257-269. Casida, J. E., Kimmel, E. c., Elliott, M., and Janes, N. F. (1971). Oxidative metabolism ofpyrethrins in mammals. Nature 230,326-327.
References
Chamberlain, K, Matsuo, N., Kaneko, H., and Khambay, B. P. S. (1998). Pyrethroids. In "Chirality in Agrochemicals" (N. Kurihara and J. Miyamoto, eds.), pp. 9-84. Wiley, New York. Class, T. J., Ando, T., and Casida, J. E. (1990). Pyrethroid metabolism: Microsomal oxidation metabolites of (S)-bioallethrin and the six natural pyrethrins. J. Agrie. Food Chem. 38, 529-537. Cole, K. M., Ruzo, L. 0., Wood, E. J., and Casida, J. E. (1982). Pyrethroid metabolism: Comparative fate in rats of tralomethrin, tralocythrin, deltamethrin and (IR,aS)-eis-cypermethrin. J. Agrie. Food Chem. 30,631-636. Crawford, M., and Hutson, D. H. (1977). The metabolism of the pyrethroid insecticide (±)-a-cyano-3-phenoxybenzyl 2,2,3,3tetramethylcyclopropanecarboxylate, WL41706, in the rats. Pestie. Sci. 8,579-599. Crawford, M. J., Croucher, A, and Hutson, D. H. (1981a). Metabolism of cisand trans-cypermethrin in rats. Balance and tissue retention study. J. Agrie. Food Chem. 29, 130-135. Crawford, M. J., Croucher, A, and Hutson, D. H. (1981b). The metabolism of the pyrethroid insecticide cypermethrin in rats: Excreted metabolites. Pestic. Sci. 12, 399-411. Edwards, R., Millburn, P., and Hutson, D. H. (1990). Comparative metabolism and disposition of [14C-benzyll-cypermethrin in quail, rat and mouse. Pestic. Sci. 30, 159-181. Elliott, M., Janes, N. F., Kimmel, E. C., and Casida, J. E. (1972). Metabolic fate of pyrethrin I, pyrethrin n, and allethrin administered orally to rats. J. Agrie. Food Chem. 20,300-313. Elliott, M., Janes, N. F., Pulman, D. A., Gaughan, L. C., Unai, T., and Casida, J. E. (1976). Radiosynthesis and metabolism in rats of the lRisomers of the insecticide permethrin. J. Agrie. Food Chem. 24, 270-276. Food and Agriculture Organization-World Health Organization (FAOIWHO) (1985). Flucythrinate. In "Pesticide Residues in Food: 1985 Evaluations. Part I. Residues." FAO Plant Production and Protection Paper, Food and Agriculture Organization of the United Nations, Rome. FAOIWHO (1984). Cyhalothrin. In "Pesticide Residues in Food: 1984 Evaluations. Part n. Toxicology." FAO Plant Production and Protection Paper, Food and Agriculture Organization of the United Nations, Rome. FAOIWHO (1986). Cyfluthrin. In "Pesticide Residues in Food: 1986 Evaluations. Part n. Toxicology." FAO Plant Production and Protection Paper, Food and Agriculture Organization of the United Nations, Rome. FAOIWHO (1992). Bifenthrin. In "Pesticide Residues in Food: 1992 Evaluations. Part n. Toxicology." FAO Plant Production and Protection Paper, Food and Agriculture Organization of the United Nations, Rome. FAOIWHO (1993). Etofenprox. In "Pesticide Residues in Food: 1993 Toxicology Evaluations." FAO Plant Production and Protection Paper, Food and Agriculture Organization of the United Nations, Geneva. FAOIWHO (1996). Flumethrin. In "Pesticide Residues in Food: 1996 Toxicology Evaluations." FAO Plant Production and Protection Paper, Food and Agriculture Organization of the United Nations, Geneva. Gaughan, L. C., Unai, T., and Casida, J. E. (1977). Permethrin metabolism in rats. J. Agrie. Food Chem. 25, 9-17. Huckle, K R., Hutson, D. H., and Millburn, P. (1981). Species differences in the metabolism of 3-phenoxybenzoic acid. Drug Metab. Dispos. 9, 352-359. Hutson, D. H., and Casida, J. E. (1978). Taurine conjugation in metabolism of 3-phenoxybenzoic acid and the pyrethroid insecticide cypermethrin in mouse. Xenobiotiea 8, 565-571. Hutson, D. H., Gaughan, L. C., and Casida, J. E. (1981). Metabolism of the eisand trans-isomers of cypermethrin in mice. Pestie. Sci. 12, 385-398. International Programme on Chemical Safety (IPCS) (1989). "AllethrinsAllethrin, d-Allethrin, Bioallethrin and S-Bioallethrin," Environmental Health Criteria 87. World Health Organization, Geneva. IPCS (1990a). "Cyhalothrin," Environmental Health Criteria 99. World Health Organization, Geneva. IPCS (1990b). "Deltamethrin," Environmental Health Criteria 97. World Health Organization, Geneva. Isobe, N., Kaneko, H., Shiba, K, Saito, K, Ito, S., Kakuta, N., Saito, A., Yoshitake A, and Miyamoto, J. (1990). Metabolism of esfenvalerate in rats and
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mice and effects of its isomers on metabolic fates of esfenvalerate. 1. Pestie. Sei. 15, 159-168. Isobe, N., Suzuki, T., Nishikawa, J., Kaneko, H., Nakatsuka, I., and Yoshitake, A. (1992). Metabolism of empenthrin isomers in rats. J. Pestie. Sci. 17,27-37. Izumi, T., Kaneko, H., Matsuo, M., and Miyamoto, J. (1984). Comparative metabolism of the six stereoisomers of phenothrin in rats and mice. J. Pestic. Sci. 9,259-267. Kaneko, H., lzumi, T., Matsuo, M., and Miyamoto, J. (1984a). Metabolism of fenvalerate in dogs. J. Pestie. Sci. 9, 269-274. Kaneko, H., lzumi, T., Ueda, Y., Matsuo, M., and Miyamoto, J. (1984b). Metabolism and placental transfer of stereoisomers of tetramethrin isomers in pregnant rats. J. Pestie. Sci. 9,249-258. Kaneko, H., Matsuo, M., and Miyamoto, J. (1984c). Comparative metabolism of stereoisomers of cyphenothrin and phenothrin isomers in rats. J. Pestie. Sci. 9,237-247. Kaneko, H., Matsuo, M., and Miyamoto, J. (1986a). Differential metabolism of fenvalerate and granuloma formation. I. Identification of a cholesterol ester derived from a specific chiral isomer of fenvalerate. Toxieol. Appl. Pharmaeol. 83, 148-156. Kaneko, H., Ohkawa, H., and Miyamoto, J. (1981a). Comparative metabolism of fenvalerate and the [2S,aSl-isomer in rats and mice. J. Pestie. Sei. 6, 317-326. Kaneko, H., Ohkawa, H., and Miyamoto, J. (1981b). Adsorption and metabolism of dermally applied phenothrin in rats. 1. Pestie. Sci. 6, 169182. Kaneko, H., Ohkawa, H., and Miyamoto, J. (1981c). Metabolism of tetramethrin isomers in rats. J. Pestie. Sei. 6, 425-435. Kaneko, H., Shiba, K, Yoshitake, A, and Miyamoto, J. (1987). Metabolism of fenpropathrin (S-3206) in rats. J. Pestie. Sci. 12, 385-395. Kaneko, H., Takamatsu, Y., Kitamura, N., Yoshitake, A., and Miyamoto, 1. (1986b). In vivo and in vitro conversion of tralomethrin to deltamethrin in larvae of tobacco cutworm. Spodoptera litura. J. Pestie. Sei. 11, 533-540. Kaneko, H., Takamatsu, Y., Okuno, Y., Abiko, J., Yoshitake, A., and Miyamoto, J. (1988). Substrate specificity for formation of cholesterol ester conjugates from fenvalerate analogues and for granuloma formation. Xenobiotiea 18, 11-19. Leahey, J. P. (1985). "The Pyrethroid Insecticides." Taylor & Francis, London. Leng, G., Leng, A., Kuhn, K-H., Lewalter, J., and Pauluhn, J. (1997). Human dose-excretion studies with the pyrethroid insecticide cyfluthrin: Urinary metabo1ite profile following inhalation. Xenobiotiea 27, 1273-1283. Miyamoto, J. (1976). Degradation, metabolism and toxicity of synthetic pyrethroids. Environ. Health Perspeet. 14, 15-28. Miyamoto, J. (1981). The chemistry, metabolism and residue analysis of synthetic pyrethroids. Pure Appl. Chem. 53, 1967-2022. Miyamoto, J., Kaneko, H., and Takamatsu, Y. (1986). Stereose1ective formation of a cholesterol ester conjugate from fenvalerate by mouse microsomal carboxyesterase(s). J. Bioehem. Toxieol. 1,79-94. Miyamoto, J., Kaneko, H., Tsuji, R., and Okuno, Y. (1995). Pyrethroids, nerve poisons: How their risks to human health should be assessed. Toxieol. Left.
82/83, 933-940. Miyamoto, J., Nishida, T., and Ueda, K (1971). Metabolic fate ofresmethrin, 5-benzyl-3-furylmethyl dl-trans-chrysanthemate in the rats. Pestie. Bioehem. Physiol. 1, 293-306. Miyamoto, J., Sato, Y., Yamamoto, K, Endo, M., and Suzuki, S. (1968). Biochemical studies on the mode of action of pyrethroidal insecticides. Part 1. Metabolic fate of phthalthrin in mammals. Agrie. BioI. Chem. 32, 628-640. Miyamoto, J., Suzuki, T., and Nakae, C. (1974). Metabolism of phenothrin or 3-phenoxybenzyl d-trans-chrysanthemumate in mammals. Pestie. Bioehem. Physiol. 4, 438-450. Ohkawa, H., Kaneko, H., Tsuji, H., and Miyamoto, J. (1979). Metabolism of fenvalerate (Sumicidin) in rats. 1. Pestie. Sei. 4, 143-155. Ohsawa, K., and Casida, J. E. (1980). Metabolism in rats of the potent knockdown pyrethroid kadethrin. J. Agrie. Food Chem. 28, 250-255. Okuno, Y., Seki, T., Ito, S., Kaneko, H., Watanabe, T., Yamada, H., and Miyamoto, J. (1986). Differential metabolism of fenvalerate and granuloma
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formation. H. Toxicological significance of a lipophilic conjugate from fenvalerate. Toxicol. Appl. Pharmacol. 83, 157-169. Quistad, G. B., and Selim, S. (1983b). Fluvalinate metabolism by rhesus monkeys. 1. Agric. Food Chem. 31, 596--599. Quistad, G. B., Staiger, L. E., Jamieson G. c., and Schooley, D. A. (1983). Fluvalinate metabolism by rats. 1. Agric. Food Chem. 31, 589-596. Quistad, G. B., Staiger, L. E., and Schooley, D. A. (1982). Xenobiotic conjugation: A novel role for bile acids. Nature 296, 462--464. Roberts, T., and Hutson, D. (1999). "Metabolic Pathways of Agrochemicals. Part 2. Insecticides and Fungicides." The Royal Society of Chemistry Information Services, Cambridge. Ruzo, L. 0., and Casida, J. E. (1977). Metabolism and toxicology ofpyrethroids with dihalovinyl substitutuents. Environ. Health Perspect. 21, 285-292. Ruzo, L. 0., Engel, J. L., and Casida, J. E. (1979). Decamethrin metabolites from oxidative, hydrolytic and conjugative reactions in mice. 1. Agric. Food Chem. 27,725-731. Ruzo, L. 0., Gaughan, L. c., and Casida, J. E. (1981). Metabolism and degradation of the pyrethroids tralomethrin and tralocythrin in insects. Pestic. Biochem. Physiol. 15, 137-142. Ruzo, L. 0., Vnai, T., and Casida, J. E. (1978). Decamethrin metabolism in rats. 1. Agric. Food Chem. 26,918-924. Saito, K, Kaneko, H., Tomigahara, Y, Nakatsuka, 1., and Yamada, H. (1995). Metabolism of imiprothrin isomers in rats: Biotransformation and excretion.l. Pestic. Sci. 20, 529-540. Saito, K, Kaneko, H., Tomigahara, Y, Nakatsuka, 1., and Yamada, H. (1996). Metabolism of imiprothrin isomers in rats: Absorption and distribution. 1. Pestic. Sci. 21,49-55. Seguchi, K, Asaka, S., Katoh, Y, and Yamaguchi, 1. (1991). Metabolism of cycloprothrin in rats. 1. Pestic. Sci. 16, 591-598. Shiba, K, Kakuta, N., Kaneko, H., Nakatsuka, 1., Yoshitake, A., Yamada, H., and Miyamoto, J. (1988). Metabolism of the pyrethroid insecticide S-4068F in rats. 1. Pestic. Sci. 13,557-569. Shiba, K, Kaneko, H., Kakuta, N., Yoshitake, A., and Miyamoto, J. (1990). Placental transfer of esfenvalerate and fenvalerate in pregnant rats. 1. Pes tic. Sci. 15, 169-174. Shono, T., Ohsawa, K, and Casida, J. E. (1979). Metabolism of trans- and cispermethrin, trans- and cis-cypermethrin and deltamethrin by microsomal enzymes. 1. Agric. Food Chem. 27, 316--325.
Soderlund, D. M., and Casida, J. E. (1977). Effects of pyrethroid structure on rate of hydrolysis and oxidation by mouse liver microsomal enzymes. Pestic. Biochem. Physiol. 7, 391--401. Staiger, L. E., and Quistad, G. B. (1984). Fluvalinate metabolism in rats. 1. Agric. Food Chem. 32, 1130-1133. Suzuki, T., and Miyamoto, J. (1978). Purification and properties of pyrethroid carboxyesterase in rat liver microsome. Pestic. Biochem. Physiol. 8, 186-198. Suzuki, 1., Ohno, N., and Miyamoto, J. (1976). New metabolites of (+)-cis fenothrin, 3-phenoxybenzyl (+ )-cis chrysanthemumate, in rats. 1. Pestic. Sci. 1,151-152. Tomigahara, Y, Mori, M., Shiba, K., Isobe, N., Kaneko, H., Nakatsuka, 1., and Yamada, H. (1994a). Metabolism of tetramethrin isomers in rat. I. Identification of a sulphonic acid type of conjugate and reduced metabolites. Xenobiotica 24, 473--484. Tomigahara, Y, Mori, M., Shiba, K, Isobe, N., Kaneko, H., Nakatsuka, I., and Yamada, H. (1994b). Metabolism of tetramethrin isomers in rat. H. Identification and quantitation of metabolites. Xenobiotica 24, 1205-1214. Tomigahara, Y, Onogi, M., Miki, M., Yanagi, K., Shiba, K, Kaneko, H., Nakatsuka, 1., and Yamada, H. (1996). Metabolism oftetramethrin isomers in rat. HI. Stereochemistry ofreduced metabolites. Xenobiotica 26, 201-210. Tomigahara, Y, Shiba, K, Isobe, N., Kaneko, H., Nakatsuka, I., and Yamada, H. (1994c). Identification of two new types of S-linked conjugates of Etoc in rat. Xenobiotica 24, 839-852. Tomlin, C. D. S. (1997). "A World Compendium, The Pesticide Manual," 11th ed. British Crop Protection Council, Berks. Veda, K, Gaughan, L. c., and Casida, J. E. (1975a). Metabolism of (+)-transand (+)-cis-resmethrin in rats. 1. Agric. Food Chem. 23,106--115. Veda, K, Gaughan, L. c., and Casida, J. E. (1975b). Metabolism of four resmethrin isomers by liver microsomes. Pestic. Biochem. Physiol. 5, 280294. White, I. N. H., Verschoyle, R. D., Moradian, M. H., and Barnes, J. M. (1976). The relationship between brain levels of cismethrin and bioresmethrin in female rats and neurotoxic effects. Pestic. Biochem. Physiol. 6,491-500. Woollen, B. H., Marsh, J. R., Laird, W. J. D., and Lesser, J. E. (1992). The metabolism of cypermethrin in man: Differences in urinary metabolic profiles following oral and dermal administration. Xenobiotica 22, 983-991.
CHAPTER
59 Pyrethroid Insecticides: Mechanisms of Toxicity, Systemic Poisoning Syndromes, Paresthesia, and Therapy David E. Ray MRC Applied Neuroscience Group
59.1 MECHANISMS OF TOXICITY The widespread use of the pyrethroids in agricultural and public health applications has acted as a stimulus for study of their mode of action in insects and also of their toxicity to nontarget organisms. Generally speaking, these two actions are similar, the molecular targets present in insects being qualitatively similar to those seen in mammals. The relative resistance of mammals to pyrethroids is attributable to a combination of their faster metabolic disposal, higher body temperature, and a lower sensitivity of the analogous target sites (Song and Narahashi, 1996b). The toxicology of the pyrethroids has been reviewed by several authors (Aldridge, 1990; Narahashi et aI., 1998; Soderlund and Knipple, 1995; Vijverberg and van den Bercken, 1990). Many different mechanisms of action have been proposed for the pyrethroids, which are potent agents with a wide range of actions on biological systems (Table 59.1). Any evaluation of these competing mechanisms of action must be informed by knowledge of potency, because only those effects that are only seen at the lowest concentrations are likely to have much biological significance. Unfortunately, much of the published data on pyrethroid concentrations must be used with caution, because the pyrethroids have very low water solubilities, readily partition into lipids, and show potent binding to glass and plastics. This means that it is very difficult to be certain what the true concentration is (Ray et aI., 1997) or to compare directly across in vitro systems with different physicalchemical characteristics. Furthermore, whereas actions at some of these target sites have appreciable biological consequences at less than 1% modification (Song and Narahashi, 1996b), others may require 90% modification before an effect is seen. Hence, in constructing Table 59.1, the lowest concentration needed to Handbook of Pesticide Toxicology Volume 2. Agents
produce a significant biological response has been cited, rather than the ED5o, which can underestimate potency. Pyrethroids are primarily functional toxins (Narahashi et aI., 1998), readily causing hyperexcitation, but having little or direct cytotoxic potential in mammalian cells, although insect cells are more vulnerable to pyrethroid toxicity (Meola et al., 1997). Thus, inexcitable mammalian cells are little affected by pyrethroids: A number of pyrethroids only produce growth inhibition at 10-5 M and no cytotoxicity at 10-4 M (Squiban et aI., 1986). In contrast, the interaction with the sodium channel shows dissociation constants on the order of 10- 8 M (Soderlund, 1985), producing profound hyperexcitation in nerve or muscle cells. Similarly, hepatocytes in culture showed decreased viability after treatment with permethrin only at 100 ng/1 06 cells (El Tawil and Abdel Rahman, 1997), a level that is approximately 10-100 times higher than that reached in the brain of intoxicated animals. This marked resistance of inexcitable cells to pyrethroids has enabled purely in vitro studies to be made at pyrethroid concentrations many orders of magnitude higher than could be survived in vivo. Obviously, such high-concentration studies are of very limited value in any extrapolation to human toxicology.
59.1.1 ACTIONS ON SODIUM CHANNELS Voltage-gated sodium channels are vital to the function of most excitable cells and are seen in all organisms from jellyfish upward. They are responsible for the generation of the inward sodium current that produces the action potential in most cells and are closed at normal resting potentials. Their structure and function have been recently reviewed (ConIey and Brammar, 1999; Marban et aI., 1998). They consist of an ex subunit, which
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Copyright © 2001 by Academic Press. All rights of reproduction m any form reserved.
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CHAPTER 59 Pyrethroid Insecticides
Table 59.1 Some Potential Targets of Pyrethroids in Mammals Target
Concentrationa
Reference
Protein phosphorylation
10- 13 M
Enan and Matsumura (1993)
Voltage-gated sodium channels
10- 10 M
Ghiasuddin and Soderlund (1985)
Voltage-gated chloride channels
Ray et al. (1997)
Noradrenaline release
10- 10 M 10- 10 M
Membrane depolarization
10- 8 M
Eells et al. (1992), Rekling and Theophilidis (1995)
Voltage-gated calcium channels GABA-gated chloride channels
Brooks and Clark (1987)
<10- 7 M 10- 7 M
Hagiwara et al. (1988) Lawrence et al. (1985)
10- 7 M 10- 7 M
Sherby et al. (1986)
Mitochondrial complex I Lymphocyte proliferation
10-6 M
Diel et al. (1998)
Mitochondrial ATPase
10- 5 M
Prasada Rao et al. (1984)
Intercellular gap junctions
10- 5 M 10- 4 M
Hemming et al. (1993), Tateno et al. (1993)
Nicotinic receptors
Chromosomal damage
Gassner et al. (1997)
Barrueco et al. (1992)
aNominal pyrethroid concentration needed to produce a significant effect. Actual concentrations may, however, vary with experimental conditions.
confers pore function, resembles those of other voltage-gated ion channels, and can take several possible isoforms; and the fh and fh subunits, which may be absent in nonmammalian channels and which modify the basic function of the Cl subunit. There are many variant forms of the Cl subunit, 10 being characterized in the rat, and channels are also subject to glycosylation and phosphorylation, which further modify function. The channel is highly ion selective, favoring sodium over potassium ions by a factor of 30, yet maintains a relatively high conductance of 10-25 pS. Depolarization-driven activation [via the m-gate of Hodgkin and Huxley (1952)] is sensed by the four S4 segments in the Cl subunit and probably mediated by the S6 segment. Time-dependent inactivation of the channel [via the h-gate of Hodgkin and Huxley (1952)] is less well defined in molecular terms, but is probably related to the cytoplasmic domain between the S3 and S4 segments of the Cl subunit. Unfortunately, a standardized nomenclature for the many channel isoforms has not been accepted, and descriptions based on pharmacological properties (e.g., tetrodotoxin resistant) or tissue source (e.g., brain I, 11, Ill) are widely used. Expression is controlled by many different genes. The interaction of pyrethroids with the sodium channel has the effect of slowing both the activation and the inactivation properties of the sodium channel (Ginsburg and Narahashi, 1993), leading to a stable hyperexcitable state. This effect is amplified by the high level of expression of sodium channels in most excitable cells, which means that only about 0.1 % of sodium channels need to be modified by a pyrethroid in order for the extra current generated to render a cell hyperexcitable (Song and Narahashi, 1996b). Although activation is slowed at the single channel level, this high density of sodium channels means that sufficient unmodified channels are always present to ensure that the activation phase of the action potential is not appreciably delayed. However, in the falling phase of the action potential, even a low proportion of modified channels can
generate enough extra current to delay inactivation. This slower rate of inactivation of pyrethroid-modified channels generates a prolonged depolarizing "tail" current that follows the normal action potential. This "tail" can trigger a second action potential if the current is both large enough and lasts for the 0.5-1 ms needed for the unmodified sodium channels to recover excitability. Hence, what would normally be a single-action potential can become converted into double or continuous discharges (Fig. 59.1). This produces a profound disruption of neuronal function. Action potential amplitude generally remains constant, the effect of pyrethroids in delaying the recruitment of open channels after depolarization (which would decrease action potential amplitude) and in delaying the closing of these same channels with time (which would increase ampli-
_----------------1,2 ~ ~4 Jt---------' lis W
~~~--~~~i~[~--~~--~~ Figure 59.1 Repetitive firing in the electromyogram (EMG) and increased muscle twitch (tension) induced by deltamethrin (3 mglkg iv) in rat muscle are only seen if sufficiently long intervals are allowed for recovery between stimuli. At the shorter intervals (lower records), the abnormal repetitive activity is completely lost and the muscle twitch also becomes normal. The interstimulus interval is indicated on the left of each record (in seconds) and the time course of the individual responses on the horizontal axis (1- and IO-ms markers).
59.1 Mechanisms of Toxicity
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Table 59.2 Prolongation of Sodium Channel Currents by Type I and Type 11 Pyrethroids Time constant (ms)
Pyrethroid (class)
System
Control
TTX-sensitive rat cells
1.3
Tetramethrin (I)
TTX -sensitive rat cells
7.1
Fenvalerate (11)
TTX-sensitive rat cells
Control
TTX-resistant rat cells
1.9
Tetramethrin (I)
TTX-resistant rat cells
3.2
Fenvalerate (11)
TTX -resistant rat cells
Permethrin (I)
Frog nerve fibefl
Cismethrin (I)
Frog nerve fibefl
112
388
7.3
Song et al. (1996) Song et al. (1996) Song et al. (1996) Song et al. (1996) Song et al. (1996) Song et al. (1996) Vijverberg and de WeiIle (1985)
21
Vijverberg and de Weille (1985) Vijverberg and de Weille (1985)
Fenvalerate (11)
Frog nerve fibefl
463
Deltamethrin (11)
Frog nerve fibefl
1772
Fenfluthrin (I)
Rat skeletal musclea Rat skeletal musclea
Cismethrin (I)
Reference
Vijverberg and de Weille (1985)
2.3
Wright et at. (1988)
5.8
Wright et at. (1988)
9.3
Wright et at. (1988)
Fenvalerate (11)
Rat skeletal musclea Rat skeletal musclea
14
Wright et al. (1988)
Deltamethrin (11)
Rat skeletal musclea
33
Wright et al. (1988)
Cyphenothrin (IIII)
aTime constant of the abnormal pyrethroid-induced current.
tude), cancelling out in most systems (Ginsburg and Narahashi, 1993). Another significant feature is that, after modification by pyrethroids, sodium channels retain many of their other normal functions, such as their selectivity for sodium ions and conductance. Their link with membrane potential is also retained, although shifted so as to increase excitability (Narahashi et aI., 1998). This means that, after exposure to moderate doses of pyrethroids, cells can still continue to function-but in a new and relatively stable state of abnormal hyperexcitability. In insects, this state corresponds to the incapacitating but sublethal level of effect known as "knock down." The amplitude of the sodium current continues undiminished until the level of sodium entry associated with this hyperexcitability overwhelms the capacity of the sodium pump to remove it (Narahashi, 1985). Very high concentrations of pyrethroids, or levels of hyperactivity beyond those that the cell can sustain, can thus cause depolarization and conduction block (Vijverberg and de Weille, 1985). This depolarization is more readily produced by those pyrethroids that hold the sodium channel open longest. Hence, although 2-Hz stimulation presents no difficulty for normal rat skeletal muscle, the prolonged repetitive discharges and enhanced contractions induced by deltamethrin cannot be sustained at this rate (Fig. 59.1), although they are sustained at lower frequencies. An important characteristic of this pyrethroid-generated tail current is that the amplitude and duration are independent. The current amplitude is dependent only on the proportion of sodium channels modified and hence shows a sigmoidal relationship with pyrethroid concentration or dose. The current duration is dependent only on the pyrethroid structure, some pyrethroids, such as permethrin, holding the channel open for
relatively short times and others, such as deltamethrin, holding it open for much longer (Table 59.2). Individual pyrethroids thus generate a characteristic time constant for prolongation of the sodium channel tail current that is virtually independent of dose (Brown and Narahashi, 1992). There is a continuous distribution of time constants across the range of pyrethroid structures, with type I pyrethroids producing the shorter time constants and type 11 pyrethroids producing the longer time constants. This holds true in both insects and amphibia, where the time constants vary from tens of milliseconds to seconds (Vijverberg and de Weille, 1985), and in mammals, where the absolute time constants are rather shorter (Table 59.2). In all cases, the threshold of hyperexcitability is reached once the decaying tail current remains above the threshold to precipitate a second abnormal action potential for as long as the time needed for excitability to return. This can be achieved by increasing either the amplitude or the duration of the tail current. The different forms of the sodium channel show differential sensitivity to pyrethroids. Pyrethroids act most readily on the tetrodotoxin-resistant subtype of the sodium channel (Song and Narahashi, 1996a), which is expressed in the developing mammalian brain and in the adult dorsal root ganglia. The tetrodotoxin-resistant channels were 10 times more sensitive than tetrodotoxin-sensitive channels in the same cells (Ginsburg and Narahashi, 1993). Insect sodium channels are 100 times more sensitive than the rat brain lIa channel (Warmke et aI., 1997), which explains in part the resistance of mammals. The rat brain lIa form of the sodium channel is sensitive to type 11, but not type I pyrethroids, an effect enhanced 20-fold by the presence of the accessory f31 subunit (Smith and Soderlund, 1998). Channels expressing only the a subunit are capable of showing all of the the characteristics of pyrethroid modi-
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CHAPTER 59
Pyrethroid Insecticides Permethrin toxicity
3.5 3.0
0;
0
2.5
0
.:..:
S 0
2.0
Q
1.5
It)
~
0 0
1.0 0.5
0
0.0 0
20
40
60
80
100
% cis isomer Figure 59.2 Effect of isomeric composition on the oral LDso of permethrin to rats. Data taken from WHOIIPCS (1990).
fication, but require relatively high pyrethroid concentrations of lO-8_lO-7 M (Trainer et aI., 1997). It has been proposed that some of the regional selectivity of action of the pyrethroids parallels the distribution of sensitive sodium channel subtypes, although there are at present only limited data to support this attractive idea. Thus, striatal brain slices showed neurotransmitter release in response to type 11 pyrethroids that was not shown by hippocampal slices of similar capacity (Eells and Dobocovich, 1988). When these brain regions were examined in vivo, the striatum is the first site to show electroencephalogram (EEG) discharges and hyperexcitability (Ray, 1980), whereas the hippocampus showed only enhanced inhibition (Joy et aI., 1989). A similar differential was seen when the activation of earlyresponse genes by mild pyrethroid intoxication was examined, hippocampus showing very little response, but specific cortical, thalamic, and hypothalamic areas showing marked activation (Hassouna et aI., 1996). Unfortunately, the distribution of the different sodium channel subtypes across these brain regions is not yet known. Peripheral nerve (SNSIPN3) sodium-channels are highly sensitive to pyrethroids, especially to type 11 compounds, which produce effects at lO-9 M (Soderlund et aI., 2000), and action at these channels may be relevant to the production of paresthesia. Unfortunately, all such studies are difficult to interpret at present because it has proved difficult to reproduce the high pyrethroid sensitivity of in vivo systems in single-channel electrophysiological experiments. Hence, it is not possible to rank the various sodium channel types in terms of absolute sensitivity, other than when they are co-expressed in the same test system. Pyrethroid action on the sodium channel shows a marked stereospecificity that results in some isomers being far more toxic than others (Soderlund, 1985). For example, the lR and IS cis isomers bind competitively to one site, and the lR and IS trans isomers bind noncompetitively to another (Narahashi, 1986). The IS forms do not modify the channel function but do block the effect of the 1 R isomers. In whole mammals, the 1R isomers are thus active and the 1S isomers inactive and essentially nontoxic. Isomerism at the third carbon of the cyclo-
propane ring gives cis and trans isomers that show insecticidal activity (Elliott et aI., 1978) but differential mammalian toxicity, with the cis isomers being about lO times more potent than the trans ones (Gray, 1985). A final chiral center is generated if a cyano substituent is added to the alcohol, giving eight possible isomers. Again, this affects potency, witn only the a-S and not the a-R forms being toxic to both insects and mammals. This stereospecificity has been exploited in the synthesis of pure isomers such as deltamethrin to produce a remarkable degree of selective toxicity (Glomot, 1982). A practical consequence of this is that the toxicity of products such as permethrin, which are commonly sold as mixtures, can vary from batch to batch. Thus, the rat oral LDso value of commercial permethrin varies from 430 to 8900 mg/kg (FAOIWHO, 1980). This is illustrated in Fig. 59.2, which shows that the toxicity of permethrin samples is largely determined by their cis isomer content. 59.1.2 OTHER ACTIONS Many target sites other than the sodium channel may be relevant to poisoning. These are summarized in Table 59.1. To put these into context, the tissue concentration of pyrethroids in the brain during intoxication varies from 1 to 30 nmol/g tissue (Anadon et aI., 1996; Rickard and Brodie, 1985) in rats, although the free concentration is probably lower. The complex nature of the effects of pyrethroids on the central nervous system has led various workers to suggest that they also act via antagonism of y-aminobutyric acid (GABA)-mediated inhibition, modulation of nicotinic cholinergic transmission, enhancement of noradrenaline release, or direct actions on calcium or chloride ion channels. However, because neurotransmitter-specific pharmacological agents offer only poor or partial protection against poisoning (see Section 59.5), it is likely that these other effects represent only secondary mechanisms of action of the pyrethroids. Indeed, most neurotransmitter release is secondary to increased sodium entry (Eells et aI., 1992). Action on the voltage-dependent chloride channel has also been proposed as a target of pyrethroids (Forshaw and Ray, 1990). Voltage-sensitive chloride channels are found in brain, nerve, muscle, and salivary gland, and their function is to control cell excitability, chloride and sodium conductance having reciprocal effects on membrane excitability (Adrian and Marshall, 1976). Unlike the sodium channel, many functionally different kinds of chloride channels are seen, but the channels that have been shown to be sensitive to pyrethroids belong to the maxi chloride channel class (Franciolini and Petris, 1990). Maxi channels have not been characterized at the molecular level, but are activated by depolarization, have high conductance, are calcium-independent, and are inactivated by protein kinase C phosphorylation (Forshaw and Ray, 1993). A similar, but calcium-dependent maxi channel is unaffected by pyrethroids. Type 11 pyrethroids decrease chloride channel currents at low enough concentrations to be relevant to mammalian poisoning (Ray et aI., 1997), and reduce chloride current in vagus nerve removed from rats during poisoning (Forshaw and
59.1 Mechanisms of Toxicity Ray, 1990). The pyrethroid-induced decrease in maxi chloride channel current is brought about by a fall in open-state probability, which serves to increase excitability and therefore would synergize pyrethroid actions on the sodium channel. Of the few pyrethroids that have been tested, only those producing type 11 poisoning seem to affect maxi chloride channels (Ray et aI., 1997). Because agents such as ivermectin and pentobarbitone, which open chloride channels, have antagonist actions on pyrethroid-evoked salivation, choreoathetosis, and repetitive firing in skeletal muscle (Forshaw et aI., 2000), it seems likely that chloride channel actions contribute to most components of the type 11 poisoning syndrome. Indeed, an action on voltagegated chloride channels may play a major part in the generation of the salivation and myotonia, although its role in the centrally generated signs of poisoning appears to be limited to synergizing the primary action of pyrethroids on the sodium channel. Voltage-dependent calcium channels have been proposed as a target of pyrethroids and are good candidates in insects, with primary effects on T-type calcium currents being seen at 10-7 M (Duce et a!., 1999). However, some mammalian calcium channels appear to be less sensitive, with tetramethrin producing 75% block ofT and 30% block ofL currents in neuroblastoma cells only at 10-5 M, and type 11 pyrethroids having no effect (Narahashi, 1988). In contrast, rabbit sinoatrial node cells were far more sensitive, 10- 7 M tetramethrin (a type I pyrethroid) producing complete block of T current, but having no effect on L current until 10-5 M (Hagiwara et a!., 1988). Unfortunately, the effect of lower concentrations was not reported. The action on voltage-gated calcium channels is interesting as one of the rare instances of a type I pyrethroid-specific effect. At relatively high concentrations, pyrethroids can also act on GABA-gated chloride channels (Bloomquist et a!., 1986), and this effect may contribute to the seizures that accompany severe type 11 poisoning. Several other reports have suggested a role for the GABAA receptor-ionophore complex in components of type 11 pyrethroid toxicity. Deltamethrin inhibits binding of ligand to the mammalian GABA complex by 50% at 10-7 M (Lawrence et a!., 1985), and GABA-stimulated chloride flux is reduced by 72% by 10- 6 M cis-cypermethrin (Abalis et a!., 1986). As with voltage-gated chloride channels, these effects are specific for type 11 pyrethroids. Contrary evidence against a major GABA antagonist effect of pyrethroids is that deltamethrin does not reduce GABAA -mediated hippocampal inhibition (Joy and Albertson, 1991), nor does the in vivo toxicity of deltamethrin correlate well with its potency in inhibiting GAB A-induced 36CI influx (Ramadan et aI., 1988). An even more marked differential is seen in invertebrates, with sodium effects at 10- 12 M and GABA effects at 10-6 M (Chalmers et a!., 1987). Pyrethroids, however, potentiate pentylenetetrazole convulsions by interaction with benzodiazepine binding sites (Devaud et a!., 1986) at more reasonable doses, and this may indicate a potential for action via GABA. Similarly, type 11 pyrethroids show dose additivity with GABA antagonists in terms of acoustic startle (Crofton and Reiter, 1987). An additive effect on startle response shows that both GABA antagonists and pyrethroids induce hyperexcitable states, although
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it does not prove a common mechanism. Indeed, deltamethrin also enhances tryptamine and (to a lesser extent) strychnine toxicity as well as that of the GABA antagonist pentazol (Chanh et aI., 1984). A further problem for the GABA hypothesis is that the choreoathetosis, which is seen at lower doses than those that evoke seizures, has been shown to be of spinal origin (Bradbury et a!., 1983). GABA is not an important neurotransmitter in the spinal cord, and Ogata et a!. (1988) have shown that deltamethrin does not block the GABAA -gated chloride current in dorsal root ganglia. Of pharmacological agents acting on the GABA receptor, baclofen had no therapeutic properties in deltamethrin-poisoned rats (Bradbury et aI., 1983). GABA antagonists were ineffective at preventing the increased hippocampal inhibition produced by deltamethrin (Joy and Albertson, 1991). Diazepam (2 mg/kg iv) was ineffective at controlling deItamethrin choreoathetosis in rats (Cremer et a!., 1980), although it does prevent the seizures associated with late-stage poisoning (Gammon et al., 1982). Diazepam was also ineffective in fenvalerate-poisoned humans (He et a!., 1989). It appears that the undoubted potential of type 11 pyrethroids to act at the GABA receptor is of limited practical significance other than in cases of severe intoxication. The peripheral benzodiazepine receptor has also been proposed as a target of pyrethroids, Devaud et a!. (1986) finding that the proconvulsant action of low doses of both type I and type 11 pyrethroids were blocked by the peripheral benzodiazepine receptor antagonist PKll195. This antagonist also reduced the inhibitory effect of iverrnectin on deltamethrinevoked salivation in rats (Forshaw et a!., 2000), suggesting that some component of the salivation is mediated via the peripheral benzodiazepine receptor. Competitive binding at both central and peripheral-type benzodiazepine receptors has been shown to occur in rat parotid and submandibular glands (Yamagishi and Kawaguchi, 1998). The mechanism whereby pyrethroids interact with ion channels is not known, but type 11 pyrethroids stimulate protein kinase C-dependent protein phosphorylation at as Iowa concentration as 10- 13 M in vitro by a direct mechanism (Enan and Matsumura, 1993). Because sodium and chloride ion channel activity is modulated by the phosphorylation state, this is likely to be a very important mechanism of action. Pyrethroids are capable of acting directly in systems with no phosphorylation capacity, but at somewhat higher concentrations (Forshaw et aI., 1993). Pyrethroids have no direct anticholinesterase activity (Ray and Cremer, 1979) and have little effect on acetylcholine sensitivity of muscle (Sherby et aI., 1986), but do inhibit acetylcholine-activated calcium flux at 10- 7 M and slow the desensitization of the nicotinic receptor (Sherby et a!., 1986). This may lead to some degree of potentiation of nicotinic transmission in the central nervous system (CNS). The marked increase in neocortical blood flow seen during pyrethroid intoxication is cholinergically mediated (being blocked by atropine or lesioning the cholinergic innervation), but this probably represents an indirect effect of pyrethroids on the cholinergic system (Lister and Ray, 1988).
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A more sensitive effect is the simulation of noradrenaline release from brain synaptosomes, which has an ED50 of only 2.9 x 10-9 M for deltamethrin, although type I pyrethroids were ineffective (Brooks and Clark, 1987). Cultured cells proved much less sensitive, requiring 10-5 M (Bickmeyer et aI., 1994). This noradrenaline release parallels increased calcium entry and may indicate an action via voltage-sensitive calcium or sodium channels. Related effects are seen in whole animals, where deltamethrin markedly increases plasma noradrenaline (Cremer and Seville, 1982) and enhances the noradrenalinemediated contraction of mesenteric blood vessels and cardiac contractility (Forshaw and Bradbury, 1983). It has been suggested that the calcium effects may be mediated by an action on calmodulin, but calmodulin-stimulated phosphodiesterase activity is inhibited by pyrethroids only at 10-6 _10- 4 M (Rashatwar and Matsumara, 1985). Narahashi (1986) has also described a decrease in fast calcium current in neuroblastoma cells at concentrations similar to those prolonging sodium current. The significance of this is not yet clear, but it may be related to the depression of neurosecretion measured in vitro after in vivo dosing of pyrethroids (Dyball, 1982). Other biochemical changes are the marked increases in cerebellar cyclic GMP (glutamine monophosphate) seen during type 11 intoxication (Aldridge et aI., 1978). This effect would appear to be entirely secondary to the increased motor activity, however (Brodie, 1983), and could not be reproduced in cerebellar slices (Lock and Berry, 1981). Similarly, there are changes in the concentration of several neurotransmitters in the brain, which parallel the motor symptoms (Aldridge et aI., 1978; Hudson et al., 1986) but which are probably only of secondary importance. Pyrethroids have been shown to inhibit mitochondrial Mg2+ -ATPase activity, but only at 2 x 10-5 M (Prasada Rao et aI., 1984). Just subtoxic levels of type 11 pyrethroids have a tumor promoter-like activity in rat liver, although not acting as tumor initiators (Hemming et aI., 1993). This action has been attributed to an inhibition of gap-junction communication, but the high pyrethroid concentrations needed to produce this effect in vitro (Table 59.1) cast doubt on the hypothesis (Hemming et aI., 1993; Tateno et aI., 1993).
59.2 SYSTEMIC POISONING Fortunately, there have been relatively few reports of systemic poisoning, because use of adequate protective clothing prevents intoxication even under tropical conditions (Moretto, 1991). However, systemic poisoning can occur under conditions of misuse or inadequate user protection (He et aI., 1989), and, in a study of sprayers where dermal contamination was poorly controlled, 0.3% showed signs of mild pyrethroid poisoning (Chen et al., 1991). When systemic toxicity does occur, the central signs of poisoning can be difficult to control and may be confused with intoxication by other pesticides such as anticholinesterases, which also cause salivation and hyperexcitability. Given the common formulation of pyrethroids with volatile
solvents such as xylene, symptoms of poisoning can be complicated by solvent toxicity, and solvents may also introduce additional skin effects. Mild poisoning symptoms may also be amplified by anxiety (Lessenger, 1992), which may itself be precipitated by fear or by the disconcerting paresthesia resulting from dermal contact with pyrethroids (Flannigan et aI., 1985). In some low-exposure cases, poisoning may be more apparent than real. Thus, of 64 cases of self-diagnosed pyrethroid intoxication following very low level exposure to a variety of pyrethroids, 58 showed either no somatic effects or ones that were unrelated to pyrethroid exposure, and 6 showed reversible changes that were of ambiguous etiology. No CNS or PNS (peripheral nervous system) lesions were found in any of these cases (Altenkirch et aI., 1996). Such cases are in marked contrast with frank occupational poisonings (He et aI., 1989) and may relate more to public perception of hazard than to real toxicity. The blood half-life of pyrethroids is on the order of tens of hours (Anadon et al., 1996; Gray et aI., 1980), and intoxication by the oral route is correspondingly short lasting. Some investigators have found even shorter half-lives. Thus, cyfiuthrin was found to have a plasma half-life of 19-86 min in humans (Leng and Lewalter, 1999). Inherent toxic potential can be high though, as intravenous LD50S range from more than 250 to 0.5 mg/kg (Ray, 1991), but toxicity is limited in practice by rapid hydrolysis in blood and liver. Toxicity by the dermal route is further limited by low absorption through the skin and the capacity for dermal metabolic destruction of pyrethroids (Bast et aI., 1997). In humans, the bioavailability of dermal pyrethroids is about 1% (Woollen et aI., 1992), compared to 36% for gastric absorption. Hence, the dermal route of exposure presents relatively little risk of systemic poisoning although, in cases of very severe skin contamination, intoxication has lasted for several weeks (He et aI., 1989), possibly due to a reservoir of pyrethroid bound to the epidermis. Once absorbed, pyrethroids are rapidly distributed throughout the body, their high lipophilicity and lack of exclusion by the multidrug transporter glycoprotein (Bain and LeBlanc, 1996) ensuring ready entry into the brain (Gray et al., 1980). All of the motor signs of systemic pyrethroid intoxication are generated at the spinal level, being relatively unchanged by destruction of the brain (Bradbury et al., 1983), although other brain areas show a range of secondary responses to intoxication. Recording from central and peripheral sites shows that a modest level of hyperactivity in sensory fibers is amplified at the first synapse and becomes further amplified in polysynaptic pathways (Forshaw et aI., 1987). Animal studies have shown that the major neurotoxic hazard presented by pyrethroids to adults is acute excitation. Nearlethal doses of all classes of pyrethroids can give rise to an axonal degeneration in peripheral nerve closely resembling Wallerian degeneration (Aldridge, 1990), but this effect is inherently reversible and is only seen at dose levels that produce prolonged and severe motor signs. Central neuropathology has been described in one study of adult rats given 15 daily doses of the type 11 pyrethroid deltamethrin at doses just be-
59.2 Systemic Poisoning Table 59.3 Signs of Pyrethroid Poisoning and Classification of Some Pyrethroids Type I (T) poisoning"
Type II (CS) poisoning" Profuse watery salivation
Severe fine tremor
Coarse tremor Increased extensor tone
Marked reflex hyperexcitability
Moderate reflex hyperexcitability
Sympathetic activation
Sympathetic activation Choreoathetosis Seizures
Paresthesia (dermal exposure)
Paresthesia (dermal exposure)
Type I pyrethroids
Type II pyrethroids
Permethrin Pyrethrins b
Cyhalothrin
FenvaIerate
Bioallethrin
Deltamethrin
Cismethrin
Cypermethrin
a As
seen in the rat. bPyrethrins are the natural plant product, as distinct from the synthetic pyrethroids.
low the threshold for motor signs (Husain et aI., 1994). The animals were stated to show 40% increases in the weight of pons/medulla and hippocampus, apparently without morphological change, and also degeneration of cerebellar Purkinje cells. These findings do not appear to be internally consistent, and their significance must be considered questionable. Other workers have found no central pathology associated with the maximum survivable dose of pyrethroids (Holton et aI., 1997). Despite the continuous variation in duration of the abnormal sodium current with pyrethroid structure described in Section 59.1, the effects of all pyrethroids can be described in terms of just two well-demarcated poisoning syndromes (Table 59.3) that are seen in both mammals and insects. The longer current prolongations are far more disruptive, and pyrethroids with time constants of more than about 10 ms (the normal time constant of the unmodified sodium channel being about 0.5 ms) cause incoordination, choreoathetosis, seizures, and direct effects on skeletal and cardiac muscle and salivary gland. Reflex hyperexcitability can be seen, as a dose-dependent combination of enhancement and suppression (Hijzen and Slangen, 1988; Wright et aI., 1988). This is called the type 11, or choreoathetosis/salivation, syndrome (Aldridge, 1990). Those pyrethroids (Table 59.2) producing shorter prolongations of less than about 10 ms in mammals (Wright et aI., 1988) cause a less complex syndrome of simple reflex hyperexcitability and fine tremor, which is termed the type I, or tremor, syndrome, and closely resembles that produced by dichlorodiphenyltrichloroethane (DDT) (Aldridge, 1990). Pyrethroids producing intermediate time constants of about 10 ms produce a complex, mixed syndrome in which the characteristic signs of both type I and type 11 poisoning are superimposed (Wright et al., 1988). Almost all reports of human poisoning relate to the more potent type 11 pyrethroids, so it is not certain how well this description of
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the two syndromes applies to humans, although what has been described fits quite well with the earlier experimental animal observations. A similar division into two classes has been made by a number of authors for systemic effects in insects (Clements and May, 1977; Gammon et aI., 1981; Salgado et al., 1983), in amphibians (Ruigt and van den Bercken, 1986), and in mammals (Barnes and Verschoyle, 1974; Verschoyle and Aldridge, 1980; Wright et aI., 1988). Generally speaking, there is fairly good agreement across species (Wright et aI., 1988), with most noncyano substituted pyrethroids falling into group I and most of the cyano pyrethroids into group 11. Some authors have objected to such a division on the grounds that qualitatively similar electrophysiological changes are produced by all active pyrethroids in both invertebrates (Leake et aI., 1985) and mammals (StaatzBenson and Hosko, 1986) or because some pyrethroids show intermediate characteristics (Gammon et aI., 1981; Verschoyle and Aldridge, 1980). All type 11 pyrethroids have a cyano substituent, but not all type I pyrethroids lack one, the trans and cis isomers of flurocyphenothrin somewhat confusingly producing type I and 11 effects, respectively. Both pyrethroid classes have a similar range of mammalian toxicity but, for commercial pesticides, the type 11 pyrethroids such as deltamethrin and cypermethrin are generally more toxic than the type I pyrethroids such as permethrin. A classification of the more common pyrethroids is given in Table 59.3. Both type I and 11 pyrethroids cause marked adrenal activation in rats, probably by a direct stimulation of noradrenaline release (Brooks and Clark, 1987), with the increases in blood adrenaline and noradrenaline accompanying motor signs (Cremer and Seville, 1982). The type 11 pyrethroid, deltamethrin, causes increased corticosteroid secretion at even lower doses (de Boer et aI., 1988). Hence, it is important to remember that even moderate pyrethroid poisoning occurs against a background of profound adrenal activation. Adrenal activation has been proposed as the mechanism whereby low doses of fenvalerate reduce conditioned avoidance behavior in rats (Moniz et al., 1994). 59.2.1 TYPE I POISONING The type I pyrethroids produce the simplest poisoning syndrome and produce sodium tail currents with relatively short time constants (Wright et aI., 1988). Poisoning closely resembles that produced by DDT and was first clearly distinguished by Verschoyle and Aldridge (1980). It involves a progressive development of fine whole-body tremor, exaggerated startle response, incoordinated twitching of the dorsal muscles, hyperexcitability, and death (Ray, 1982b). The tremor can be so severe as to double the whole-body metabolic rate in the rat at sublethal dose levels (Cremer and Seville, 1982) and can lead to prostration and death. At sublethal dose levels, respiration and blood pressure are well sustained (Ray, 1982b), but plasma noradrenaline, lactate, and, to a lesser extent, adrenaline are greatly increased (Cremer and Seville, 1982). Type I effects
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are generated largely by action on the central nervous system, as shown by the good correlation between brain levels of cismethrin and tremor (White et aI., 1976) and the induction of tremor by small quantities of cismethrin directly injected into the CNS (Gray and Rickard, 1982; Staatz et aI., 1982). Poisoning is associated with marked increases in both spinal (Carlton, 1977; Staatz-Benson and Hosko, 1986) and brain stem (Forshaw and Ray, 1986) excitability although not with marked effects on the higher centers. Thus, lethal doses of cismethrin do not induce cortical EEG spiking (Ray, 1982a), although supralethal doses in paralyzed, ventilated animals do (Staatz and Hosko, 1985). Also, when cismethrin is injected into the lateral ventricles, tremor is seen only when enough is given to reach the brainstem (Gray and Rickard, 1982) and, although primary increases in reflex excitability are seen in the brainstem and spinal cord, only secondary effects are seen at the cerebellar, thalamic, and cerebral cortical levels (Ray, 1982a). In addition to these central effects, there is evidence for repetitive firing in sensory nerves, which, although a small effect in one study (StaatzBenson and Hosko, 1986), was more pronounced in others (Forshaw and Ray, 1986; Wright et aI., 1988). Such repetitive firing is analogous to that seen in amphibians (Vijverberg et aI., 1982) and probably contributes to the hyperexcitable state produced by the central actions of the type I pyrethroids. 59.2.2 TYPE 11 POISONING
The type 11 pyrethroids produce a more complex poisoning syndrome and act on a wider range of tissues. They give sodium tail currents with relatively long time constants (Wright et aI., 1988), which may be the reason for their ability to act on the whole range of excitable tissues. Human type 11 poisoning seems to be characterized by paresthesia (if via the dermal route), dizziness, nausea, listlessness, and muscular fasciculations (Chen et aI., 1991). More severe poisoning caused epigastric pain, nausea and vomiting (if via the oral route), hypersalivation and pulmonary edema, opisthotonos, seizures, and coma (He et aI., 1989). First noted by Barnes and Verschoyle (1974), type 11 poisoning in rats involves progressive development of nosing and exaggerated jaw opening, similar to that seen in response to an irritant placed on the tongue; salivation, which may be profuse; increasing extensor tone in the hind limbs, causing a rolling gait; incoordination progressing to a very coarse tremor; choreoform movements of the limbs and tail often precipitated by sensory stimuli; generalized choreoathetosis (writhing spasms); tonic seizures; apnea; and death (Ray, I 982b). At lower doses, more subtle repetitive behavior is seen (Brodie and Aldridge, 1982) and learned behavior is impaired (Moniz et aI., 1994). In dogs, similar symptoms are seen, but salivation and upper airway hypersecretion and gastrointestinal symptoms are more prominent (Thiebault et aI., 1985). Unlike the type I pyrethroids, type 11 pyrethroids generally decrease rather than increase the startle response to sound (Crofton and Reiter, 1984), although this is a complex response and at low doses
some type 11 pyrethroids give an increased startle (Hijzen and Slangen, 1988). The cerebral cortical response to sound is also depressed (Ray, 1980) and the latency of the visual response increased (Dyer, 1985). As in type I poisoning, plasma noradrenaline is increased by type 11 pyrethroids, but there is also a large increase in adrenaline and in blood glucose (Cremer and Seville, 1982; Ray and Cremer, 1979), which is not seen in type I poisoning. Type 11, but not type I, pyrethroids also increase cardiac contractility both directly by action on cardiac muscle and via circulating and locally released catecholamines (Forshaw and Bradbury, 1983). They also cause repetitive firing and potentiate contraction in skeletal muscle (Forshaw and Ray, 1986). These effects are limited to a large degree by physiological compensation, which maintains blood pressure at normal levels in intact rats despite some arrhythmias, and by the inability of skeletal muscle to sustain repetitive firing for more than a few seconds at normal discharge frequencies (Fig. 59.1). Both effects are therefore of secondary importance in the normal animal. As with type I pyrethroids, the primary action is on the central nervous system, because symptoms correlate well with brain concentrations (Rickard and Brodie, 1985) and can be reproduced in part by microinjection into the CNS (Brodie, 1985; Staatz et aI., 1982). The former injection studies showed, however, that actions at all levels of the neuroaxis are needed to reproduce the full range of effects. Thus, although choreoathetosis can be reproduced in spinal rats (Bradbury et aI., 1983), other symptoms are generated at higher levels and are associated with EEG spiking at cortical and subcortical sites, which ultimately progresses to slow-wave activity and loss of consciousness not seen with type I pyrethroids (Condes Lara et aI., 1999; Ray, 1980). By this stage, many neurotransmitter systems have become involved, because both specific and nonspecific pharmacological interventions can control the seizures. Although there is evidence for increased neuronal activity in both the spinal cord (Staatz-Benson and Hosko, 1986) and the brain (Ray, 1980; Staatz and Hosko, 1985), the type 11 pyrethroids do not produce the repetitive activity in sensory nerves seen with type I pyrethroids (Wright et aI., 1988). This again is analogous to the effects seen in amphibians (Vijverberg et aI., 1982). As might be expected, both classes of pyrethroids produce large increases in brain glucose utilization, this being most marked in motor areas (Cremer and Seville, 1985; Cremer et aI., 1983). Such increases seem to be secondary to increased neuronal activity and are paralleled by increased brain blood flow except in the cerebral cortex, where the flow increase is disproportionally large (Lister and Ray, 1988). 59.2.3 MIXEDIINTERMEDIATE POISONING
Complex, mixed signs representing a combination of type I and type 11 poisoning are produced by some pyrethroids. These appear to represent a true superimposition of the I and 11 poisoning syndromes (Wright et aI., 1988) and to represent a transitional state. Evidence in support of this is given by measurement of the time constants of the sodium after-potential produced by
59.4 Developmental Neurotoxicity the pyrethroids. These are short for the type I and long for the type IT pyrethroids (Table 59.2). For the whole range of pyrethroids, time constants range from 5 to 1772 ms in amphibians (Vijverberg and de Weille, 1985) and from 2.3 to 33 ms in mammals (Wright et at., 1988). The structures producing mixed signs fit in the middle of these ranges. The related question of what might happen in the case of simultaneous exposure to a type I and a type II pyrethroid appears not to have been addressed in the whole animal. However, Song et al. (1996) used tetramethrin (~ type I pyrethroid) and fenvalerate (a type IT pyrethroid) and compared their actions on tetrodotoxin-resistant sodium channels in isolated neonatal rat dorsal root ganglion cells. The measure used was the degree of prolongation of the time constant, which reflects the nature of the pyrethroid effect and is more or less independent of dose. If tetramethrin was added to preparations showing a characteristic fenvalerate response, this disappeared and was replaced by a response very similar to that generated by tetramethrin alone. When the preparation was washed, it reverted to a fenvalerate response (the fenvalerate response being poorly reversible in this system). Although these data should be interpreted with some caution because both pyrethroids were used at very high nominal concentrations, the result implies that type I and II actions may be mutually exclusive.
59.3 PARESTHESIA AND LOCAL IRRITATION In addition to systemic toxicity, pyrethroids can produce important local effects: skin contamination producing paresthesia, ingestion producing gastrointestinal irritation (Thiebault et al., 1985), and inhalation upper respiratory tract irritation (Pauluhn and Machemer, 1998). The gastrointestinal irritation is rare (being limited to cases of ingestion) and has not been well studied, but is presumably a similar phenomenon to the more common dermal paresthesia. Respiratory tract irritation can be produced at comparable thresholds in rats and humans, but is rarely reported. Dermal exposures far below the threshold for systemic poisoning can lead to a local paresthesia, which is evoked by all classes (the pyrethrins and type I and II pyrethroids), with a severity roughly in proportion to their systemic toxic potential (Aldridge, 1990). Although the plant products associated with impure pyrethrum extracts can give rise to classical contact dermatitis, the pure synthetic pyrethroids produce only a simple paresthesia, not inflammation or erythema (Flannigan et aI., 1985). However, inflammation can be evoked by some pesticide solvents or by scratching. One case of a probable immune reaction to a pyrethroid has been described (Box and Lee, 1996). The simple paresthesia is dose dependent in severity and duration, lasting for 4-30 h after a single application. When mild, the sensation is of continuous tingling or pricking or, when more severe, burning. The effect is annoying but not disabling and does not appear to be associated with any lasting ill effects on nerve conduction (LeQuesne et al., 1980). An animal model of paresthesia has been developed (Cagen et aI.,
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1984). In addition, an electrophysiological test (enhancement of the supernormal phase in two-pulse stimulation) can detect peripheral nerve hyperexcitability up to 24 h after exposure in animals (Parkin and LeQuesne, 1982). This test has also proved useful for monitoring pyrethroid effects in humans (He et al., 1991). The mechanism of paresthesia has not been studied directly, but presumably the sensation is caused by abnormal pyrethroidinduced repetitive activity in skin nerve terminals, which are exposed to higher concentrations than penetrate the rest of the body. Such an idea is supported by the observation that veratrine produces a similar response to the pyrethroids in a guinea pig model of paresthesia (McKillop et al., 1987). Repetitive activity in sensory nerves is certainly seen during systemic intoxication in rats (Wright et al., 1988), but the effect is much less than that seen in the central nervous system, suggesting that nerve terminals are not markedly sensitive to pyrethroids in mammals. The reversible peripheral nerve damage produced by pyrethroids is an unrelated effect, because it appears only when severe systemic poisoning is repeatedly produced (Aldridge, 1990; Rose and Dewar, 1983).
59.4 DEVELOPMENTAL NEUROTOXICITY Neonatal rats are 4-17 times more vulnerable than adults to the acute toxicity of type T and II pyrethroids, probably due to their lesser capacity for metabolic detoxification (Cantalamessa, 1993), an observation that is consistent with neonates and adults having similar brain concentrations at different, but equitoxic, doses (Sheets et al., 1994). A number of more specific effects of exposure to pyrethroids during early development have been described in rats or mice. Cypermethrin at 4% of the LDso over postnatal days 1016 caused an increase in the renal Dj receptor density in rats, which persisted at least until day 9 (Cantalamessa et aI., 1998). The pyrethroids permethrin and deltamethrin [in addition to DDT, polychlorinated biphenyls (PCBs), nicotine, and paraquatl have also been reported to induce permanent changes in behavior and neurochemistry of adult mice when administered directly to the neonate (Eriksson and Fredriksson, 1991; Eriksson et al., 1990a, b, 1992; Fredriksson et aI., 1993; Talts et al., 1998). These effects were seen at dose levels that are not acutely toxic. It should be noted that, at much higher doses, DDT acts directly on sodium channels in the same way as type I pyrethroids (Woolley, 1982), but that the other agents have very different mechanisms of action. It was proposed that the pyrethroid effects resulted from exposure at a critical period of rapid brain growth, during which the developing motor and sensory systems are believed to be especially vulnerable to chemical insult (Eriksson et al., 1990a, 1992). Indeed, neonatal rats are known to be sensitive to neuroactive pharmaceuticals such as haloperidol, postnatal exposure to which causes similarly enduring changes in dopamine receptor-mediated behavior that persist into adulthood (Cuomo et al., 1983; Rosengarten and Friedhoff, 1979; Thiel et al., 1989). However, in these
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taminated skin with oil (Malley et aI., 1985), as pyrethroids bind avidly to skin and cannot be removed by soap and water. Vitamin E cream has been found particularly effective for treating paresthesia in clinical trials (Flannigan et al., 1985; Malley et aI., 1985) and also in some (Song and Narahashi, 1995), but not all (Oortgiesen et aI., 1990) in vitro preparations. The vitamin E was effective if applied to the skin from 29 h before to 15 min after the pyrethroid, and protection lasted for more than 5 h. The concentration required to give greater protection than that of the corn oil solvent alone was very high (50%) and so the specificity of vitamin E protection is unclear. Similarly, in humans, considerable relief can be obtained by use of presumably inert preparations such as corn oil or Vaseline as well as vitamin E (Tucker et aI., 1983). It is not clear how vitamin E produces this effect, but it may in part be due to sequestation of lipophilic pyrethroids into the vitamin E (Song and Narahashi, 1995) or to a more specific membrane-stabilizing action. Topical treatment with local anesthetic has been described in humans and in animals (FAOIWHO, 1982; Malley et aI., 1985), but the loss of sensation may be more inconvenient than the paresthesia itself. Systemic poisoning has proved rather more difficult to deal with although, because pyrethroids produce no morphological damage and are rapidly removed from the body, only symptomatic treatment is needed. Two approaches are possible: attempting to antagonize the primary ion channel effects of the pyrethroids or controlling the secondary consequences mediated by specific chemical neurotransmitters. Pyrethroids do not, however, seem to single out a particular neurotransmission system: Both type I (Staatz et aI., 1982) and type 11 (Chugh et al., 1991) pyrethroids act on multiple neurotransmitter systems. Thus, micromolar concentrations of type 11 pyrethroids evoke release of both dopamine and acetylcholine from brain slices via a sodium channel-dependent mechanism (Eells and Dobocovich, 1988). Both type I and type 11 pyrethroids can act as proconvulsants (Chanh et al., 1984; Devaud et aI., 1986) via GABA-ergic and glutamatergic systems. However, diazepam is only poorly effective against systemic pyrethroids in rats (Staatz et aI., 1982), effective only against the terminal seizures of the type 11 syndrome in another study (Cremer et aI., 1980), and only moderately effective in the dog (Thiebault et aI., 1985). Even the salivation characteristic of type II poisoning, although undoubtedly cholinergically mediated and controllable by atropine in animals (Ray and Cremer, 1979) and in humans (He et aI., 1989), is also controlled very effectively by the chloride ion channel agonist ivermectin (Forshaw et aI., 2000), which has no anticholinergic actions. Consequently, the most successful attempts at therapy have been based on ion channel or membrane-stabilizing drugs. An ideal therapeutic agent would antagonize the abnormal, pyrethroid-evoked sodium current but leave the normal one un59.5 THERAPY FOR PYRETHROID changed. In a survey of the relative ability of a range of drugs to antagonize the pyrethroid-evoked and normal sodium curPOISONING rents in neuroblastoma cells (Oortgiesen et aI., 1990), it was The most commonly encountered sign of pyrethroid poisoning found that tetracaine and lidocaine were particularly effective. is dermal paresthesia. This can be treated by lavage of the con- Lidocaine was also effective at antagonizing pyrethroid effects
cases, the dose required to produce developmental effects in the neonate is close to that producing acute effects in the adult. It has been proposed that the lasting effects of neonatal treatment of mice with bioallethrin and deltamethrin (a decrease in muscarinic receptor density and a decrease in open-field habituation) are a result of an induced lack of appropriate cholinergic inhibitory capacity in the neonate (Eriksson et al., 1992; Rekling and Theophilidis, 1995). However, the acute increases in receptor density (4 or 7% of control) produced by DDT or type I pyrethroids (Ahlbomet aI., 1994; Erikssonetal., 1992) are only modest. They appear to be too small to account for any subsequent maldevelopment, given the large receptor reserve shown by many receptors (Bencherif et aI., 1995; Ek and Antonsson, 1993; Feuerstein et aI., 1994; Zhu, 1993), including central cholinergic systems-unless the changes were concentrated in a specific sUbpopulation. These positive results also contrast with a lack of effect of longer term, higher dose dietary administration of pyrethroids in rat multi generation studies conducted for regulatory purposes (Deshmukh, 1992; Gomes et aI., 1991), except where the dose level is so high as to produce maternal toxicity (Abdelkhalik et aI., 1993). Unpublished work in my own laboratory has reproduced the lasting change in muscarinic receptor density produced by low doses of DDT reported by Eriksson et al. in mice, but not that produced by pyrethroids. The overall physiological significance of this pyrethroidinduced muscarinic receptor deficit clearly needs further investigation. It is interesting that a major type of sodium channel is found in the developing rat brain with a peak density level in brainstem on postnatal days 10-21 and with a higher binding affinity for saxitoxin than in the adult (Xia and Haddad, 1994). This could be a potential site for low-dose pyrethroid and DDT developmental neurotoxicity. However, it appears than that a clarification of the question of the reproducibility and applicability of the effects seen in mice to other species will be needed before any more general conclusions can be drawn with regard to the potential developmental toxicity of pyrethroids in humans. A different effect is the delayed development of the bloodbrain barrier in rat pups given cypermethrin. Adult rats were unaffected at these or higher doses (Gupta et aI., 1999). The minimum dose needed to produce this effect was only 2% of the adult LD50, but represented 20% of the neonatal LD50 (Cantalamessa, 1993), so it is possible that the effect was relatively nonspecific. Vascular damage and delayed neuronal development were also reported in neonatal rats given 0.7 mg/kg/day deltamethrin ip for 5 days (Patro et aI., 1997). This dose produced a marked decrease in pup body weight though, so it is likely in this case that the effects were also nonspecific in origin.
References in intact rat hippocampus (Joy et aI., 1990) and modestly effective against motor signs (Bradbury et aI., 1983), but local anesthetics are not practical for systemic therapy because of their cardiotoxicity. Phenytoin, phenobarbitone, and valproate were found to act equally on the pyrethroid-evoked and normal sodium currents (Oortgiesen et aI., 1990); diazepam, mephenesin, and urethane had less action on the abnormal pyrethroid-evoked current than on the normal one. Hence, although urethane was found to be very effective in vivo (increasing the deltamethrin LD50 by 348%), this was only at a dose close to that producing anesthesia (LeClercq et aI., 1986), a dose at which urethane is also a lung carcinogen. Similarly, mephenesin was only effective against type I poisoning at doses producing marked muscle relaxation but, surprisingly, was effective against type 11 poisoning at rather lower doses. Mephenesin is also one of very few agents capable of antagonizing ongoing pyrethroid poisoning (Bradbury et aI., 1983). The longer lasting mephenesin derivative, methocarbimol, antagonized the motor signs of both type I and type 11 pyrethroids (Hiromori et aI., 1986), although this also required use of the maximum tolerated dose. A novel approach has been to attempt to develop therapy targeted at the voltage-gated chloride channel, a site of action for type 11 (although not type I) pyrethroids (Ray et aI., 1997). The chloride channel agonist ivermectin was effective at restoring the membrane conductance of vagus nerve taken from rats given an LD50 dose of deltamethrin (Forshaw and Ray, 1990). Ivermectin also controlled deltamethrin-induced salivation and repetitive discharges in muscle (Forshaw et aI., 2000). Unfortunately, the central actions of ivermectin are severely limited by multidrug receptor pump activity at the blood-brain barrier, which largely excludes ivermectin, but not pyrethroids, from the brain (Schinkel et aI., 1996). However, pentobarbitone is another effective chloride channel agonist, which increases both channel open probability and channel number in neuroblastoma cells, and does penetrate the blood-brain barrier. Pentobarbitone was effective against all the type 11 motor signs caused by deltamethrin at only 25% of the anesthetic dose. An equisedative dose of phenobarbitone (which does not act on chloride channels) was ineffective against type 11 signs, other than the terminal seizures (Forshaw et aI., 2000), which suggests that pentobarbitone was not just acting as a membrane stabilizer, but had some more specific action-probably on the chloride channels. This anomalous action of pentobarbitone suggests that a combination of a local-anesthetic-type sodium channel blocker and a chloride channel agonist would provide an effective treatment for type 11 pyrethroid poisoning. Although phenobarbitone has been tried and found ineffective as a type 11 pyrethroid antidote in humans (He et aI., 1989), pentobarbitone does not appear to have been tested clinically. Because type 11 poisoning involves a combined action on the central nervous system, adrenals, autonomic system, and muscle, multi drug therapy may be needed. Thus, mephenesin proved effective at blocking all type 11 motor and cardiovascular effects produced by deltamethrin, but had no effect on the salivation or the increased neocortical blood flow, and only
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a modest effect on the (presumably adrenergic) increase in blood glucose (Bradbury et aI., 1983). Combined drug therapies have proved only moderately effective so far: A combination of clomethiazole (membrane stabilizer), diazepam (against seizures), and atropine (against salivation) increased the LDso of deltamethrin by 24% (LeClercq et aI., 1986) in rats. However, because pyrethroids show very steep dose-response curves, even a modest shift such as this can have a dramatic consequence: The combination of methocarbimol and atropine prevented all deaths at LD80 doses of pyrethroids in rats (Hiromori et aI., 1986).
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Food and Agriculture Organization/World Health Organization (FAOIWHO) (1982). "FAO Plant Production and Protection." Paper 42 (Monographs), Food and Agriculture Organization, Rome. Forshaw, P. J., and Bradbury, J. E. (1983). Pharmacological effects of pyrethroids on the cardiovascular system of the rat. Eur. J. Phannacol. 91, 207-213. Forshaw, P. J., and Ray, D. E. (1986). The effects of two pyrethroids, cismethrin and deltamethrin, on skeletal muscle and the trigeminal reflex system in the rat. Pestic. Biochem. Physiol. 25, 143-151. Forshaw, P. J., and Ray, D. E. (1990). A novel action of deltamethrin on membrane resistance in mammalian skeletal-muscle and non-myelinated nervefibers. Neuropharmacology 29, 75-81. Forshaw, P. J., and Ray, D. E. (1993). A voltage-dependent chloride channel in NIEI15 neuroblastoma-cells is inactivated by protcin-kinase-c and also by the pyrethroid deltamethrin. J. Physiol. 467,252. Forshaw, P. J., Lister, T., and Ray, D. E. (1987). The effects of 2 types of pyrethroid on rat skeletal-muscle. Eur. J. Pharmacal. 134, 89-96. Forshaw, P. J., Lister, T., and Ray, D. E. (1993). Inhibition of a neuronal voltage-dependent chloride channel by the type-ii pyrethroid, deltamethrin. Neuropharmacology 32, 105-111. Forshaw, P. J., Lister, T., and Ray, D. E. (2000). The role of voltage-gated chloride channels in type II pyrethroid insecticide poisoning. Toxicol. Appl. Phannacol. 163, 1-8. Franciolini, F., and Petris, A. (1990). Chloride channels of biological membranes. Biochim. Biophys. Acta 1031, 247-259. Fredriksson, A., Fredriksson, M., and Eriksson, P. (1993). Neonatal exposure to paraquat or MPTP induces permanent changes in striatum dopamine and behaviour in adult mice. Toxicol. Appl. Pharmacal. 122, 258-264. Gammon, D. W., Brown, M. A., and Casida, J. E. (1981). Two classes of pyrethroid action in the cockroach. Pestic. Biochem. Physiol. 15, 181-191. Gammon, D. W., Lawrence, L. J., and Casida, I. E. (1982). Pyrethroid toxicology: Protective effects of diazepam and phenobarbitol in the mouse and cockroach. Toxieol. Appl. Phannacol. 66, 290-296. Gassner, B., Wuthrich, A., Scholtysik, G., and Solioz, M. (1997). The pyrethroids permethrin and cyhalothrin are potent inhibitors of the mitochondrial complex I, J. Pharmacal. Exp. Ther. 281, 855-860. Ghiasuddin, S. M., and SoderIund, D.M. (1985). Pyrethroid insecticidespotent, stereospecific enhancers of mouse-brain sodium-channel activation. Pestie. Bioehem. Pnysiol. 24, 200-206. Ginsburg, K. S., and Narahashi, T. (1993). Differential sensitivity of tetrodotoxin-sensitive and tetrodotoxin-resistant sodium-channels to the insecticide allethrin in rat dorsal-root ganglion neurons. Brain Res. 627, 239248. Glomot, R. (1982). Toxicity of deltamethrin to higher vertebrates. In "Deltamethrin" (R. Lhoste, ed.), Roussel Vclaf, Marseilles. Gomes, M. D., Bemardi, M. M., and Spinosa, H. D. (1991). Pyrethroid insecticides and pregnancy-Effect on physical and behavioral-development of rats. Vet. Hum. Toxieol. 33,315-317. Gray, A. J. (1985). Pyrethroid structure-toxicity relationships in mammals. Neurotoxicology 3, 25-35. Gray, A. J., and Rickard, J. (1982). Toxicity of pyrethroids to rats after direct injection into the central nervous system. Neurotoxicology 6, 127-138. Gray, A. J., Connors, T. A., Hoellinger, H., and Nam, N. H. (1980). The relationship between the pharmacokinetics of intravenous cismethrin and biotesmethrin and their mammalian toxicity. Pestic. Biochem. Physiol. 13, 281-293. Gupta, A., Agarwal, R., and Shukla, G. S. (1999). Functional impairment of blood-brain barrier following pesticide exposure during early development in rats. Hum. Exp. Toxieol. 18, 174-179. Hagiwara, N., Irisawa, H., and Kameyama, M. (1988). Contribution of 2 types of calcium currents to the pacemaker potentials of rabbit sino-atrial node cells. J. Physiol. 395, 233-253. Hassouna, I., Wickert, H., El Elaimy, I., Zimmermann, M., and Herdegen, T. (1996). Systemic application of pyrethroid insecticides evokes differential expression of c-Fos and c-Jun proteins in rat brain. Neurotoxieology 17, 415-431.
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Pyrethroid Insecticides
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Rose, G. P., and Dewar, A. J. (1983). Intoxication with four synthetic pyrethroids fails to show any correlation between neuromuscular dysfunction and neurobiochemical abnormalities in rats. Arch. Toxieol. 53, 297316. Rosengarten, H., and Friedhoff, A. J. (1979). Enduring changes in dopamine receptor cells of pups from drug administration to pregnant and nursing rats. Science 203, 1133-1135. Ruigt, G. S. E, and van den Bercken, J. (1986). Action of pyrethroids on a nerve-muscle preparation of the clawed frog, Xenopus laevis. Pestic. Biochem. Physiol. 25, 176-187. Salgado, V. L., Irving, S. N., and Miller, T. A. (1983). The importance of nerve tenninal depolarization in pyrethroid poisoning of insects. Pestic. Biochem. Physiol. 20, 169-182. Schinkel, A. H., Wagenaar, E., Mol, C. A. A. M., and Van Deemter, L. (1996). P-glycoprotein in the blood-brain barrier of mice influences the brain penetration and pharmacological activity of many drugs. 1. Clin. Invest. 97, 2517-2524. Sheets, L. P., Doherty, J. D., Law, M. w., Reiter, L. w., and Crofton, K. M. (1994). Age-dependent differences in the susceptibility of rats to deltamethrin. Toxicol. Appl. Pharmacol. 126, 186-190. Sherby, S. M., Eldefrawi, A. T., Deshpande, S. S., Albuquerque, E. X., and Eldefrawi, M. E. (1986). Effects of pyrethroids on nicotinic acetylcholine receptor binding and function. Pestic. Bioehem. Physiol. 26, 107-115. Smith, T. J., and SoderIund, D. M. (1998). Action of the pyrethroid insecticide cypermethrin on rat brain IIa sodium channels expressed in Xenopus oocytes. Neurotoxicology 19, 823-832. Soderlund, D. M. (1985). Pyrethroid-receptor interactions: Stereospecific binding and effects on sodium channels in mouse brain preparation. Neurotoxieology 6, 35-46. Soderlund, D. M., and Knipple, D. C. (1995). Actions of insecticides on sodium-channels-Multiple-target sites and site-specific resistance. Am. Chem. Soc. Symp. Ser. 591,97-108. Soderlund, D. M. et al. (2000). Differential sensitivity of sodium channel auxiliary subunit gene from the house fly (Musca domestica). Neurotoxicology 21, 127-137. Song, J. H., and Narahashi, T. (1995). Selective block oftetramethrin-modified sodium channels by (+j-)-alpha-tocopherol (vitamin E). 1. Pharmacol. Exp. Ther. 275, 1402-141 I. Song, J. H., and Narahashi, T. (1996a). Differential effects of the pyrethroid tetramethrin on tetrodotoxin-sensitive and tetrodotoxin-resistant single sodium channels. Brain Res. 712, 258-264. Song, J. H., and Narahashi, T. (1996b). Modulation of sodium channels of rat cerebellar Purkinje neurons by the pyrethroid tetramethrin. 1. Pharmacol. Exp. Ther. 277,445-453. Song, J. H., Nagata, K., Tatebayashi, H., and Narahashi, T. (1996). Interactions of tetramethrin, fenvalerate and DDT at the sodium channel in rat dorsal root ganglion neurons. Brain Res. 708, 29-37. Squiban, A., Marano, E, and Ronot, X. (1986). The action of deltamethrine, a pyrethrinoid. BioI. Cell 57, A5 I. Staatz, G., and Hosko, N. H. (1985). Effect of pyrethroid insecticides on EEG activity of conscious, immobilized rats. Pestic. Biochem. Physiol. 24,231239. Staatz, C. G., Bloom, A. S., and Lech, J. J. (1982). A pharmacological study of pyrethroid neurotoxicity in mice. Pestic. Biochem. Physiol. 17,287-292. Staatz-Benson, C. G., and Hosko, M. J. (1986). Interaction of pyrethroids with mammalian spinal neurons. Pestic. Biochem. Physiol. 25, 19-30. Talts, U., Fredriksson, A., and Eriksson, P. (1998). Changes in behavior and muscarinic receptor density after neonatal and adult exposure to bioallethrin. Neurobiol. Aging 19, 545-552. Tateno, C., Ito, S., Tanaka, M., and Yoshitake, A. (1993). Effects of pyrethroid insecticides on gap junctional intercellular communications in balb/c3t3 cells by dye-transfer assay. Cell BioI. Toxicol. 9,215-221. Thiebault, J. J., Bost, J., and Foulhoux, P. (1985). Experimental intoxication by deltamethrin in the dog and its treatment. Collect. Med. Leg. Toxicol. Med. 131, 47-62 (in French).
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CHAPTER
60 DDT and its Analogs Andrew G. Smith Medical Research Council Toxicology Unit, United Kingdom
60.1 INTRODUCTION Chlorinated insecticides have had a "bad press" over the last few decades. Their use has declined dramatically. Even the few uses worldwide and the few countries this applies to are under tremendous environmental pressure to cease completely. It is difficult for us now to accept that in their time the chlorinated insecticides were of outstanding importance for human health. As with penicillin, it was wartime that showed their tremendous utility. For the first time, pests could be confidently controlled and importantly irradicated. It is estimated that DDT alone was responsible for the saving of tens of millions of human lives. Many areas of the world are now free of such pests as malaria purely as the result of DDT use in the postwar period. That is not to say that our present concerns of environmental toxicity of these chemicals are not perfectly valid. Clearly, the large-scale administration of these persistent chemical had significant costs for wildlife and are undesirable. Whether there are any costs for chronic human health through environmental exposure is a very difficult issue. In the last 10 years or so DDT, its analogs, and other chemicals have been considered as insidious endocrine disruptors causing cancer and reproductive disorders but on the whole, DDT and its analogs have apparently been extremely safe for humans. It would seem prudent to retain their use, in a tightly controlled manner, as an important reserve weapon in the fight against insect pests. It is not always easy to substitute with newer insecticides that are more expensive and have their own undesirable effects. This chapter describes aspects of the known toxicity of DDT and its analogs such as methoxychlor. Although much of the work reported is now some years old (WHO, 1979), it is extremely important and very pertinent to today's discussions and arguments on the hazards of environmental exposures (Smith, 2000).
60.2 DDT DDT came to widespread attention because it dramatically controlled typhus and malaria in time of war. When it became available for civilian use, it controlled flies and other pests that annoy large numbers of people and may transmit disease, and Handbook of Pesticide Toxicology Volume 2. Agents
it increased the production of important crops. Knowledge that traces of it are stored in essentially everyone in the world has kept DDT in the spotlight. Later it was implicated in the injury of a wide variety of wildlife. Under these circumstances, it is no wonder that DDT probably has been studied more thoroughly than any other pesticide and is used to illustrate many principles and concepts in toxicology including very important topics such as human exposure levels and effects on domestic and wild animals. In the following discussion, details of storage and excretion of DDT in humans are covered as well as toxic actions in humans and experimental animals. DDT was first synthesized by Zeidler (1874) who called it dimonochlorophenyltricholorathan. However, it was put to no use until its insecticidal properties were demonstrated by Paul Muller in 1939. The first sample sent to the United States arrived in September 1942. Results of the tests were so encouraging that manufacture was given high priority both in the U.S. and UK. At first, the entire production was used for the protection of troops against malaria, typhus, or certain other vector-borne diseases, or against biting flies or other insects, that are merely pests (Hayes, 1982). As the supply increased, DDT was used in the United States for control of malaria in war areas, that is, in the vicinity of military installations, ports, and transportation centers. As a result of this effort, mosquito transmission of malaria was brought to an end in the United States in 1953, even though military personnel and other infected persons from the tropics continued to reintroduce the disease extensively as late as 1972 and in diminishing numbers thereafter. The revolution in the control of malaria and typhus among allied troops and among certain civilian populations during World War II was accomplished with relatively little DDT. Far greater amounts were required for the control of agricultural and forest pests that became possible after the compound was released in the United States for commercial use on August 31, 1945. Civilian use in other countries became possible a little later with tremendous effect. An informative account of its discovery and early use, especially in Europe, can be found in West and Campbell (1946).
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Copyright © 2001 by Academic Press. All rights of reproduction in any fonn reserved.
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DDT and its Analogs
60.2.1 IDENTITY, PROPERTIES, AND USES
60.2.2 FORMULATIONS AND PRODUCTION
Technical DDT has been formulated in almost every conceivp,p' -DDT is 1, 1'-(2,2,2-trichloroethylidene)-bis( 4-chloroable form including solutions in xylene or petroleum distillates, benzene). Common nomenclatures that have been used are emulsifiable concentrates, water-wettable powders, granules, 1,1, I-trichloro-2,2-bis (p-chlorophenyl)ethane, 1,1, I-trichloro- aerosols, smoke candles, charges for vaporizers, and lotions. 2,2-bis( 4-chlorophenyl)ethane, and 1, I-bis( 4-chlorophenyl)- Aerosols and other household formulations were combined 2,2,2-trichloroethane. Because the older terminology has been with synergized pyrethrins. Production and use of DDT in the used widely in the past and continues to be so, especially since United States and other countries have been discussed previmany abbreviations are based on it (e.g., p,p'-DDT and o,p'- ously (Hayes, 1991; Smith, 1991). DDT), the and p nomenclature will be used for referring to Before 1945, all of the DDT produced in the United States DDT in its abbreviated form. was used or allocated by the military services for medical and DDT is universally accepted as the common name of the public health uses. Early in 1945, it became available for exteninsecticide. As approved by BSI, DDT refers to the technical sive experimental work in agriculture, and it was commercially product, and there is historical justification for that practice be- available in limited quantities early in the autumn of the same cause DDT is an acronym for dichlorodiphenyltrichloroethane. year. The results were so spectacular that use increased until p,p'-DDT was approved by BSI as a separate term. Zeidler 1959. In response to a demand for exports, production con(1874) called the compound di-monochlorophenyltrichlorlithan. tinued to increase until about 1963. Even before 1963, some When used as a drug, DDT was known in the United Kingdom restrictions were placed on its use, mainly to minimize residues as dicophane (BP), in Sweden as klorfenoton, and in the United in food and in the feed of animals that produced milk and meat. Among the first of these restrictions was that on its use in the States as chlorphenothane (USP). dairy cattle industry. Another important factor reducing the use DDT was sold under a variety of trade names, includof DDT was the increasing resistance of pests. One of the first ing Anofex®, Cesarex®, Didimac®, Digmar®, Diniocide®, species to be affected was the housefly; because of its abunGenitox®, Guesarol®, Gyron®, Ixodex®, Neocid®, and dance and widespread distribution, its resistance was bound to Zerdane®. Code designations for DDT include IMS-16 and be noticed by the public generally. Although many pests of pubENT-l ,506. The CAS registry number for p,p'-DDT is 50- lic importance have been resistant to DDT in some or all of 29-3. DDT has the empirical formula C4H9Cls and a molec- their range, resistance among vectors of malaria has been minular weight of 354.49. Pure p,p'-DDT is a white, tasteless, imal. Because malaria control constitutes such a large segment almost odorless crystal-line solid melting at 108.5 to 109.0°C. of vector control, the use of DDT for vector control remained Technical DDT is a waxy solid. A typical example of technical stable, while its use in agriculture continued to decline, espeDDT had the following composition: p,p'-DDT, 77.1%; o,p'- cially in temperate climates. DDT, 14.9%; p,p'-DDD, 0.3%; o,p'-DDT, 14.9%; p,p'-DDD, Prophetically when Sweden banned DDT from January 1, 0.3%; o,p'-DDD, 0.1%; p,p'-DDE, 4.0%; o,p'-DDE, 0.1%; 1970, they pointed out that "the need for insecticides is rather and unidentified compounds, 3.5%. The vapor pressure ofDDT small in Sweden compared to that in many other countries" and is 1.5 x 10- 7 mm Hg at 20 o e. DDT is highly soluble in a polar that the ban of this and certain other chlorinated hydrocarbon organic solvents: solubility per 100 ml acetone, 58 gm; ethanol, insecticides could be used as a tool to explore scientific prob2 gm; benzene, 106 gm; carbon tetrachloride, 45 gm; cyclo- lems about their movement (Hayes, 1969). In order to respond hexanone, 116 gm; ethyl ether, 28 gm; petroleum ethers and to ecologists who considered that the widespread occurrence of DDT in the environment was inherently bad and was the direct kerosene, 4-10 gm. It is practically insoluble in water. The structure of p,p'-DDT and the structures of several cause of injury to certain fish and birds, government agencies of its analogs are compared in Table 60.1. The table is con- of some other countries attempted to justify restrictions on the use of DDT by its alleged threat to human health. This did not fined to compounds that have occurred in commercial DDT prevent the same agencies from providing that DDT might be and analogs that have had some use as insecticides. It must be used, if needed, to combat any future threat from vector-borne emphasized that even the commercially available insecticidal disease within their boundaries. To this day the hazard of DDT analogs have strikingly different properties. Especially remark- and DDE to humans is still a highly debatable matter. Although able are the slow metabolism and marked storage of DDT and many countries severely restrict or ban the use of DDT, it is its metabolite DDE and the rapid metabolisms and negligible still used for both agriculture and vector control, in some tropistorage of methoxychlor. Table 60.1 does not include the wide cal countries. It is possible that complete abolishment of its use range of compounds that have been synthesized and studied in worldwide in vector control might have significant repercusconnection with structure-activity relationships, often with the sions. Information apparently is not available on how much of hope of emphasizing the good properties of DDT and reducing the agricultural use involves food protection or how much loss its undesirable properties. For a more extensive consideration of food production would result if use of DDT were discontinof analogs, see Metcalf (1973). Further information of related ued. How much of the use ofDDT is in public health is still also unknown, but the picture with malaria control is clear. In 1971, chemicals will be found in Section 60.2.3.5.
°
60.2 DDT
1307
Table 60.1 Structure of p,p' -DDT and a Few of Its Analogs that Have Had Commercial Use
oc
RI
RI
Chemical namea
RI
R2
R3
DDT
1,1, I-trichloro-2,2-bis(4-chloropheny l)ethane
Cl
H
CCI3
Bulan®b
2-nitro-l, I-bis(4-chloropheny I)butane
Cl
H
CH(N02)CH2CH3
Chlorfenethol (DMC)
I, l-bis(4-chlorophenyl)ethanol
Cl
OH
CH3
Chlorobenzilate
ethyl 4,4' -dichlorobenzilate
Cl
OH
COOCH2CH3
Chloropropylate
isopropyl 4,4' -dichlorobenzilate
Cl
OH
COOCH(CH3h
Name
DFDT
1,1, I-trichloro-2,2-bis(4-fluropheny l)ethanol
F
H
CCI3
Dicofol (Kelthane®)
2,2,2-trichloro- I, I-bis(4-chlorophenyl)ethanol
Cl
OH
CCI3
Ethylan (Perthane®)
I, I-dichloro-2,2-bis( 4-ethylphenyl)ethane
CH2CH3
H
CHCI2
Methoxychlor Prolan®b
I, I, l-trichloro-2,2-bis(4-methoxyphenyl)ethane
OCH3
H
CCI3
2-nitro-l, l-bis(4-chloropheny l)propane
Cl
H
CH(N02)CH3
TDEc
1, I-dichloro-2,2-bis(4-chlorophenyl)ethane
Cl
H
CHCI2
aChemical names: The names used here are those which are commonly encountered. b A mixture of Prolan and Bulan (I : 2) has been sold in the past as Dilan®. C As an insecticide, this compound has had the approved name of TDE; as a metabolite of DDT it usually is called DDD. It has been sold under thename Rhothane®; as a drug, the a,p'-isomer is called mitotane.
WHO calculated that substitution of malathion or propoxur for DDT would increase the cost of malaria control approximately 3.4- and 8.5 fold, respectively, and this increase could not be supported in some countries without a decrease in the coverage of control programs (WHO, 1971). Despite these increased costs, DDT use has been mainly overtaken by other pesticides. However, a proposal for its complete ban is still a controversial matter (Curtis and Lines, 1999; Attaran et aI., 2000; Roberts et aI., 2000). 60.2.3 TOXICITY TO LABORATORY ANIMALS 60.2.3.1 Symptomatology
The description of DDT intoxication in animals given by Domenjoz (1944) remains one of the best. The first perceptible effect is abnormal susceptibility to fear, with violent reaction to normally subthreshold stimuli. There is definite motor unrest and increased frequency of spontaneous movements. As poisoning increases, hyperirritability like that seen in strychnine poisoning develops, but convulsions do not appear at this time. A fine tremor, recognizable at first only as a terror reaction, is later present as an intention tremor in connection with voluntary movement, and then intermittently without observable cause. Finally it is present as a coarse tremor without interruption even for several days. Spontaneous movement is limited, and food intake stops so that surviving animals lose weight. In the later stages, especially in some species, there are attacks of epileptiform, tonic-clonic convulsions with opisthotonos. All the signs are strengthened by external stimuli and become man-
ifest at first through external stimuli. In all stages, the animals show normal position and labyrinth reflexes. The picture of poisoning in mammals recalls the disturbances of movement and tone that are known in human pathology as the amyostatic syndrome. Symptoms appear several hours after oral administration of the compound, and death may follow after 24-72 hr. The latent period after intravenous administration at about the LD 50 level is approximately 5 min; signs of poisoning reach a maximal level in about 30 min, and survivors are symptom-free in 18-24 hr. Animals that survive recover completely. In addition to the features of poisoning already mentioned, Cameron and Burgess (1945) noticed that as rats, guinea pigs, and rabbits become sick they become cold to the touch and show ruffled fur. Some show diarrhea. Muscular tremors were preceded by muscular weakness which occurred first in the back and later in the hind legs. The front legs were relatively spared so that animals showing marked weakness of the hindquarters could still drag themselves about. However, several authors have found that the tremor characteristic of DDT poisoning generally starts in the muscles of the face, including the eyelids, and spreads caudally with variable severity until all the muscles are affected. Furthermore, although weakness of hindquarters has been seen, it was not a common finding. Although there is a general similarity in the clinical effects ofDDT in all vertebrate species, some characteristic differences exist. Cats show greater extensor rigidity and opisthotonos than other laboratory animals. The stiffness appears first in the distal part of the extremities and later extends to the proximal part and to the trunk. Poisoned cats show marked pilomotor activity. Convulsions are also prominent in dogs, as is ataxia. Tremors
1308
CHAPTER 60 DDT and its Analogs
Table 60.2 Comparison of Acute LD 50 to Laboratory Animalsa
Species Rat Mouse Guinea pig
Fonnulation b
I.p.
Oral
Dermal
(mg/kg)
(mg/kg)
(mg/kg)
(mg/kg)
47
80-200
w/p
<2000 200-1500
250-3000
1000-1500
300-1600
375
300
100-800
250-500
2000
1500
w/p 900
150
w/p 250-3200
30-41
<2100
250-560
1000
275
375
300-1770
300-2820
w/p 68
w/p
>300 100-410
>650
0
Monkey
1000
113-450
w/p
0
Cat
500-2500
0
0
Dog
I.v.
(mg/kg)
0
0
Rabbit
S.c.
32
w/p 55
0
aModified from Hayes (1959). bw/p suspension in water or as powder.
0,
solution in oil.
are so pronounced in rats that it may be difficult to detect clonic convulsions in them. Poisoning produced by repeated doses of DDT differs from that produced by a single dose only insofar as the animal may be gradually debilitated, especially by malnutrition. If food intake is maintained, tremor may last for weeks or even, intermittently, for months. If the animals survive a short time after dosing stops, recovery is complete. However, food intake may be interfered with in at least two ways. Tremor and more severe signs may interfere mechanically with eating. Animals offered food containing high concentrations of DDT often eat little or nothing and lose weight rapidly. This seems to be due to taste, not an effect on appetite, as the same animals will show excellent appetites when offered the same kind of food containing no DDT just after refusing the major portion of their daily ration of contaminated food. Animals that have suffered severe weight loss as a result of DDT poisoning may die partly as a result of general debility. Even though severely ill, animals that survive a few days after the last of many doses of DDT go on to recovery. Table 60.2 summarizes the acute toxicity ofDDT to common laboratory animals. It may be concluded that dissolved DDT is absorbed by all routes, although DDT powder is absorbed through the skin to only a negligible degree. Remarkably it is frequently impossible to put enough DDT dust on the skin of animals to kill them, so that a LD 50 value for this formulation cannot be determined with precision by the dermal route. Although formulation is important in determining the toxicity of DDT by other routes, the difference is not so great as it is in connection with skin exposure. In round figures, DDT is about 4 times more toxic when given intravenously than when given orally and about 40 times more toxic intravenously than dermally. In general, DDT, like some other lipophilic chemicals, appears more toxic orally as a solution in vegetable oil or ani-
mal fat than when given in some petroleum fractions. Acute oral LD 50 values of DDT metabolites commonly found in tissues or excreta are less toxic than the most absorbable preparations of the parent compounds. At an oral dosage of 150 mg/kg, p,p'-DDT produces severe illness in rats and kills about half of them, but O,p' DDT at the same dosage produces no illness, even thought the concentrations of the two compounds in the brain at various intervals after dosing are about the same. At a dosage of 3000 mg/kg, a,p' DDT produces mild to moderate illness, and the concentration in the brain is 5-9 times the concentration of p,p'-DDT necessary to produce similar symptoms. Thus, p,p'-DDT appears to be inherently more toxic than the a,p' isomer (Dale et aI., 1966). Rats tolerate higher tissue levels of DDA than of DDT. Eighteen hours after intravenous injection of DDA at a rate of 100 mg/kg, tissue levels still were higher than are usually found in animals fatally poisoned by DDT (Judah, 1949). DDA produces somewhat less injury than DDT to the liver but, especially at high intravenous dosages, produces greater damage to the kidney (Lillie et aI., 1947). This is consistent with the finding of Spicer et al. (1947) that, following administration of DDT, DDA constitutes a higher proportion of DDT-related compounds in the kidney (25%) than in any other tissue, being 12 % in the liver, 10% in the brain, and even less in other tissues. Young animals eat more than adults in relationship to their body weight. For this and other reasons, young animals often are more susceptible than adults to poison in food. However, there is no evidence that DDT is more toxic to young animals of any species, including humans, and in the rat the young are less susceptible to a single dose (Table 60.3). They are about equally susceptible to repeated doses. According to Henderson and Woolley (1969), the relative insusceptibility of the young
60.2 DDT
1309
Table 60.3 Effect of Age on the Toxicity of DDT to Rats LD50 Number of doses
Agea
(mglkg)b
Reference
Newborn
>4000
Lu et al. (1965)
Newborn
2356
Harrison (1975)
10 days
728
Henderson and WooIley (1969)
14-16 days
437
Lu et al. (1965) Lu et al. (1965)
weaning
355
2 months
250
Henderson and WooIley (1969)
2 months
152
Mitjavila et al. (l981a)
3--4 months
194
Lu et al. (1965)
middle-aged
235
Lu et al. (1965)
adult
225
Harrison (1975)
4
reweaning
279
Lu et al. (1965)
4
adult
285
Lu et al. (1965)
a Data
from more than one strain of rat.
bTotal intake of one or more doses.
is associated with relatively poor absorption of DDT by their central nervous systems and by lesser inherent susceptibility of the young brain to DDT already absorbed by it. Further study by the same authors (Henderson and Woolley, 1970) showed that fatal poisoning of both 10- and 6O-day-old rats involves hyperexcitability and intense tremor followed by prostration and eventual respiratory failure. However, in the adult rat, DDT causes convulsions, an increase in respiration and heart rate, and a lethal increase in body temperature prior to death, but the body temperature of the immature rat decreases during acute intoxication by DDT. The authors suggested that, whereas DDT is a direct depressant of respiration in both young and old rats, the additional toxic responses manifested by seizures and hyperthermia account for the increased lethality of DDT in mature animals. No acute LD 50 could be established for hamsters (Agthe et aI., 1970), which also seem resistant to chronic effects ofDDT (Table 60.4). There is virtually no sex difference in the acute toxicity of DDT to rats; the LD 50 is 113 and 118 in males and females, respectively (Gaines, 1960). A similar situation is observed with mice (Agthe et aI., 1970). When DDT is fed to rats at ordinary dietary levels, the two sexes store it equally. However, at higher dosages, females store more of the compound; the difference is explained mainly by the lesser activity of the liver microsomal enzymes in female rats and only in part by relatively higher food intake of the females. 60.2.3.2 Response to Repeated Doses The effects of repeated doses of DDT are summarized in Table 60.4. The 90-dose oral LD 50 of technical DDT in rats is 46.0 mg/kg/day (Gaines, 1969). The chronicity index is 5.4. Thus the compound has only a moderate tendency to cause cumulative effects, and this limited tendency is fully explained by the accumulation of DDT itself in tissues as a result of contin-
uing intake. In fact, this accumulation, which is strictly dosage dependent, is detectable at all measurable levels of intake. If storage is considered undesirable per se, then DDT is without a no-injurious-effect level. However, the same may be said for all compounds that are absorbed, for the presence of all of them in the bodies of exposed organisms-perhaps at very low levels and for relatively short periods-may be assumed; failure to demonstrate low levels of storage does not depend on physiology but only on limitations of the analytical techniques employed. A number of papers have reported no-effect levels for DDT within parameters other than storage, namely rat, 0.05 mg/kg/day; dog, 8mg/kg/day; and monkey, 2.2-5.54 mg/kg/day (see Smith, 1991). There remain reports of effects in animals at the lowest dosages investigated. For example, decreased serum albumin and increased fJ- and a-globulins in the blood of rats and rabbits maintained on a dosage of 0.2 mg/kg/day for 3-11 months were reported by Kagan et al. (1969). In summary, the lowest dosages that have been studied in animals are of the same order of magnitude as those encountered by people who made or formulated DDT and, therefore, hundreds of times greater than the dosages encountered by ordinary people. The animal studies have continued long after a steady state of storage has been achieved. The results permit the conclusion that bioaccumulation sufficient to produce neurotoxicity or other clinical effects, including a reduction of the life span, can occur only at dosage levels substantially higher than those encountered by the most heavily exposed workers, let alone those exposed environmentally. DDT dosages encountered by workers produce in some groups of mice and rats a small but detectable increase of the liver changes (hypertrophy, margination, and lipospheres) characteristic of rodents at much higher doses (Smith, 1991). The same changes occur in low incidence in control mice and rats but not in other animals.
1310
CHAPTER 60
DDT and its Analogs
Table 60.4 Effect on Various Species of Prolonged Oral Administration of DDT Dosage Range
Method and
Species,a
Maximum
(mg/kg/day)
concentration (ppm)
number, and sex
duration
Results
References
41-80
800 ppm in diet
rat
2 yr
increased mortality, typical liver
Fitzhugh and Nelson (l947)b
changes, and liver carcinomas 46 mg/kg, then 140 ppm in diet
mouse
l.5 yr
36 M, 36F
hepatomas in 51 and 21 % of M and
Innes et al. (1969)
F compared with 18 and 0.6% of controls
1000 ppm in diet
hamster
l.9 yr
25M,30F 1000 ppm in diet
hamster hamster
l.5 yr
dog
2.4 yr
monkey
Graillot et al. (1975)
no liver tumors and survival as
Rossi et al. (1983)
controls 4yr
10 5000 ppm in diet
no liver tumors but decreased serum cholinesterase
35 M, 36 F 3200 ppm in diet
Agthe et al. (1970)
less than controls
30M,30F 1000 ppm in diet
no liver tumors and survival slightly
100% mortality; liver damage, no
Lehman (1951, 1952, 1965)
tumors lOwk
100% mortality
Durham et al. (1963)
14wk
100% mortality; no hematologic
Cranmer et al. (1972)
IM 50 mg/kg/day
monkey 6
21-40
400 ppm in diet
rat
effects 2 yr
24 M, 12F 500 ppm in diet
rat
increased mortality, typical liver
Fitzhugh and Nelson (1947)b
changes 2.9yr
liver tumors in 45%
Rossi et al. (1977)
2.3 yr
liver tumors in 18% F
Cabral et al. (1982b)
2 gen
risk ofliver tumor increased 3.7- and 18.5-
Tomatis et al. (1972)
37 M, 35 F 500 ppm in diet
rat 38 M, 38 F
250 ppm in diet
mouse 103 M, 90F
250 ppm in diet
mouse
fold in M and F, respectively 2gen
31M,121F 500 ppm in diet
hamster dog
1.7 yr
100 ppm in diet
mouse
no liver tumors and survival as
Cabral et al. (l982a)
controls 4yr
25% mortality; minor liver
Lehman (1951, 1952, 1965)
damage but no tumors
4 11-20
Terracini et al. (1973)
of M and F
39M,40F 2000 ppm in diet
liver tumors in 48 and 59%
2 yr
100 M, 100F
hepatomas increased in F of one
Fitzhugh (1970)
strain but no increase in hepatocarcinomas
100 ppm in diet
mouse
2 yr
30M,30F 100 ppm in diet
mouse
50 ppm in diet
mouse
Walker et al. (1973)
4.4-fold 2 yr
30M,3F 6-10
risk of liver tumors increased risk ofliver tumors increased 3.3-
Thorpe and Walker (1973)
and 4.2-fold in M and F 2 gen
127 M, 104 F
risk ofliver tumors increased 2.45-
Tomatis et al. (1972)
and 3.46-fold in M and F, respectively
50 ppm in diet
mouse
2 yr
risk of liver tumors increased 2.9-fold
Walkeretal. (1973)
4 yr
no effect
Lehman (1951, 1952, 1965)
2gen
no increase in tumors
Terracini et al. (1973)
30M,30F 400 ppm in diet
dog 2
2.6-5
20 ppm in diet
mouse 48 M, 128 F
(continues)
60.2 DDT
1311
Table 60.4 (continued) Dosage Range
Method and
Species,a
Maximum
(mg/kg/day)
concentration (ppm)
number, and sex
duration
200 ppm in diet
monkey
7.5 yr
no effects
Durham et al. (1963)
10 ppm in diet
mouse
2 gen
risk of liver tumors increased 2.26-
Tomatis et al. (1972)
1.26-2.5
Results
References
and 2.46-fold in M and F,
104 M, 124F
respectively 0.63-1.26
25 ppm in diet
rat
2 yr
no clinical effect; M survived longer
Treon and Cleveland (1955)
than controls 0.31-D.63
10 ppm in diet
rat
2 yr
typical liver changes; no effect on
Fitzhugh (1948)
reproduction 12.5 ppm in diet
rat
2 yr
no effect
Treon and Cleveland (1955)
2.8-3.0 ppm in diet
mouse
5 gen
tumors in 28.7%, including lung
Tarjan and Kemeny (1969)
carcinomas, lymphomas, and
683
leukemias 0.16-D.31
2 ppm in diet
mouse
2 gen
124 M, 111 F 2 ppm in diet
mouse
risk of liver tumor doubled in M,
Tomatis et ai. (1972)
unchanged in F 2gen
no increase in tumors
Terracini et al. (1973)
2 yr
no effect
Treon and Cleveland (1955)
58 M, 135 F 0.08-D.16
2.5 ppm in diet
rat
aVarious strains of rats were used: Osbome-Mendel (Fitzhugh and Nelson, 1947), Carworth (Treon and Cleveland, 1955), Wistar (Rossi et aI., 1977), MRCPorton (Cabral et ai., 1982b). Mouse strains used were (C57BLl6 x C3H1An)Fl and C57BLl6 x AKR)Fl (Innes et ai., 1969), CFI (Tomatis et aI., 1972; Thorpe and Walker, 1973; Walker et aI., 1973), BALB/cl andC3HeBIFeJ (Fitzhugh, 1970), BALB/c (Tarjan and Kemeny, 1969; Terracini et ai., 1973). bSlides reexamined by Reuber (1978).
60.2.3.3 Absorption Most DDT dust is of such large particle size that any that is inhaled is deposited in the upper respiratory tract and eventually is swallowed. Toxicity data indicate that respiratory exposure to DDT is of no special importance. The absorption of DDT from the gastrointestinal tract is slow. Whereas intravenous injection at the rate of 50 mg/kg produces convulsions in rats in 20 min, convulsions occur only after 2 hr when DDT is administered orally at a rate two or more times the LD 50 value. The onset of convulsions is delayed for about 6 hr when DDT is given to rats orally at approximately the LD 50 value (Dale et aI., 1963). DDT dissolved in animal or vegetable fats is absorbed from the gastrointestinal tract about 1.5-10 times more effectively than is undissolved DDT (e.g., Keller and Yeary, 1980; Palin et aI., 1982), but large doses of the compound in the gastrointestinal tract are poorly absorbed from nonabsorbable solvents. At high dosage levels, less 4 C]DDT is absorbed and stored in organs following oral than following intraperitoneal administration, and a higher proportion is excreted in the feces than after intraperitoneal administration (40 versus 0.9%) (Bishara et aI., 1972). However, in connection with small repeated doses, the kind of solvent used made little difference; apparently the occurrence of bile in the intestine and the presence of some fat in the diet are sufficient to promote absorption of the compound. Rothe et al. (1957) reported that after giving radioactive DDT by stomach tube as an emulsion of a peanut oil solution they recovered 41-57% of it in lymph. Less than 0.1%
e
was found in the urine, 7.4-37.1% was in the feces or in the intestinal contents, and 19-67% of the activity was found in the carcass. Of the administered DDT not found in feces and intestinal contents, 47-65% was found in the lymph. Fifty percent of the DDT-derived material found in the lymph was absorbed in the first 2.5-7 hr, and 95% was absorbed by 18 hr. Because the lymphatic duct in the rat is not a single vessel, Rothe et al. (1957) were unable to exclude the possibility that some or all of the DDT that they later recovered from the carcasses of their animals had been transported to the general circulation by collateral lymph vessels rather than by the hepatoportal system. They gave indirect evidence for supposing that little or no DDT is absorbed from the gastrointestina1 system by the blood, and this has been confirmed by Palin et al. (1982). Most of the DDT absorbed into the lymph is carried in the lipid core of chylomicrons and thence into the plasma proteins (Pocock and Vost, 1974; Sieber et aI., 1974). p,p'-DDT is taken up at a rate which is different from those of its metabolites and o,p'-DDT (Sieber, 1976) and which does not strictly parallel differences in lipid solubility. As already stated, dermal absorption ofDDT is very limited. 60.2.3.4 Distribution and Storage The distribution and storage of DDT in animals can be summarized as below. Original references can be found in Smith (1991).
1312
CHAPTER 60
DDT and its Analogs
1. DDT is stored in all tissues the highest concentrations of DDT usually being found in adipose tissue. Rats store DDT in their fat at all measurable dietary levels including trace concentrations. 2. Following repeated doses, storage in the fat increases rapidly at first and then more gradually until a peak or plateau is reached. Repeated doses at a moderate rate could result in greater total storage of DDT in the fat than a single dose at the highest rate that can be tolerated or even a single dose at a rate that frequently is fatal. The equilibrium storage of DDT in each tissue varies directly with the daily dosage. However (with the apparent exception of the dog), storage in the fat and perhaps in other tissues is less extensive in relation to dosage at higher dietary levels. 3. The rat apparently tends to lose a part of the DDT it has stored in fat at the peak level reached in about 6 months, even though it continued on the same diet. 4. There is a measurable difference between the storage patterns of different species; that of the dog differs most. 5. At higher dosage levels but not at ordinary residue levels, the female rat consistently stores more DDT in its fat than the male when fed the same diet. The difference is accounted for only in part by the greater food intake of the female and must depend partly on more rapid biotransformation in the male. Other species show little or no sex difference. 6. The amount of DDT stored in the tissues is gradually reduced if exposure to the compound is discontinued or diminished. Other observations regarding storage include the finding that rats whose brains contain DDT at a concentration of 25 ppm or less (wet weight) usually survive, whereas higher levels tend to be fatal regardless of whether absorption followed one or many doses. Of samples that may be collected in vivo, the concentration of DDT in serum most accurately reflects its concentration in the brain, the critical tissue. Adams et al. (1974) observed that about the same concentrations of DDT and related compounds are stored by male rats and by females that reproduce successfully. The lower storage in mated females probably is accounted for by transfer to the young via the placenta and the milk. However, other factors may be involved. When DDT, some of its analogs, and several other chlorinated hydrocarbon insecticides were fed to male and female rats for four generations, there was little variation in storage of the materials from one generation to another and no evidence of a continuing increase in succeeding generations (Adams et aI., 1974). The concentrations of DDT in the blood and other tissues of the fetus are lower than those in corresponding tissues of the mother (Dedek and Schmidt, 1972). DDE constituted about 4% of technical DDT. Most species convert some of the DDT they ingest to DDE. Finally, most species, including humans, store DDE more tenaciously than they do DDT, the greater part of which is metabolized by a different pathway from that of DDA and excreted more rapidly.
The result is that DDE, expressed as a percentage of total DDTrelated compounds, increases in individuals after DDT intake decreases and increases in successive steps of the food chain. The Rhesus monkey apparently is an exception. Monkeys store DDE when it is fed to them. However, when feeding is stopped, the rate of loss of DDE stored in fat is more rapid than that of DDT (Durham et aI., 1963). Whether it is relative inability to form DDE, unusual ability to excrete it, or a combination of both that accounts for the fact that little or no DDE can be found in monkeys fed DDT is not entirely clear. 60.2.3.5 Metabolism The chemical nature of the chief metabolite excreted in the urine was first elucidated by White and Sweeney (1945). 2,2bis(4-chlorophenyl) acetic acid (DDA) was isolated from the urine of rabbits chronically administered DDT. Later work by many authors confirmed that DDA isomers are the major urinary metabolites of p,p'-DDT and o,p'-DDT in all mammals, including humans, but the nature of all excreted metabolites still may not have been elucidated fully. The ability of phenobarbital and especially diphenylhydantoin to promote the excretion of DDT was discovered in humans (Davies et aI., 1969) and later confirmed in animals (Alary et aI., 1971; Cranmer, 1970; Fries et aI., 1971). This is of course consistent with our current knowledge of the induction of drug metabolizing enzymes. That portion of the metabolism of DDT that leads to DDA in rats was explored by Peters on and Robinson (1964), who gave evidence for the sequence of changes leading to DDA involving reduction to 1,1-dichloro-2, 2-bis(4-chlorophenyl)ethane (DDD) followed by dehydrochlorination to 1-chloro-2,2-bis(4chlorophenyl)ethane (DDMU), which was apparently converted to 2,2-bis-(4-chlorophenyl)ethanol (DDOH) via 2,2-bis(4chlorophenyl)ethane (DDNU); see Fig. 60.1. The compound identified as a "probable" intermediate aldehyde between p,p'DDA was later synthesized and shown to be highly labile (McKinney et aI., 1969; Peterson and Robinson, 1964), confirming that it is unlikely to accumulate in tissues in measurable amounts. Kujawa et al. (1985) obtained evidence for its formation from p,p' DDD by rat liver homogenates and its presence in the urine of rats injected with DDD. Two additional metabolites, bis(p-chloro-phenyl)methane (DDM) and bis(pchlorophenyl) methyl ketone (DBP), were identified in chicks (Abou-Donia and Menzel, 1968). Not only was DBP found to result from the metabolism of DDA with DDM as an intermediate, but DBP was the only metabolite of DDE administered. Organ perfusion studies indicated that the liver is capable of biotransformation of DDT, DDE, DD, DDMU, and other possible metabolites (Datta and Nelson, 1970). Cultures of human embryonic lung cells are capable of metabolizing DDT to DDA via DDD (North and Menzer, 1973). On the other hand perfusi on of rat, guinea pig, pig and human skin samples has shown poor percutaneous passage of DDT «1 %) and no evidence for cutaneous metabolism (Moir et aI., 1994). When DDA was discovered, it was postulated that DDE was a step in its formation (White and Sweeney, 1945); however,
60.2 DDT
1313
CCI3
~ clN ~CI p.p'.DDT*
CCI,
~...... -
HOV
VCI
"',
~
CIV
,,
~ ~CI
CHC,
Cl
~ I ~I 0..
~
Cl
CIN
p.p'.DDD*
,
/
CI~ HO V
--------~.~
VCI
p.p'·DDE*
4.Hydroxy· p'·DDE
CC,
\
/
p.p'·DDCHO
/
/
"
/
......
COCI
~
/
~ clN ~CI
VCI
4·Hydroxy·m. p'.DDE
Acyl chloride intenncdiate
dY;
CIY OH
~CI
3·Hydroxy·p. p'·DDE
CH20H
+
CI~OHVCI 2·Hydroxy·p. p'·DDE
~ clN ~CI
COOH
0..
Cl
~ I~
du i
p.p'.DDOH
Cl
p.p'·DDA*
Figure 60.1 Metabolites of p,p' ·DDT and the postulated route of metabolism in the rat. The metabolites indi· cated by an asterisk have been found in humans.
rats which produced both DDE and DDA from DDT were said by Peterson and Robinson (1964) to be incapable of forming DDA when fed preformed DDE. This finding was contradicted by Datta (1970) and by Datta and Nelson (1970), who claimed that 14C-labeled p,p' DDE was converted by rats to 1-chloro2,2-bis(4-chlorophenyl)ethene (p,p' -DDMU), which then underwent further metabolism to p,p' -DDA. Datta suggested that the predominance of detoxication via DDE or DDD may depend on physiological response or the amount of toxic ant used. DDE is stored in tissue as first demonstrated in connection with human fat (Mattson et aI., 1953; Pearce et al., 1952). In fact DDE is stored more tenaciously than DDT. The way in which DDE is lost from storage remained something of a mystery. In humans (Cueto and Biros, 1967), seals, and guillemots (Jansson et aI., 1975) part of it is excreted unchanged, but the fact that its elimination is promoted by inducers of drug metabolism enzymes strongly suggests that much undergoes metabolism, conjugation, or both. That metabolism does occur was first demonstrated by identification of two hydroxylated derivatives of DDE in the feces of wild seals and guillemots and in the bile of seals (Jansson et aI., 1975). When p,p' -DDE was fed to rats, the same metabolites and one other were isolated from the feces, accounting for about 5% of the dose (Sundstrom et aI., 1975). Later, a fourth hydroxylated derivative was identified in the feces of rats fed p,p' -DDE. The metabolites are m-hydroxy-p,p'-DDE [1,1O-dichloro-2-
(p-chloro-m-hydroxyphenyl)-2,2(p-chlorophenyl)ethylene, the major metabolite], o-hydroxy-p,p'-DDE, p-hydroxy-m,p'DDE (the product of an NIB shift), and p-hydroxy-p'-DDE. A scheme involving m,p-epoxy- p,p' -DDE and o,m' -epoxyp,p' -DDE was proposed for the formation of these metabolites as well as a fifth metabolite (Sundstrom, 1977). In mice, feeding DDE increased the hepatic levels of radioactivity from [14C]DDE and decreased that in the urine and feces (Gold and Brunk, 1986). The only metabolite identified was the a-hydroxylated product. DDE is metabolized not only to easily excretable phenols but also to m-methylsulfone-p,p'-DDE. In the blubber of seals from the Baltic, this compound was found in a concentration of 4 ppm along with DDE (138 ppm), DDD (10 ppm), DDT (78 ppm), and various polychlorinated biphenyls (PCBs) and their metabolites (150 ppm) (Jenson and Jansson, 1976). Sulfurcontaining metabolites of halogenated aliphatic and aromatic chemicals usually arise by initial conjugation with glutathione. The possibility of glutathione-derived conjugates of DDT requires further attention. In vitro the reductive dechlorination of p,p' -DDT to DDD can occur with a cytochrome P-450 system, especially under anaerobic conditions (Esaac and Matsumura, 1980; Hassall, 1971; Zaidi, 1987). A one-electron reduction of DDT to the 1, 1-dichloro-2,2-bis(p-chlorophenyl)ethy1radical seems to occur, followed by abstraction of a hydrogen atom, possibly from
1314
CHAPTER 60
DDT and its Analogs
lipid, to give DDD (Kelner et aI., 1986). The reduction of DDT to DDD is stimulated by thiols in an unknown manner. The formation of an intermediate radical explains binding to microsomal lipid, especially under anaerobic conditions (Baker and Van Dyke, 1984). DDD, on the other hand, needs aerobic conditions for binding, implying that further metabolism is required. Other studies with mouse liver microsomes have shown the formation of 2,2-bis(p-chlorophenyl)-1 ,2-ethanediol (DDNU-diol) from DDNU, suggesting that a reactive epoxide intermediate might be formed (Planche et aI., 1979). When synthesized, however, the ethylene oxide (DDNU-oxide) was not mutagenic. Gold and colleagues examined the metabolism of DDT metabolites in mice in vivo. The results seem to be a little different from that previously accepted for rats. It is thought that DDMU can undergo epoxidation; the resulting mutagenic epoxide is hydrolyzed and oxidized to 2-hydroxy-2,2-bis(4chlorophenyl)acetic acid (aOH-DDA), which is excreted in the urine (Gold et aI., 1981; Gold and Brunk, 1982). Another route of metabolism of DDT in both the mouse and hamster (Gold and Brunk, 1982, 1983, 1984) seems to be the formation of DDA by a route involving hydroxylation on the C-l side chain carbon of DDD. Loss of HCl gives an intermediate acyl chloride, 2,2-bis(4-chlorophenyl)acetyl chloride (CI-DDA), capable of reacting with cellular proteins, DNA, etc., or losing water to give DDA. Since this work, the metabolism of DDT in rats has been reexamined (Fawcett et aI., 1981, 1987) and seems to be similar to that described above for hamsters and mice. The conversion of p,p'-DDD to p,p'-DDA occurs primarily by hydroxylation leading to CI-DDA, which on hydrolysis gives DDA. This acyl chloride may also be formed from DDE via an epoxidation route. Although DDMU is converted to DDA (Gold and Brunk, 1984; Fawcett et aI., 1987), there is now considerable doubt as to whether it is an important intermediate in DDT metabolism. In addition, there is evidence to suggest that DDOH is a reduction product of DDCHO formed directly from DDT and not a precursor. A current scheme for the metabolism of p,p'-DDT in rats is shown in Fig. 60.1 and is still probably incomplete. For instance, the role of DDOH still appears to be uncertain (Kujawa et aI., 1985). The interconversions of o,p'-DDT and p,p' DDT have been reported (Abou-Donia and Menzel, 1968; French and Jeffries, 1969; Klein et aI., 1965), but there is considerable doubt as to whether these occur in vivo (Cranmer et aI., 1972). Compared to p,p'-DDT, the more rapid excretion of o,p'DDT is explained at least in part by the observed ring hydroxylation of the parent compound in rats (Feil et aI., 1973) and chickens (Feil et aI., 1975) and of its metabolite o,p'-DDD in rats (Reif and Sinsheimer, 1975) and humans (Reif et al., 1974) (see Fig. 60.2). At least l3 metabolites were detected in rats. Ring hydroxylation, which has not been observed with p,p'-DDT or p,p'-DDD (but has been seen with p,p'-DDE), occurs in all species but with species differences. For example, o,p'-DDE and three hydroxylated o,p'-DDEs were found in the exc-
reta of chickens but not in the excreta of rats. In two patients with adrenal carcinoma for which they were receiving o,p'-DDD at a rate of 2000 mg/day, as much as 46-56% of the daily intake was recovered in the urine. Just over half of the recovered material was in the form of o,p'-DDA, but the remainder was in the form of hydroxylated derivatives, mainly m-hydroxy-, p-hydroxy-, m-hydroxy-p-methoxy-, and p-hydroxy-m-methoxy-o,p'-DDA. All hydroxylation had occurred on the ring that had its chlorine in the 0 position (Reif et al., 1974). When the metabolism of a single 100-mg oral dose of o,p'-e 4 C]DDD was studied in rats, averages of 7.1 and 87.8% of the activity were recovered in the urine and feces, respectively, within 8 days (Reif and Sinsheimer, 1975). The high recovery indicated rapid excretion with little storage. o,p'-DDD is specifically toxic for the adrenal cortex in a number of species including humans. In vitro studies suggest that this is due to its activation in adrenal mitochondria to a metabolite which binds covalently. Unlike the situation in liver, a metabolite more polar than DDA is also produced (Martz and Straw, 1977, 1980; Pohland and Counsell, 1985). More recently, Lund et al. (1988) showed that 3-methysulfonyl- p,p'DDE is selectively covalently bound and toxic to the adrenal zona fasciculata of mice. A single dose of 3-methylsulfonylp,p'-DDE to mouse dams caused high binding in the adrenals of suckling pups with extensive vacuolation and necrosis of the zona fasciculata. Slight degenerative changes were seen in fetal adrenals after dosing mothers with 50 mgikg (Jonnsson et aI., 1992). The binding and damage probably results from cytochrome P450 activation (CYP11B) in adrenal mitochondria (Jonnsson et aI., 1991, 1995). A similar cytochrome P450-mediated activation may account for the covalent binding of o,p'-DDD in mouse lung (Lund et aI., 1986, 1989) and may be related to the acyl chloride formation already reported for p,p'-DDT in rats and mice (Fig. 60.1). Cytochrome P450mediated activation to the acyl chloride has also been proposed as partly accounting for the toxicity of DDD to isolated rabbit Clara cells and human bronchial epithelial cells (Nichols et aI., 1995). Of the compounds shown in Figs. 60.1 and 60.2 only DDT, DDD, DDE, and DDA commonly are reported in the tissues or excreta of animals, including humans. Conjugates of DDOH with fatty acids in the livers and spleens of rats given DDT have been reported (Leighty et aI., 1980) and can be removed in vivo by treatment with bile salts, heparin, or lecithin (Leighty, 1981). Although microorganisms, plants, insects, and birds produce many of the same metabolites found in mammals, there are interesting differences. Nearly 20 derivatives (including mammalian metabolites) have been identified, and the chemical structures of others are still unknown. Some aspects of nonmammalian, as well as mammalian, metabolism have been reviewed (Fishbein, 1974; Klein and Korte, 1970; Korte, 1979; Menzie, 1969; Schroeder and Dorozalska, 1975). The metabolism of microorganisms and plants, as well as that of domestic animals, may influence the composition of DDTderived re sidues in human food, but there is no evidence that
60.2 DDT Cl
CCI3
Cl
CH30~", ~I
HO
1...-:;
lJJ~ Cl
-
4-Hydroxy-3-methoxy-o,p'-DDT
o,p'-DDT*
Cl
3-Hydroxy-o, p '-DDT
CHCI2
~ U V-- Cl
Cl
3-Hydroxy-o, p '-DDD* 4-Hydroxy-o, p '-DDD* 5-Hydroxy-o, p '-DDD*
o,p'-DDD
Cl
o,p'-DDMU
COOH
Cl
~ V--CI
HoM
H°tni'" HO
~
Cl
CHCI2
1
1
...-:;
Cl
3,4-Hydroxy-o, p '-DDD (4,5-Hydroxy also found)
Cl
-
/"1
HO
I'" ...-:;
~
Cl
3,4-Hydroxy-o, p '-DDA * (4,5-Hydroxy-o, p'-DDA* also found)
Cl
~
1
,
,..:;.
Cl
COOH
H°trU'" ~I
'...-:;
Cl
3-Hydroxy-o, p '-DDA (5-Hydroxy also found)
Cl
COOH
CH30 r 6 r U ' " HO
o,p'-DDA*
Cl
Hor6rU
OH
~ U ~CI
COOH
COOH
~ Cl UV--
4-Hydroxy-o, p '-DDA
Cl
CCI3
HO~ Cl
HO~
1315
COOH
I ~ ~
I'" ...-:;
Cl
OH
0,
p '-Dichlorobcnzhydrol
4-Hydroxy-3-methoxy-o, p'-DDA
5-Hydroxy-o, p '-DDA
Figure 60.2 Metabolism of o,p'-DDT in the rat. Compounds indicated by an asterisk have been found in humans, including those humans treated with large doses of o,p'-DDD. In rats, glycine and serine conjugates of o,p'-DDA have been found in the urine, and the aspartic acid conjugate of o,p'-DDA has been found in the feces.
these residues contain a significant amount of any compound not formed from DDT by human metabolism.
60.2.3.6 Excretion When large doses of DDT are ingested, some of the compound is unabsorbed. Only traces of unaltered DDT may be found in the feces when exposure is by any route other than oral. However, true fecal excretion of DDT metabolites was established irrespective of the route of administration (Hayes, 1965), although either DDT metabolites are not excreted by humans in the feces to any important degree, or they are excreted in one or more forms different from those demonstrated in rats. The bile appears to be the principal source of DDT metabolites in the feces of rats. When the bile duct was cannulated before intravenous injection of radioactive DDT, 65% of the dose was recovered in the bile, 2% in the urine, and only 0.3% in the feces (Jensen et aI., 1957). The different routes of excretion
are not unrelated. Burns et al. (1957) found that there was an increase in urinary excretion of radioactive material following ligation of the bile duct in rats fed radioactive DDT. This supports the finding by Jensen and his colleagues that most of the metabolites in bile are DDA or closely related to it. Although an enterohepatic circulation of the metabolites of DDT has not been proved directly, it seems likely that such a circulation exists. The difference between the excretion of DDT and its metabolites in rats and the slower excretion in birds seems to be the reduced ability of birds to further metabolize DDE (Fawcett et aI., 1981). The excretion of DDE in rats is dependent on dose and probably involves induction of drug-metabolizing systems (Ando, 1982). The excretion of DDT in milk was first published by Woodard et al. (1945) in connection with a dog fed at the rate of 80 mg/kg/day. Within a short time, excretion of DDT in milk was reported in rats, goats, and cows, and in 1951 it was demonstrated in women (Laug et aI., 1951). Telford and
1316
CHAPTER 60
DDT and its Analogs
Guthrie (1945) reported that rats fed a diet containing 1000 ppm produced milk toxic to their young. Following these early studies, the presence of DDT was demonstrated repeatedly in the milk of cows. Cows fed substantial, not nontoxic, residues of DDT commonly excrete 10% or more of the total dose in their milk (Hayes, 1959). The proportion of the mother's DDT intake that could be recovered from her milk varied from 12.6 to 30.2% and averaged 24.6% in rats receiving the compound from their diet at an average rate of 32.4 mg/kg/day. Under these circumstances, the dosage of the young was somewhat less than half of that of their mothers on a milligram per kilogram basis. The oral dosage of 32.4 mg/kg/day was well tolerated by both dams and pups, as was also true of an intraperitoneal dosage of 100 mg/kg/day. An intraperitoneal dosage of 200 mg/kg/day killed some dams, but most of the pups of other dams survived. All of the pups of these mothers experienced reduced milk intake and reduced weight gain. The concentration of DDT in the brains of these pups was much lower than in pups killed by oral administration of the compound, indicating that the young of mothers receiving massive dosages of DDT suffer malnutrition but not poisoning (Hayes, 1976). Wilson et al. (1946) showed that DDT was secreted from the skin of a cow maintained on an oral dosage of about 53 mg/kg/day. Because DDA is the main form in which DDT is excreted, it might be expected that, following its direct administration, DDA would be excreted relatively efficiently. During the first several days after oral dosing, rabbits excreted DDA in the urine approximately 15 times faster than animals given DDT at an equivalent dosage. Although the rate of DDA excretion associated with DDT increased more rapidly, so that the values differed by a factor of only 5 after day 20 of feeding (Smith et al., 1946). 60.2.3.7 Biochemical Effects The main mechanism of action of DDT is its effect on membranes in the nervous system, especially axonal membranes. The effect on axons may be related to inhibition of Na+ -, K+ -, and Mg2+ -adenosine triphosphatase derived from a nerve ending fraction of rabbit brain that is inhibited by DDT. A similar enzyme that binds DDT was isolated from the synapses of rat brain (Bratowski and Matsumura, 1972). There has been considerable interest in a Ca-ATPase which may regulate calcium levels at the axon surface (Ghiasuddin and Matsumura, 1979), and DDT is known to cause prolonged opening of the ion gates of the sodium channel perhaps by affecting phosphorylation in the a-subunit protein (Ishikawa et al., 1989). Song et al. (1996) showed that DDT appeared to interact with sodium channels of rat dorsal, not ganlion neurons, in the same manner as type I and type 11 pyrethroids. At a supralethal dosage of 600 mg/kg, DDT caused in rats a marked decrease in the concentration of cortical and striatal acetylcholine and of brain stem 3-methoxy-6-hydroxphenylglycol and 5-hydroxyindoleacetic acid (Hrdina et al., 1973; Hudson et al., 1985; Tilson et al.,
1986). p-Chlorophenylalanine blocked all of the neurotoxic signs of poisoning, and other inhibitors blocked one or another but not all of the effects. It was suggested that changes in the metabolism of 5-hydroxytryptamine and norepinephrine may be responsible for DDT-induced hypothermia and acetylcholine may be related to tremors and convulsions (Hrdina et al., 1973). Although spinal a-adrenoceptors have been proposed as modulating DDT-induced tremor (Herr and Ti1son, 1987), attenuation of DDT-induced motor dysfunction requires blockade of a-adrenoceptors in regions other than solely the spinal cord (Herr et al., 1989). At a lower dose of DDT (180 mg/kg), but one which still induced convulsive tremor, acetylcholine and cyclic GMP were increased in the cerebellum (Aldridge et al., 1978). In adult rats and mice there is a decrease in the cholinergic muscarinic receptors of rat brain (Eriksson et al., 1984), particularly in the cerebellum (Fonseca et al., 1986). The palmitic acid conjugate of DDOH can also have this effect (Eriksson and Nordberg, 1986). Disturbances of brain lipid metabolism have been observed in monkeys after chronic exposure to DDT (Sanyal et al., 1986). Khaikina and Shilina (1971) reported that administration of DDT to rats at only one-fifth of the LD 50 for 20 days increased by 188% the amount of 5-hydroxyindo1eacetic acid excreted in their urine. This indicated a change in the metabolism of serotonin, but it probably does not support a serotonin deficiency as a DDT mode of action (Chung et al., 1981). It is evident that many of the side effects of DDT are the result of its induction of drug metabolizing enzymes. Oral administration of o,p' -DDT to dogs at a rate of 50 mg/kg/day stimulates the microsomal enzymes of the liver. These changes in the liver are initially accompanied by an increase in the size of the adrena1s and of the cells of the zone faciculata; these cells become vacuolated and devoid of acidophilic cytoplasm, and their nuclei become hyperchromatic and often peripheral in position. Synthesis of cortocosteroids by the adreanal is not blocked (Copeland and Cranmer, 1974). Thus, the effect of a substantial dosage of o,p'-DDT is quite different from that of o,p'-DDD, although part of the metabolism of o,p'-DDT must be by that route. The tissue level of p,p'-DDE necessary to induce liver microsomal enzymes is lower in the rat than in the quail (and possibly in other birds). Thus Bunyan et al. (1972), using residues in the heart as an index, found a maximal increase in cytochrome P-450 per weight of liver and a maximal activity of aniline hydroxylase activity at tissue levels of approximately 3 ppm DDE in rats and 40 ppm DDE in quail. However, at any given dietary level, higher tissue levels were reached by quail than by rats, so the dosage responses of the two were similar. These authors concluded that DDE is more important that DDT in inducing microsomal enzymes, but in humans the opposite appears to be true. The significance of the induction of hepatic enzymes and any correlation with the potential hepatocarcinogenicity of DDT can be found elsewhere in this chapter and is particularly discussed in Smith (1991). In the female rat multiple bi-daily doses of DDT induced hepatic CYP2B and 3A proteins but not CYP1A1 or lE1 (Li et al., 1995) and caused el-
60.2 DDT
evated hydroxylation at 16 and 6 f3 of testosterone. DDT, DDE, and DDD all induced CYP2B and 3A in male rat liver to not disimilar degrees despite marked differences in bio retention (Nims et al., 1998). In squirrel monkeys (and presumably in other species) only 2 days on a vitamin C-deficient diet impairs both the induction of o-demethylase and the stimulation of the glucuronic acid system by DDT (5 mg/monkey/day) (Chadwick et aI., 1971). In guinea pigs, maintenance of induction of microsomal enzymes requires a higher dietary level of vitamin C than does prevention of scurvy (Wagstaff and Street, 1971). Since lipids are associated with the function of microsomal enzymes and DDT induces these enzymes, it might have been expected that DDT and essential fatty acids would interact. Tinsley and Lowry (1972) found that the growth of female rats receiving p,p'-DDT at a dietary level of 150 ppm was depressed if they received a diet deficient in essential fatty acids but was slightly stimulated if they received the same diet supplemented with these acids. It was suggested that DDT influenced essential fatty acid metabolism by increasing the demand for them. Sampson et al. (1980) found that DDT did not exacerbate aspects of essential fatty acid deficiency but did alter lipid metabolism in an unexplained way. Exposure of rats to DDT by the intratracheal route has shown lung lipid metabolic changes but the significance is unclear (N arayan et al., 1990a, b). In contrast, a variety of diets (containing fats that may occur in the human diet and that were in approximately the same proportion as fats in typical human food in the United States) had little or no influence on the storage of DDT and a wide range of pesticides fed to rats for four generations in combination at rates only 200 times those found in food in the United States (Adams et aI., 1974). Fat mobilization can cause rapid release of stored DDT, but this does not seem to be associated with any major toxic effect assessed pathologically or biochemically (Mitjavila et aI., 1981 b). DDT has been shown in vitro and sometimes in vivo to influence some enzymes of intermediary metabolism and other miscellaneous enzymes. For instance, DDT and a variety of analogs have been shown to affect isolated rat liver mitochondria but the significance of this in vivo is uncertain (Ohyama et aI., 1982). Whether this is linked to the effects caused by 3-methyl sulfonyl-p,p'-DDE in adrenal mitochondria is not known. The hyperglycemia observed during much of the early part of acute poisoning may be associated with an increase in four gluconeogenic enzymes (pyruvate carboxylase, phosphoenolpyruvate carboxykinase, fructose-1,6-diphosphatase, and glucose-6-phosphatase) (Kacew and Singhal, 1973). Increase in these enzymes in the renal cortex of rats have been observed after a single dose at a rate as low as 100 mg/kg or greater or following 45 daily doses at rates of 5 or 25 mg/kg/day. The changes are not mediated through release of corticosteroids from the adrenal glands. The fact that 100 mg/kg is the smallest single dosage that produced a statistically significant change in these enzymes indicates that their alteration is a complication rather than a cause of poisoning. High concentrations of DDT inhibit phosphatidase, muscle phosphatases, carbon an-
1317
hydrase, and oxaloacetic carboxylase and increase the activity of cytochrome oxidase and succinic dehydrogenase. However, none of these changes with the possible exception of inhibition of carbonic anhydrase appear to have any connection with the toxic action of DDT or even with its side effects (Hayes, 1959). Neal et al. (1944) reported a small but consistent increase in the volume of urine excreted in 24 hr when dogs were dosed orally or by insufflation at the rate of 100 mg/kg/day. The possibility that increased urinary output is related to the inhibition of carbonic anhydrase (Torda and Wolff, 1949) may deserve attention, but data from volunteers receiving 3.5 or 35 mg/person/day indicated no increase in urinary volume compared with controls (Hayes et al., 1971). Many enzymes including plasma amylase, aldylase, glutamic pyruvic transaminase, and isocitric dehydrogenase were not changed in squirrel monkeys given dosages from 0.05 to 50 mg/kg/day, the latter of which proved fatal within 14 weeks (Cranmer et aI., 1972).
60.2.3.8 Effects on the Nervous System The major toxic action of DDT is clearly on the nervous system, probably by slowing down closing of "gates" in axon sodium channels (Dubois and Bergman, 1977; Hong et aI., 1986; Woolley, 1982, 1985), and it requires an intact organism for full expression. For other biochemical mechanisms related to the nervous system see Section 60.2.3.7. The fact that DDT causes a myotonic response in muscle and substitution of a train of spikes for the normal diphasic electroneurogram (Eyzaguire and Lilienthal, 1949) is in marked contrast to the absence of detectable injury or, in fact, any response in other isolated tissues. In spite of the importance of the nervous system, a detailed review of early literature indicates that although the presence of some specialized nervous function may be necessary for the manifestation of DDT poisoning, the mere occurrence of specialized nervous fibers in certain protozoa or the occurrence of a rather complex nervous system in mollusks is not sufficient to render these forms susceptible. Just as there is little explanation still for the effect of DDT in susceptible species, the fact that certain species and even entire phyla are inherently resistant to the compound is still not entirely understood. A review (Hayes, 1959) of literature on the effects of DDT on the nervous system showed that all major parts, both central and peripheral, are affected. Whereas effects on specific portions, notably the cerebellum and the motor cortex, have been viewed as of greatest importance, it probably is more accurate to emphasize the interaction of functions, all modified to some degree. Farkas et al. (1968) found that electrocardiogram wave frequency showed considerable increase in resting rats that had received 20 mg/kg/day as a result of dietary intake. Rats that had received 5mg/kg/day did not exhibit this change while at rest, but even these exhibited abnormalities when exposed to a rhythmic light stimulus. Electrical activity may become abnormal only a minute or two after administration of a large dose of DDT. Four stages culminating in generalized seizure were
1318
CHAPTER 60 DDT and its Analogs
described by Joy (1973). Phenobarbital, but not diphenylhydantoin or trimethadione, was effective in stopping seizures. The most characteristic effect of DDT in contrast to dieldrin, for example, is the production of tremor. Sufficient dosages of DDT produce tremor even at ambient temperatures that approach body heat. However, dosages of DDT that produce no other clinical effect make rats more sensitive to low temperatures. This sensitivity may be demonstrated by having the rats swim to exhaustion in cool water. The ability of the rat to keep afloat is more dependent on coordination than on physical strength. DDT appreciably reduces the swimming time (Hayes, 1982; Smith, 1991). Like tremor, the coldness of the skin and ruffling of the fur seen in acute poisoning probably represent an indication of disturbed thermal regulation. Apparently, it was not until the work of Hrdina et al. (1975) that an increase of almost 3°C in body temperature was reported in rats following a fatal (600 mg/kg) oral dosage of DDT. The central nervous systems of mice and hamsters are equally sensitive, the concentration of DDT in their brains at death being similar (Gingel and Wallcave, 1974). However, after an oral dosage of 500 mg/kg, the DDT concentration of the mouse brain was twice that of the hamster. This cannot be explained by a difference in absorption, metabolism, or excretion but apparently is due to a difference in permeability of the blood-brain barriers of the two species. When animals received DDT at a dietary level of 205 ppm for 6 weeks, the residues in fat and liver were seven to eight times higher in the mouse, a fact only partially explained by the greater food intake of mice relative to body weight. Although urinary excretion of [14C] DDT was similar in previously unexposed hamsters and mice, this excretion was stimulated in the hamster but little affected in the mouse by previous dietary exposure to DDT. Careful studies have shown that neurotoxic actions of DDT following oral dosing of rats are significantly affected by the volume of the dosing solution. Not surprisingly this is probably the consequence of higher partitioning of DDT in oil and greater gut motility but it illustrates the importance of pharmacokinetic knowledge in comparative studies (McDaniel and Moser, 1997). Both DDT and DDE interact similarly with model and native membranes, causing disordering effects in cholesterol rich membranes, such as brain microsomes (Antunes-Madeira et al., 1993; Antunes-Madeira and Madeira, 1993). Whether these effects have any bearing on neurotoxicity in vivo is unknown.
60.2.3.9 Cause of Death Death from DDT poisoning is usually the result of respiratory arrest. The heart continues to beat to the end and in some instances continues a little while after respiration stops. Deichmann et al. (1950) found that the onset of hyperirritability was accompanied by an increase in the frequency and amplitude of respiration. Later, with the occurrence of tremors, the depth of respiration frequently returned to a more normal level, but the rate remained high. In some animals respiration stopped suddenly after a deep inspiration during a tonic convulsion. In other
animals the rate and amplitude decreased progressively and finally ceased without any terminal spasm. Animals that die of respiratory failure caused by DDT do so after a relatively long period of muscular activity that leaves them exhausted. It was shown by Phillips and Gilman (1946) and Phillips et al. (1946) that the hearts of dogs given large intravenous doses of DDT were sensitized to epinephrine. This was true not only of injected epinephrine but also of the compound released by the adrenal glands during a seizure. Stimulated in this way, the sensitized hearts of dogs developed an irreversible, fatal ventricular fibrillation. However, the hearts of monkeys were able to recover from fibrillation and resume normal rhythm. It is not clear how important sensitization of the myocardium is when DDT is administered by other routes, but ventricular fibrillation may be the cause of death in animals that die suddenly soon after onset of poisoning. Thus, DDT not only sensitizes the myocardium in a way similar to that of halogenated hydrocarbon solvents but also, through its action on the central nervous system, produces the stimulus that increases the likelihood of fibrillation. There is no evidence that repeated, tolerated doses of DDT sensitize the heart (Jeyaratnam and Forshaw, 1974). Rats were fed DDT at a dietary level of 200 ppm (about 10 mg/kg/day) for 8 months, during which they received weekly intraperitoneal doses of vasopressin which causes a temporary myocardial ischemia. Electrocardiograms showed no significant increase in cardiac arrhythmias in the DDT-fed rats compared with controls. Intravenous noradrenaline given at the end of the 8-month period did not produce a greater incidence of arrhythmias in the DDT-fed rats.
60.2.3.10 Mutation and Carcinogenesis DDT has been tested in a number of ways for possible mutational effect. Much of this work has been reviewed in detail together with most of the carcinogenicity studies shown in Table 60.4 (Coulston, 1985). For example, Shirasu et al. (1976) listed DDT as a negative chemical in microbial mutagenicity screening studies including metabolic activation systems with both DDT and DDE (McCann and Ames, 1976; Shirasu et al., 1977). At a dosage of 105 mg/kg DDT produced no increase of dominant lethals in mice (Epstein and Shafner, 1968). However, concentrations of 10 ppm or greater produced chromosome breaks and exchange figures in a marsupial somatic cell line (Palmer et al., 1972). A slight mutagenic effect in mammals has been reported by Markarian (1966). Deletions plus gaps were reported to be more common in the chromosomes of mice that had received DDT. On the whole, in vitro tests of the mutagenicity of DDT have given only negative or dubious results (Coulston, 1985). An unconventional test for mutagenicity involved examination of explants of pulmonary tissue from embryonic mice whose dams had been fed dietary concentrations of 10 and 50 ppm DDT (Shabad et aI., 1972). An increase of diffuse hyperplasia and focal proliferation was observed, but a dosageresponse relationship was not clear. Some of the embryos were
60.2 DDT
allowed to live and the experiment was repeated in subsequent generations. There was no continuing progression of the reported changes in succeeding generation. The question of whether DDT is carcinogenic really seems to be restricted, experimentally, to its action in the liver of some rodents. Some of the positive findings shown in Table 60.4 have not been found in other studies (National Cancer Institute, 1978a). However, there is still the evidence that DDT can act as a promoter of liver carcinogenesis initiated by aflatoxin and of other chemicals in vitro and in vivo (Peraino et a!., 1975; Rojanapo et a!., 1987; Schulte-Hermann, 1985; Sugie et a!., 1987; Williams and Numoto, 1984). DDT causes inhibition of intercellular communication in cultured rat liver epithelial cells and in hamster cell lines (Flodstrom et al., 1990; Tsushimoto et a!., 1983; Warngard et aI., 1989; Williams et aI., 1981). This could be protected against by extracts of green tea (Sigler and Ruch, 1993). Using freezefracture analysis of hepatocytes from rats exposed in vivo to DDT Sugie et at. (1987) showed that both gap junction size and number were reduced. Tateno et a!. (1994) provided evidence that in these in vivo experiements with rats changes in connexin 32 and connexin 26 expression might be responsible. Additional studies with the rat "oval" cell line WB-F344, which does not express connexin 32 but rather connexin 43 predominantly, showed not a decreased expression but a decrease in the phosphorylated form (Ruch et aI., 1994). Subsequent investigations suggest this is due to endocytosis of gap junctions and lysosomal connexin 43 P2 degradation (Guan and Ruch, 1996). Some evidence suggests that o,p'-DDT can support the growth of an estrogen-responsive tumor (Robison et aI., 1985a). The significance of the effects of DDT and other chlorinated hydrocarbon insecticides on the liver in relation to hepatocarcinogenicity have been discussed in detail previously (Smith, 1991) and will not be discussed here.
1319
those with 10 or more years of occupational exposure whose plasma DDT levels were reported that after a single large dose (100 mg/kg) to rats thyroidal 131 I release was completely inhibited for more than 12 hr. The view of Clifford and Weil (1972) was that there was no evidence that occupational exposure to DDT has had any effect on human endocrine organs. What at first appeared to be an immunological response to DDT in guinea pigs really involved a quite different, predictable effect. Animals sensitized to diphtheria toxoid were less susceptible to anaphylaxis in response to a challenge dose of the toxoid if they were pretreated with DDT at a dosage of only 1020 mg/kg/day (Gabliks et aI., 1973, 1975). Direct measurement of antitoxin production indicated little or no difference between protected and unprotected animals. Furthermore, some protection was given by DDT administered for only 3 days prior to the induction of anaphylaxis. Further study showed that DDT treatment reduced the histamine levels in the lungs of both immunized and nonimmunized animals. The number of detectable mast cells was also reduced; this was true whether the count was made in tissues from guinea pigs dosed systemically with DDT or in lungs and mesenteries from untreated animals exposed to DDT in vitro at concentrations ranging from 10 to 45 ppm. These results indicated that the protection offered by DDT was the result of a reduction of the amount of histamine available for sudden release in response to a challenge dose of toxoid (Askari and Gablicks, 1973). Regardless of exposure to DDT, immunization leads to an increase in detectable mast cells (Gabliks et aI., 1975). Banerjee et al. (1996) have compared DDT with some of its metabolites in suppressing aspects of humoral and cellular immune response in rats: DDE and DDD but not DDA appeared to play a role. DDT has been reported to cause acute renal failure in rats after intravenous infusion (Koschier et a!., 1980). 60.2.3.12 Effects on Reproduction
60.2.3.11 Other Miscellaneous Effects on Organs and Tissues Many early reports reviewed by Hayes (1959) indicate that large doses of DDT may have no effect on the blood or they may produce a moderate leukocytosis and a decrease in hemoglobin, with or without a decrease in the concentration of red cells. The leukocytosis probably is secondary to stimulation of the sympathetic nervous system, while the loss of hemoglobin may be nutritional in origin. A later study with squirrel monkeys did not confirm the early results (Cranmer et aI., 1972). A range of hematologic parameters remained unchanged in squirrel monkeys dosed orally at rates of 0, 0.05, 0.5, 5, and 50 mg/kg/day, even though the highest dosage was fatal within 14 weeks. Average protein-bound iodine levels of 5.42 and 6.93 J.Lg/%, respectively, were reported in the sera of 42 workers occupationally exposed to chlorinated hydrocarbon insecticides and 51 workers not so exposed. The differences were statistically significant even though all values fell within the normal range of 4-8 J.Lg/% (Wassermann et aI., 1971). It was not recorded whether the workers involved were from the same factory as
In recent years the possibility that environmental levels of DDT isomers and metabolites might have effects on human reproductive function and carcinogenesis in breast and testes has become a major issue. It was shown very early that DDT produces a striking inhibition of testicular growth and secondary sexual characteristics of cockerels when injected subcutaneously in dosages as high as 300 mg/kg/day (Burlington and Linderman, 1950). Changes in the testis involve the tubules and not the interstitial tissues, and they have been attributed to an estrogen-like action of DDT. Before the problem of residues became evident, DDT was used extensively for control of lice and common mites on chickens without any adverse effects on egg production or other aspects of reproduction. It must be emphasized that many rats would be killed the first day if they were given the dosage of DDT that has been shown to affect the testis of cockerels. The report that under special conditions DDT has a gonadotoxic effect is of questionable significance in view of the results of multigeneration tests in rats, mice, and dogs (Rybakova, 1968). Dean et al. (1980) were unable to demonstrate any changes in either serum androgens or testicular synthesis of
1320
CHAPTER 60
DDT and its Analogs
testosterone in young rats after exposure to DDT despite significant induction of metabolism of testosterone by isolated hepatic microsomes. However, more recent evidence supports the view that DDT and its metabolite DDE can disrupt male reproductive development and act as anti androgens by binding androgen receptors in a nonproductive manner (Kelce et aI., 1995, 1998). DDE, like vinclozolin and flutamide, changed the expression of androgen receptor regulated genes, such as the prostatic mRNA prostatein C3, in castrated-tetosterone treated male rats (Kelce et aI., 1997). Exposure of male rats to DDE in utero and by lactation showed that by some parameters Long-Evans rats were more sensitive than Sprague-Daw1ey, e.g., ano-genital distance, but DDE had no effects on testes, epididymus, seminal vesicles, or vental prostrate weights (You et aI., 1998). Effects were minimal at maternal doses <10 mg/kg/day. Confirmation that Long-Evans rats were more sensitive to DDE than some other strains was reported from studies designed to detect antiandrogen endocrine disrupters and appeared to act centrally rather than peripherally (O'Connor et aI., 1999). The effects of p,p'DDE on the development of the rat prostate was confirmed with Holtzman rats exposed in utero or by lactation (Loeffler and Peterson, 1999) without changes in serum androgen concentrations. This suggests effects on androgen signalling pathways within the prostate. Some evidence for a partial interaction with the antiandrogenic effects of TCDD was suggested from dual dosing experiments. The significance of these studies with regard to human male development is really not clear given the considerably greater experimental exposures. Intraperitoneal injection of as little as 5 mg/kg of technical DDT or 1 mg/kg of o,p'-DDT causes a significant increase in weight of the uterus of normal immature female rats or of ovariectomized adult females (Welch et aI., 1969). A very much smaller stimulation is caused by p,p'-DDT. Treatment of rats with DDT, especially o,p' -DDT, inhibited uptake of the hormone by the uterus in vivo, possibly by competition for binding sites. Isomers ofDDD and DDE do not influence uterine weight or the binding of estradiol (Welch et aI., 1969). It seems unlikely that metabolic activities of o,p' -DDT is necessary as is true of activation of o,p' -methoxychlor (Kupfer and Bulger, 1979). The action of o,p' -DDT on the uterus seems to be as a long-acting agonistic estrogen interacting with the same receptor as 17,B-estradiol (Galand et al., 1987; Ireland et aI., 1980; Robison et aI., 1984). However, some differences from estradiol have been recorded (Robison et aI., 1985b). The lesser enantiomer of o,p'-DDT seems to be the active isomer (McBain, 1987). The binding and estrogenic activity of DDT analogs in rats is only about 1/10,000 as great as that of diethylstilbestrol (Nelson, 1973), an important point when considering potential effects of trace levels relative to endogenous oestrogens. o,p'DDT inhibited DNA synthesis in cultured bovine oviductal and uterine cells to a greater degree than methoxchlor, especially uterine epithelial and stromal cells (Tiemann et aI., 1996). However, at lower concentrations both o,p' -DDT and methoxchlor stimulated DNA synthesis. A considerably smaller dosage of o,p' -DDT resulting from a dietary level of 10 ppm for 2-9 months had no effect on repro-
duction in ewes (Wrenn et aI., 1971a). In a similar way, dietary levels of o,p'-DDT as high as 40 ppm, giving a dosage level of about 2.1 mg/kg/day in rats, failed to interfere with reproduction and lactation in these animals, although dosage was continued through two pregnancies (Wrenn et al., 1971b). The report (Heinriks et aI., 1971) that o,p'-DDT significantly advances puberty, induces persistent vaginal estrus after a period of normal estrus cycles, and causes other reproductive abnormalities in female rats would at first appear inconsistent with the lack of effect of technical DDT or o,p'-DDT on reproduction cited above. The same is true of other effects of o,p' -DDT subcutaneously on the second, third, and fourth days of life (counting the day of birth as zero). Because rat pups on the third day weighed about 12 gm or less each, it follows that the subcutaneous dosage was about 83.3 mg/kg/day or more, that is, about 40 times greater than the highest oral dosage of o,p' -isomer fed to breeding rats and more than 105 times greater than human dietary exposure. Ottoboni (1969) found that female rats reproduced normally when fed DDT for two generations at dietary levels as high as 200 ppm (about 10 mg/kg/day except during lactation, when intake is increased about threefold). In fact, at a dietary level of 20 ppm, the dams had a significantly longer reproductive life span (14.5 months) than their littermate controls (8.9 months); the number of females becoming pregnant after the age of 17 months and the number of successful pregnancies after that age were significantly different in the two groups (Ottoboni, 1972). In a study focused mainly on DDT in milk, the full ability of rats to reproduce at a dietary level of 200 ppm was confirmed, and the ability of dams injected intraperitoneally at levels as high as 100 mg/kg/day to rear their young was demonstrated (Hayes, 1976). A six-generation test of reproduction in mice showed no effect of DDT at a dietary level of 25 ppm on fertility, gestation, viability, lactation, and survival (Keplinger et aI., 1970). A level of 100 ppm produced a slight reduction in lactation, and survival in some generations but not all, and the effect was not progressive. A level of 250 ppm was distinctly injurious to reproduction. The dietary concentrations used determine dosages of 3.33,13.3, and 33.2 mg/kg/day in nonpregnant, nonlactating, adult, female mice. The intake is much higher in both young and lactating mice. Four female dogs of unstated age that previously had received DDT at the rate of 12 mg/kg/day, 5 days/week, for 14 months were bred when they went into heat (Deichmann et aI., 1971). The males involved had been fed aldrin (0.15 mg/kg/day) plus DDT (60 mg/kg/day) for 14 months prior to breeding but not during breeding. Two of the females went into heat but failed to become pregnant, and one failed to come into heat during 12 months after feeding stopped. Four of six pups born to the fourth female died within 1 week of birth; the other two were weaned successfully even though only two posterior mammae of the mother were functional. A three-generation study failed to confirm any of the injuries suggested by the study of four dogs. In the three-generation study, male and female dogs were fed technical DDT from weaning at rates of
60.2 DDT
0,1,5, and 10 mg/kg day. Observations were made on 135 adult females, 63 adult males, and 650 pups. There were no statistically significant differences among controls and DDT-treated dogs in length of gestation, fertility, success of pregnancy, litter size, or lactation ability of the dams; in viability at birth, survival to weaning, sex distribution, and growth of pups; or in morbidity, mortality, organlbody weight ratios, or gross histological abnormalities in all the animals studied. The only clear difference was that DDT-treated females had their first estrus 2 or 3 months earlier than the control dogs. There was a slight increase in liver/body weight ratio in some DDT-treated animals but the difference was not statistically significant, not dosage related, and not associated with any histological change (Ottoboni et aI., 1977). When p,p'-DDTwas administered to pregnant mice at a rate of I mg/kg on days 10, 12, and 17 of gestation, it was not teratogenic but did alter the gonads and decrease the fertility of the young, especially the females (McLachlan and Dixon, 1972). A single dose at the rate of 15 mg/kg or repeated doses of 2.5 mg/kg/day given during pregnancy may be embryotoxic but not teratogenic to mice (Schmidt, 1973). DDT was shown to be more toxic than methoxychlor to preimplantation mouse embryos in culture (AIm et al., 1996). Teratogenic effects of DDT have not been seen in studies of reproduction, including those for two generations in rats, six generations in mice, and three generations in dogs (Smith, 1991). Because of the estrogenic properties of large doses of DDT, the compound was considered as a possible cause of abortion in dairy cattle, but no evidence for a relationship was found (Macklin and Ribelin, 1971). A similar conclusion was reached regarding human abortions (O'Leary et aI., 1970). 60.2.3.13 Behavioral Effects Behavioral changes may be demonstrated in animals receiving DDT daily at rates too low to produce illness. Khairy (1959) detected ataxia in the form of changes in gait in rats that had been fed DDT at dietary levels of 100 ppm or more for 21 days. The results were recorded in terms of the tangent, that is, the ratio of the width and length of step. At a dosage of about 5 mg/kg/day the ratio was less than normal, a change attributed to an exaggeration of the stretch reflex. At dosages of about 10, 20, and 30 mg/kg/day, the ratio was progressively increased above normal as a result of broadening of the gait and shortening of the steps. These same dosage levels did not affect problem-solving behavior or speed of locomotion. The experimental animals were found to be generally less reactive to stress than normal ones. The acoustic startle response of rats is significantly increased after a 12.5 mg/kg dose of p,p'-DDT but can be attenuated by phenytoin and an adrenergic receptor antagonist, phenoxy benzamine (Herr and Tilson, 1987; Herr et aI., 1987; Saito et aI., 1986; Tilson et aI., 1985, 1985), which also decreased DDTinduced myoclonus (Hwang and Van Woert, 1978). See also Section 60.2.3.8.
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60.2.3.14 Pathology Morphological changes are inadequate to account for death from DDT poisoning. Changes that occur in the liver have been discussed previously (Smith, 1991). Mild to moderate morphological changes have been reported in the kidneys of animals that had received massive single doses or repeated doses; examples are fatty degeneration, necrosis, and calcification (Lillie et al., 1947; Stohlman and Lillie, 1948) or slight brown pigmentation of the convoluted tubular epithelium (Fitzhugh and Nelson, 1947). However, it sometimes has happened that a complete absence of change in the kidney has been reported in connection with other studies carried out in the same laboratories (Lillie and Smith, 1944; Nelson et al., 1944). 60.2.3.15 Treatment of Poisoning in Animals The more successful studies of treatment of animals poisoned by DDT involve the nervous system. A full discussion of this can be found in Smith (1991). In essence sodium phenobarbitol affords little help in rats but in dogs and cats, to a lesser extent monkeys, it protected against tremors and death (Phillips and Gilman, 1946). More recently, Tilson et al. (1985, 1986) have reported that phenytoin attenuates the tremor produced in rats by DDT and permethrin but not by lindane and chlordecone. Vaz and his colleagues (1945) were apparently the first to note the antidotal effect of calcium in DDT poisoning. Dogs were given DDT orally as a 10% oily solution at a daily dosage of 100 mg/kg until signs of intoxication appeared. The same dosage could then be repeated to produce intense symptomatology from which the animals would recover spontaneously in 12-24 hr. For the actual tests, a larger challenge dosage of DDT (150-200 mg/kg) was used. Each dose of calcium gluconate (30 ml of a 10% solution) was injected intravenously into dogs weighing 8-18 kg. Dogs that were injected with calcium gluconate daily for 4 days and challenged with a large dose ofDDT on the fourth day developed no symptoms or only slight ones. Dogs receiving a single dose of calcium gluconate showed symptoms of short duration and survived following a dosage of DDT large enough to kill two controls. Koster (1947) studied cats poisoned by the intravenous injection of a soya lecithin-corn oil emulsion of DDT. A comparison was made of several aspects of intoxication, including number of convulsions, general severity (tremors, prostration, dyspnea), duration, and mortality. Both calcium gluconate and sodium gluconate reduced mortality but not severity. Gluconic acid increased the survival time and reduced mortality but did not reduce convulsions or severity. Calcium chloride reduced convulsions but not mortality or tremors. The lifesaving capacity of calcium gluconate at a rate of 40 mg/kg was confirmed by ludah (1949), even though he found normal blood calcium values in most poisoned but unmedicated animals. One animal showed a high calcium value, and Cameron and Burgess (1945) reported a similar result. Calcium has, then, an antidotal action against DDT in intact animals of several species. The hypothesis has been advanced that certain neurotoxins, including DDT,
1322
CHAPTER 60 DDT and its Analogs
act by delaying the restoration of calcium ions to a surface complex following breaking of the chelate linkage of calcium ions to surface polar groups by an initial exciting impulse (Gordon and Welsh, 1948). This action of the neurotoxin is conceived as depending largely on its physical rather than on its chemical properties. The hypothesis is still helpful in explaining the fact that a wide variety of chemically unrelated compounds produce repetitive responses in excitable tissue and also the fact that many compounds that show a high toxicity for arthropods and mammals are fat-soluble and chemically relatively inert. On the other hand, calcium may help to offset the effects of DDT on calcium-dependent ATPases, especially in the neuronal axons (see Section 60.2.3.7). Having observed the effect of DDT on the metabolism of glucose and glycogen, Laiiger and colleagues (1945a, b) investigated the use of glucose as an antidote. All of the 10 dogs given 2000 mg of DDT per kilogram of body weight orally in the form of an oil solution died within 8-24 hr. Five of the 10 dogs treated with one or more 20-ml doses of 20% glucose survived the same dosage of DDT. Koster (1947) found that glucose given before or after a LD 33 dosage reduced convulsions and mortality and, when given before the poison, reduced tremors, prostration, and dyspnea in cats, but was ineffective against a LD 95 dosage except to increase the time of survival.
60.2.4 TOXICITY TO HUMANS 60.2.4.1 Experimental Oral Exposure Table 60.5 summarizes the effects of one or a few oral doses of DDT. The results are consistent with those in accidents reported by Garrett (1947) and Hsieh (1954) in which it was possible to estimate accurately the amount ingested. It may be concluded that a single dose of 10 mglkg produces illness in some but not all subjects even though no vomiting occurs. Smaller doses generally produce no illness. Persons who were made sick by 10 mg/kg did not show convulsions, but convulsions have occurred in accidents when the dosage level was 16 mg/kg or greater (Hsieh, 1954). Rarely, a dosage as high as 20 mg/kg may be taken without apparent effect (MacCormack, 1945). Dosages at least as high as 285 mg/kg have been taken accidentally without fatal result (Garrett, 1947). However, large doses lead to prompt vomiting so the amount actually retained cannot be determined accurately. In acute poisoning a slight decrease in hemoglobin and a moderate leukocytosis without any constant deviation in the differential white count have been observed in volunteers (Velbinger, 1947a, b). These findings are considered secondary to the neurological effects. In the course of tests with volunteers, dilute colloidal aqueous suspensions of DDT are apparently odorless and tasteless (Domenjoz, 1944; Hoffman and Lendle, 1948). Saturated alcoholic solutions of DDT have a weak aromatic taste, or rather odor. Some people find these solutions slightly anesthetic to the
tongue (Hoffman and Lendle, 1948). The taste of DDT in vegetable oil is so slight that many persons cannot identify capsules containing 0, 3.5, and 35 mg of DDT when they are presented separately but can arrange them in proper order when one of each is available for comparison. The possible clinical effects of many repeated doses of DDT were first explored by Fennah (1945). Because of his interest in predicting the results of indiscriminate use, he expressed the exposures in terms of environmental levels rather than in dosage units. The exposures were clearly higher than those ordinarily encountered. In one test, lasting a total of 11.5 months, Fennah daily inhaled 100 mg of pure DDT and drank water dusted at the rate of 3240 mg/m2 . Much of the inhaled dust must have been deposited in the upper respiratory tract and swallowed. Later, for 1 month, Fennah ate food all of which had been sprayed at the rate of 2160 mg/m 2 after it had been served. No ill effect of any kind was observed. However, it must be said that these days we would examine more closely possible subtle chronic effects. Some studies of DDT in volunteers have been designed to explore the details of storage and excretion of the compounds in humans and to search for possible effects of doses considered to be safe. In the first of these studies, men were given 0,3.5, and 35 mg/person/day (Hayes, 1969). These administered dosages plus DDT measured in the men's food resulted in dosage levels of 0.0021-0.0034,0.038-0.063, and 0.36-0.61 mg/kg/day, respectively, the exact value depending on the weight of each individual. Six volunteers received the highest dosage of technical DDT for 12 months, and three received it for 18 months. A smaller number of men ingested the lower dosage of technical DDT or one of the dosages of p,p'-DDT for 12-18 months. No volunteer complained of any symptom or showed any signs of illness that did not have an easily recognizable cause clearly unrelated to the exposure of DDT. At intervals, the men were given a systems review, physical examination, and a variety of laboratory tests. Particular attention was given to the neurological examination and liver function tests, because the major effects of DDT in animals involve the nervous system and the liver. The same result was obtained in a second study in which the same dosages were given for 21 months and the volunteers were observed for a minimum of 27 additional months (Ha yes et aI., 1971). In the first study, the storage of DDT was proportional to dosage, but there was a then unexplained difference in the storage of the p,p' -isomer and of technical DDT. Following dosing for 12 months, the pure material was stored in fat at an average concentration of 340 ppm, but the technical material was stored at an average of only 234 ppm. The difference was statistically significant for the 3.5 mg/person/day dosages given for 3-6 and for 7-18 months. The difference was significant for the 35 mg/person/day doses after 7-18 months of dosing but not after only 3-6 months. Men who ate p,p' -DDT showed a definite increase in the absolute amount of DDE stored. After 6 months at a dosage of 35 mg/person/day, eight men showed an average DDE fat storage of 32.6 ± 7.0 ppm as compared to 12.3 ± 1.5 ppm for the same individuals upon entering the investigation. There was a further increase of DDE storage as
60.2 DDT
1323
Table 60.5 Summary of the Effects of One or a Few Oral Doses of DDT on Volunteers Dose (mg) and Result
Reference
2S0 x 9, suspension
no effect
Domenjoz (1944)
lS,OO butter solution
no effect, but mice killed when fed 6 and 12 hr after dose
MacCormack (194S)
SOO, oil solution
no clinical effect
Neal et al. (1946)
770, oil solution
no clinical effect; DDA measured in urine
Neal et al. (1946)
2S0, suspension
none except slight disturbance of sensitivity of mouth
Velbinger (l947a, b)
2S0, oil solution
variable hyperesthia of mouth
Velbinger (1947a, b)
SOO, oil solution
variable hypercsthia of mouth
Vclbingcr (l947a, b)
7S0, oil solution
disturbance of sensitivity of lower part of face; uncertain gait; peak
Velbinger (1947a, b)
formulation
reaction (6 hr after ingestion) characterized by malaise, cold moist skin, and hypersensitivity to contact; reflexes normal 1000, oil solution
same as above; no joint pains, fatigue, fear or difficulty in seeing
Velbinger (l947a, b)
or hearing lSOO, oil solution
prickling of tongue and around mouth and nose beginning 2.S hr after
Velbinger (1947a, b)
dose; disturbance of equilibrium; dizziness; confusion; tremor of extremities; peak reaction (l0 hr after ingestion) characterized by great malaise, headache, and fatigue; delayed vomiting; almost complete recovery in 24 hr
exposure progressed. However, DDT was stored in so much greater concentration that the relative storage ofDDE decreased sharply. Thus, after 6 months at a dosage of 35 mg/person/day, eight men stored only 14% of their total DDT-derived material in the form of DDE as compared to 65% for the same persons at the beginning of the investigation. The storage of DDE by men who ate technical DDT presented a different picture. Until 18 months after exposure, there was no clear evidence that these men stored any more DDE after exposure than they did before. However, at 18 months the only three samples available showed DDE concentrations ranging from 28 to 85 ppm, all substantially above general population levels. Thus, both the total amount stored and the rate at which DDT converted to DDE served to distinguish the metabolism of p,p'-DDT and technical DDT in humans (Hayes et aI., 1956). This was true even though later study showed that the concentration of DDE in serum increased immediately in persons ingesting technical DDT at rates of 10 and 20 mg/person/day. Of course, daily values were subject to considerable variation, but the upward slopes of the graphs recording the results were apparent in 60 days or less and apparently the graphs were straight throughout the 50 month feeding period. Under the same conditions, the level of DDT in serum increased within 1 day and continued to increase in a curvilinear fashion for 5 months (Apple et aI., 1970). A similar rapid increase reaching its maximum in 30 hr after a single exposure has been observed in workers (Apple et aI., 1970). The more rapid excretion of o,p'-DDT was demonstrated by Morgan and Roan (1972). In a second study in which the volunteers received 0, 3.5, and 35 mg/person/day, the storage of DDT was again proportional to dosage with the real but very gradual accumulation
of DDE (Hayes, 1969, 1982; Smith, 1991). A steady state of storage was approached later in the second study (18.8-21.5 months) than in the earlier one (about 12 months). The second study was superior in that more men were observed for a longer period but inferior in that dosing was less regular. Because of the latter difficulty, it seems impossible to decide whether 12 months or 21.5 months is a more valid estimate of the time necessary for people to approach a steady state of storage when intake is uninterrupted and unvarying in amount. It is interesting that the storage levels eventually reached at the same dosage in the two studies were statistically indistinguishable in most instances. In the one instance in which a statistical difference existed, the greater storage by men in the second study may have been explained by the fact that some of them inadvertently received higher doses than intended. DDT was lost slowly from storage in fat after dosing was stopped. The concentration remaining following 25.5 months of recovery was from 32 to 35% of the maximum stored for those who had received 35 mg/person/day but was 66% for those who had received only 3.5 mg/person/day, indicating slower loss at lower storage levels (Hayes et aI., 1971). Morgan and Roan (1971) fed volunteers not only technical DDT but also p,p'-DDD. They found that DDE is stored more tenaciously than the other compounds in humans, the order being p,p'-DDE > p,p'-DDT > o,p'-DDT 2: p,p'-DDD. The slow metabolism of DDT to DDE was confirmed. It was noted that p,p' -DDT is lost from storage in adipose tissue much more slowly in humans than in the monkey, dog, or rat. Less than 18% of p,p' -DDE is carried in human erythrocytes. In plasma of ordinary fat content, less than 1% of all DDT-related compounds is carried by the chylomicrons. p,p'-
1324
CHAPTER 60
DDT and its Analogs
DDT and p,p'-DDE are found mainly in the triglyceride-rich, low-density, and very low-density lipoproteins. Following continuous electrophoresis, these compounds are found mainly in association with plasma albumin and a-globulins (Morgan and Roan, 1972). DDA is the main urinary metabolite of DDT. In humans, it was found first in a volunteer by Neal et at. (1946), who reported that, following ingestion of 770 mg of p,p' -DDT, excretion rose sharply to 4.0 mg/day during the second 24-hr period, decreased rapidly on the third and fourth days, decreased gradually thereafter, but was still above baseline on day 14. Later studies in volunteers who received smaller but repeated doses confirmed the very rapid rise in excretion of DDA (Hayes et aI., 1971; Roan et al., 1971) and showed that a steady state of excretion was reached after about 6-8 months. During a 56week period of continued dosing after equilibrium was fully established, the concentration of DDA associated with technical DDT at the rate of 35 mg/person/day varied from 0.18 to 9.21 ppm and averaged 2.98 ppm; corresponding values for p,p' -DDT were 0.40-6.27 ppm with a mean of 1.88 ppm. Thus technical DDT, as compared to p,p' -DDT, was excreted more effectively and stored less. During the latter half of the dosing period, it was possible in the two groups receiving recrystallized and technical DDT at the rate of 35 mg/person/day to account for an average of 13 and 16%, respectively, of the daily dose in terms of urinary DDA. The excretion of DDA was relatively constant in each individual, but marked differences were observed between men receiving the same dose (Hayes, 1969; Smith, 1991). 60.2.4.2 Experimental Dermal Exposure Depending on dosage, oral administration of DDT to volunteers has produced either no illness or brief poisoning entirely similar to that seen in experimental animals. The oral dosage necessary to produce any clinical effect was almost always 10 mg/kg or more. It is a strange coincidence that, in two studies involving only three subjects in all, experimental dermal exposure to DDT was followed by fatigue, aching of the limbs, anxiety or irritability, and other subjective complaints. Recovery was delayed a month or more (Case, 1945; Wigglesworth, 1945). In neither study was there an independent control. Although the dosage was unmeasured, the amounts of DDT absorbed must have been much smaller than those involved in the oral studies. One of the studies involved self-experimentation by one man. A somewhat more severe test on six volunteers produced no toxic or irritant effect at all (Dangerfield, 1946). In view of all other experiments and extensive practical experience, it is probable that the illnesses reported by Wigglesworth and Case were unrelated to DDT. With the exceptions just mentioned, dermal exposure to DDT has been associated with no illness and usually no irritation (Cameron and Burgess, 1945; Chin and T' Ant, 1946; Dangerfield, 1946; Domenjoz, 1944; Draize et aI., 1944; Fennah, 1945; Haag et aI., 1948; Wasicky and Unti, 1944). In fact, Hoffman and Lendle (1948) reported that even subcutaneous
injection of colloidal suspensions of DDT in saline in concentration up to 30 ppm caused no irritation. Zein-el-Dine (1946) reported that DDT-impregnated clothing caused a slight, transient dermatitis, but the method of impregnation was not stated and the absence of solvent was not guaranteed. Other more thorough studies of DDT-impregnated clothing have found it nonirritating (Cameron and Burgess, 1945; Domenjoz, 1944). Chin and T' Ant (1946) applied smallpads impregnated with different formulations of DDT to the inner surface of the forearm of 32 volunteers whose cutaneous sensation had previously been measured for a period of 5 weeks. Pads impregnated with all the elements of the formulation except DDT were applied to the corresponding position on the other arm as a control. Powdered DDT and 5% solutions of DDT showed little effect. Ten percent and 20% solutions in olive oil and petroleum showed no remarkable effect on sensation of pain, cold, or heat but reduced tactile sensation in most cases so that the minimal pressure that could arouse this sensation was 1-2.5 gmJcm 2 higher than in the control. 60.2.4.3 Experimental Respiratory Exposure Neal et al. (1944) reported almost continuous daily exposure to aerosols sufficient to leave a white deposit of DDT on the nasal vibrissae of the volunteers. This exposure produced moderate irritation of the nose, throat, and eyes. Except for this irritation during exposure, there were no symptoms, and laboratory tests and physical examination, including neurological evaluation, failed to reveal any significant changes. The studies by Fennah (1945) that involved both respiratory and oral exposure, produced no detectable ill effect, as discussed above. Stammers and Whitfield (1947) reported tests in which volunteers were exposed to DDT dispersed into the air either by volatilizing units or by continuously or intermittently operated aerosol dispensers. In some instances, a slight odor and some dryness of the throat were noticed, but otherwise the results were negative. 60.2.4.4 Therapeutic Use The early use of DDT for treating human body lice, head lice, and scabies was reviewed by Simmons (1959). Obviously, these uses offered a possibility of dermal absorption, but such absorption of dry DDT is very limited. Persons who had DDT blown into their clothing as they wore it must have inhaled some of the compound, and this was especially true of persons who used hand or power equipment to apply the dust to hundreds of people per day in mass delousing stations set up to control typhus (West and Campbell, 1946). However, the dosages absorbed cannot have been so large as in some instances in which DDT has been administered by mouth. Even smaller absorbed dosages for the general population were involved in the use of DDT for the control of other vector-borne diseases, especially malaria. These facts must not lead us to forget the tremendous contribution that DDT has made to human health through control of the vectors of typhus, malaria, plague, and several lesser diseases (Coulston, 1985; Hayes, 1991; Spindler, 1983).
60.2 DDT
DDT has been used on an experimental basis at oral dosage rates varying from 0.3 to 3 mg/kg/day for periods up to 7 months in an attempt to decrease serum bilirubin levels in selected patients with jaundice (Thompson et aI., 1969). No side effects were observed. No improvement was noted in patients with jaundice based on cirrhosis who had no demonstrated liver enzyme deficiency. However, in a patient with familial, nonhemolytic, unconjugated jaundice based on a deficiency of glucuronyltransferase, treatment with DDT rapidly reduced the plasma bilirubin level to the normal range and relieved the patient of nausea and malaise from which he had suffered intermittently. The liver function tests as well as other laboratory findings remained normal. The improvement was maintained during the 6 months when DDT was administered and had persisted for 7 additional months at the time the report was written. In this case, a dosage of 1.5 mg/kg/day produced a steady rise in plasma levels of p,p' -DDT from an initial level of 0.005 ppm to a maximum of 1.33 ppm at the end of treatment. At this time, the concentration in body fat was 203 ppm. Plasma levels fell slowly after dosing was stopped (Thompson et aI., 1969). The highest daily intake in this series was six times greater than the highest level administered in earlier studies of volunteers and about 7500 times greater than the DDT intake of the general population at that time and even more than would be expected these days. The highest value for p,p'-DDT in serum observed in the entire series was 1.330 ppm, compared to 0.996 ppm, the highest value reported by Laws et al. (1967) for formulationplant workers. A lesser induction of the microsomal enzymes has been observed in workers also (Kolmodin et aI., 1969; Poland et al., 1970). Rappolt (1970) used a single dose of 5000 mg of DDT to promote the metabolism of phenobarbital, of which his three patients had taken an overdose. The treatment appeared useful. Neither Rappolt (1970) nor Thompson et al. (1969) encountered any side effects ofDDT. However, in addition to whatever action it may have had in promoting the metabolism of phenobarbital, the DDT administered by Rappolt must have acted largely as a pharmaceutical antidote for the barbituate. The largest dose previously administered intentionally was 1500 mg, which caused moderate poisoning in a volunteer, who, of course, had received no barbiturate (Table 60.5). 60.2.4.5 Accidental and Intentional Poisoning The earliest symptom of poisoning by DDT is hyperesthesia of the mouth and lower part of the face (Hayes, 1982). This is followed by paresthesia of the same area and of the tongue and then by dizziness, an objective disturbance of equilibrium, paresthesia and tremor of the extremities, confusion, malaise, headache, fatigue, and delayed vomiting. The vomiting is probably of central origin and not due to local irritation. Convulsions occur only in severe poisoning. Onset may be as soon as 30 min after ingestion of a large dose or as late as 6 hr after smaller but still toxic doses. Recovery from mild poisoning usually is essentially complete in 24 hr, but recovery from severe poisoning requires several days. In two instances, there was some residual weakness and ataxia of the hands 5 weeks after ingestion.
1325
Involvement of the liver has been mentioned in only a small proportion of cases of accidental poisoning by DDT. In three men who ate pancakes made with DDT and who ingested 50006000 mg each, slight jaundice appeared after 4-5 days and lasted 3-4 days (Naevested, 1947). Hepatic involvement and convulsions were reported in an unsuccessful attempt at suicide by ingesting DDT and lindane (Eskenasy, 1972). Cases of individual and suicidal poisoning in which effects were clearly caused by DDT ingestion are summarized in Table 60.6. The signs and symptoms of poisoning were entirely consistent with those observed in volunteers, except that the spectrum of effects was broader because some of the accidental and suicidal doses were very high. A few persons apparently have been killed by uncomplicated DDT poisoning, but none of these cases was reported in detail. Death has been caused much more frequently by the ingestion of solutions of DDT, but in most of these instances the signs and symptoms were predominantly or exclusively those of poisoning by the solvent (Hayes, 1959). This does not mean that the toxicity of the solvent always predominates. For example, the recurrent convulsions in a case reported by Cunningham and Hill (1952), though more characteristic of poisoning by one of the cyclodienes, was certainly not typical of solvent poisoning. A 2-year-old child drank an unknown quantity of fly spray of which 5% was DDT, but the nature of the other active ingredients or the solvent was unknown. About 1 hr after consumption the child became unconscious and had a generalized, sustained convulsion. Convulsions were present when the child was hospitalized 2 hr after taking the poison, but the fits were controlled by barbiturates and other sedatives. Convulsions reoccurred on day 4 and again on day 21 but were stopped each time following renewal of treatment. On day 12, it was noted that the patient was deaf. Hearing began to improve about day 24 and was normal, as were other neurological and psychic findings, when the patient was seen about 2.5 months after the accident. Clinical effects of one toxicant may be modified by combining it with another. For example, one would not expect prolonged illness from DDT at a rate of 27 mg/kg. However, when DDT and lindane were ingested in a suicidal attempt at dosages thought to be 27 and 18 mg/kg respectively, clinical remission of convulsions and of liver involvement was delayed until day 20, and the EEG did not return to normal until day 39 (Eskenasy, 1972). What little is known about the effect of DDT on the human heart fails to show whether cardiac arrhythmia might be a possib1e cause of death in acute poisoning, as is true in some species of laboratory animals. Palpitations, tachycardia, and "irregular heart action" have been noted in some but not all cases of acute poisoning (Hsieh, 1954; Mackeras and West, 1946; Naevested, 1947). There do not seem to be any accidents or suicides involving respiratory or dermal exposure leading to recognized signs and symptoms of DDT poisoning. This is true even though sufficient respiratory exposure to aerosols or sufficient dermal exposure to solutions can cause poisoning in animals, and the difference is certainly one of dosage.
1326
CHAPTER 60 DDT and its Analogs
60.2.4.6 Use Experience
Despite its bad press and concern over its environmental impact the safety record for humans in the use of DDT is phenomenally good considering the huge quantitites distributed [Coulston, 1985; Food and Agriculture Organization/World Health Organization (FAOIWHO), 1985]. It has been used for mass delousing in such a way that the bodies and inner clothing of thousands of people of all ages and states of health were liberally dusted with the compound. By necessity, the applicators worked in a cloud of the material. Other applicators have sprayed the interior of hundreds of millions of homes in tropical and subtropical countries under conditions that Wolfe et al. (1959) showed involved extensive dermal and respiratory exposure. A smaller number of people have made or formulated DDT for many years. Extensive experience and numerous medical studies of groups of workers have been reviewed (Hayes, 1959). Dermatitis was commonly observed among workers who used DDT solutions. The rashes were clearly due to the solvent, especially kerosene. As often happens with rashes caused by petroleum distillates, they were most severe in people when they first started work and cleared in a few days unless contamination was exceptionally severe. A smaller number of workers experienced mild narcotic effects (vertigo and nausea) from solvents when working in confined spaces. Gil and Miron (1949) reported that some persons suffered temporary irritability, fatigue, and other ill-defined symptoms after exposure in the dusty atmosphere of a delousing station, but the relation of these atypical findings to DDT was not clear. With these exceptions due largely to solvents, no illnesses clearly attributable to the formulations, much less to DDT, were revealed by the early studies. Mild moderate poisoning by DDT itself may have occurred among a group of factory workers exposed to air concentrations of 5-4200 mg/m3 , but no measurements were made of DDT in blood, fat, or urine. The workers complained of parethesia of the extremities, headache, dizziness, and some other difficulties less clearly linked to DDT (Aleksieva et aI., 1959). Even higher concentrations in air have been associated with tremor of the tongue and hands as well as with numerous subjective findings (Burkatzkaya et aI., 1961). Ortlee (1958) carried out clinical and laboratory examinations of 40 workers, all of whom were exposed to a number of other pesticides. They had been employed at this work, with heavy exposure, for 0.4-6.5 years with slightly less exposure for as much as 8 years. Exposure was so intense that during working hours many of the men were coated with a heavy layer of concentrated DDT dust. By comparing their excretion of DDA with that of volunteers given known doses of DDT, it was possible to estimate that the average absorbed dosages of three groups of the workers with different degrees of occupational exposure were 14,30, and 42 mg/person/day. With the exception of the excretion of DDA and the occurrence of a few cases of minor irritation of the skin and eyes, no correlation was found between any abnormality and exposure to the insecticide. Since very large doses of DDT injure the nervous system and liver of
experimental animals, special attention was given to a complete neurological examination and to laboratory tests for liver function. Although a few abnormalities were revealed, none related to DDT was detected. Laws et at. (1967) studied 35 men employed from 11 to 19 years in a plant that had produced DDT continuously and exclusively since 1947 and, at the time of the study, produced 2722 metric tons per month. Findings from medical history, physical examinations, routine clinical laboratory tests, and chest x-ray films did not reveal any ill effects attributable to exposure to DDT. No case of cancer or blood dyscrasia was found among the 35 heavily exposed workers in a DDT factory, nor did the medical records of 63 men who had worked there for more than 5 years reveal these diseases. Two men were employed who had a history of successfully treated cancer before they came to work, but no employee had contracted cancer during the 19 years the plant had operated; during this period, the workforce varied from 110 to 135. A study ofliverfunction of the heavily exposed men is discussed near the end of this section. Measurement of storage offered direct evidence of the men's heavy exposure. The overall range of storage of the sum of isomers and metabolites of DDT in the men's fat was 38647 ppm, compared to an average of 8 ppm for the general population. Based on their storage of DDT in fat and excretion of DDA in urine, it was estimated that the average daily intake of DDT by the 20 men with high occupational exposure was 17.5-18 mg/person/day, compared to an average of 0.028 mg/person/day then found for the general population. There was significant correlation between the concentrations of total DDT-related material in the fat and serum of the workers. The concentration in fat averaged 338 times greater than that in serum, a factor about three times greater than that for people without occupational exposure. Compared to people in the general population, workers were found to store a smaller proportion of DDT-related material in the form of DDE; the difference was shown to be related chiefly to intensity rather than duration of exposure. DDE is relatively much less important and DDA much more important as excretory products in occupationally exposed men than in men of the general population. After Laws et at. (1967) had completed their study, it was found that the 36 most heavily exposed workers involved had fathered 58 children before they began working at the DDT factory and 93 children afterward (Wilcox, 1967). Laws et at. (1973) made a detailed study of the liver function of 31 men who had made and formulated DDT and who had been the subjects of the earlier study already discussed. Judging from their excretion and storage, the men's exposure was equivalent to an oral intake of DDT at rates ranging from 3.6 to 18 mg/man/day for periods ranging from 16 to 25 years and averaging 21 years. All tests were in the normal range for total protein, albumin, total bilirubin, thymol turbidity, and retention of sufobromophthalein sodium. One man had mild elevation of alkaline phosphatase and SGPT. Another man had elevated alkaline phosphatase concentration of 14 units, while a third man had an elevated SGPT.
60.2 DDT
Comparison of the residue levels of DDE, o,p'-DDT, and p,p'-DDT in blood of factory workers exposed to DDT formulations and showing apparent tempory clinical symptoms and those without symptoms showed no significant differences. Levels were approximately lO-fold greater than those of unexposed controls (Chand et aI., 1991). By far the largest number of heavily exposed workers whose health has been investigated are those associated with malaria control (WHO, 1973). In Brazil, periodic clinical examinations were made of 202 sprayers exposed to DDT for 6 or more years, 77 sprayers exposed for 13 years ending in 1959, and 406 controls. During a 3-year period, a survey of illnesses requiring medical care during the 6 months preceding each periodic medical examination failed to demonstrate any difference between exposed and control groups. A small number of analyses indicated that the concentration of DDT in the blood of sprayers was about three times higher than that of controls. In India, the blood levels of 144 sprayers were 7.5-15 times greater than those in controls and were at least as high as those reported for workers who had made and formulated DDT elsewhere (Misra et al., 1984). When the sprayers were examined, no differences from controls were found except that knee reflexes were brisker, slight tremor was more often present, and a times Romberg test was more poorly performed by the sprayers. The positive results led to the selection of 20 men for re-examination by a neurologist, who concluded that the differences found initially were not real or that the tests had returned to normal within the few months between the two examinations. In any event, the signs were not dosage related, since they showed no correlation with serum levels of DDT. Subsequently cognitive functions of Indian DDT sprayers were tested and DDT levels were 8.5 times higher than those in controls and visuomotor functions were significantly depressed. Perhaps in contrast to the above studies, the levels of DDT and its metabolites in the sera from 23 applicators in malaria control in Natal were significantly higher than in the population protected by the spraying (Bouwman et aI., 1991a). Although serum GGT was not statistically different from controls the mean in applicators was greater than the maximum laboratory mean level and ALT values were significantly greater in the applicators although not deemed clinically significant. Members of households which had been sprayed internally with DDT had significantly greater levels in their serum than people from nonsprayed households (sum of DDT and metabolites was 140.0 f.l.g/l compared to 6.4f.l.g/gl). Athough GGT levels were greater in the high DDT group this seemed to be associated with alcohol consumption (Bouwman et aI., 1991b). The induction by DDT level of microsomal enzymes of human liver was demonstrated first in workers (see Section 60.2.4.4) and DDT may be more important than DDE in this regard, as indicated by the fact that Poland et al. (1970) observed induction in men with average serum levels of 0.573 and 0.506 ppm for DDT and DDE, respectively, while Morgan and Roan (1974) found no induction in men with corresponding values of 0.052 and 0.222 ppm.
1327
As noted under Section 60.2.4.4 DDT has been used successfully to induce microsomal enzymes in order to promote metabolism of bilirubin in a case of congenital defect and to promote metabolism of phenobarbital in a case of overdose. DDT promotes its own metabolism in some species of laboratory animals. That the same is true in humans is indicated by the fact that storage of DDT is relatively less at higher dosages. However, the metabolism and subsequent excretion of DDT can be promoted even more by phenobarbital and especially diphenylhydontoin (see Section 60.2.3.5) and by some other drugs (McQueen et al., 1972). Establishment of a reduced equilibrium appeared to require about 2 months. Within this period, the regression of the level of DDT plus DDE on duration of treatment with diphenylhydantoin was highly significant. In addition to the studies already mentioned regarding workers with extensive storage and/or excretion of DDT as a result of truly heavy exposure to DDT, studies also have been made of a larger number of workers with lesser storage and/or excretion following lesser exposure to DDT but greater exposure to other insecticides. Continuing, meticulous study discussed by Hayes (1975) under Community Studies as well as the work of Tsutsui et al. (1974), Ouw and Shandar (1974), and Morgan and Lin (1978) failed to reveal effects of clinical significance among workers with prolonged, moderate exposure to a wide variety of pesticides. In a review of results for 2620 persons exposed to pesticides and lO49 persons not occupationally exposed, Morgan and Lin (1978) found that, apart from serum pesticide concentrations, the only significant and consistent charge associated with occupational exposure was a depression of serum bilirubin. This presumably was a reflection of a slight induction of liver microsomal enzymes. In addition, there was a tendency for serum alkaline phosphatase to increase with increasing concentrations of DDT plus DDE in the serum, but the differences were small in all instances and statistically significant for SGOT and LDH only. Wong et al. (1984) could find no significant overall cause specific mortality excess among men potentially exposed at work to DDT from 1935 to 1976. Similarly, a population of 499 persons living downstream from a defunct DDT-manufacturing plant showed no DDT-specific illnesses or ill health despite total DDT serum levels three times the national mean (Kreiss et al., 1981). There was, however, a possible association between serum DDT and serum cholesterol, triglyceride, and y-glutamyl transpeptidase levels. A positive linear correlation has been reported for the concentrations of vitamin A and of DDT-related compounds in the serum of men with at least 5 years of occupational exposure to DDT. However, the workers' DDT levels were little higher than those of persons in the general population (see Table 7.15 in Hayes, 1975, and Table 15.11 in Smith, 1991), and their vitamin A levels were within normal limits (Keil et aI., 1972b). Evidence regarding mutagenic activity of DDT and its significance in humans in uncertain. Comparing samples collected in winter and during the peak season of pesticide application, a slight increase in chromatid breaks was reported in the cultured lymphocytes of workers exposed to a wide variety of insecticides said to include DDT (Yoder et al., 1973). A somewhat
1328
CHAPTER 60
DDT and its Analogs
larger increase was reported for men exposed mainly to herbicides (You et aI., 1998). The paper failed to explain why exposure to DDT was claimed at a time when its use was banned. In another study lymphocytes cultured from workers with an average DDT plasma level of 0.999 ppm showed significantly more chromosomal and chromatid abberations than did cells cultured from controls with an average plasma level of 0.275 ppm (Rabello and Pereira, 1975). The difference was not significant in other comparisons in which the average plasma levels were 1.030 versus 0.380 ppm and 0.240 versus 0.030 ppm, respectively. Examination of all of the data presented by the authors suggests that a simple dosage-effect relationship was present, with a detectable effect starting somewhere between 0.2 and 0.4 ppm and increasing at levels higher than 0.4 ppm. Some chromosomal aberrations have also been observed with human lymphocyte cultures by Preston et al. (1981), but DDT did not cause unscheduled DNA synthesis in SV40-transformed human cells (Ahmed et aI., 1977). Although there is a lot of evidence against DDT causing liver cancer in humans in Western countries, there is still the outside possibility of it acting as a promoter of potent carcinogens. Aflatoxin is a well-known human carcinogen in areas of Southeast Asia such as Thailand, where DDT and other chlorinated insecticides have been used more recently. In Denmark, Unger and Olsen (1980) found significantly higher levels of DDE in adipose tissue from terminal cancer patients than in tissue from patients who died from other causes. In the United States, DDT and DDE levels were measured in 919 subjects in 1974 and 1975 (Austin et aI., 1989). After 10 years there was no correlation between these levels and overall mortality or cancer mortality except a slight correlation with respiratory cancer death. Of course, increased storage often correlates with emaciation of whatever cause (Hayes, 1975). A study has been made of 1043 deaths that occurred between 1956 and 1992 among men who used DDT in an antimalarial campaign in Sardinia in the late 1940s (Cocco et aI., 1997). Workers had a significant increased risk for liver and biliary tract cancers and mUltiple myeloma. However, nonexposed workers also showed elevated incidences of cancer. In the past decade there has been continued speculation and controversy as to whether many environmental chemicals, including chlorinated pesticides, can act as so-called "endocrine disrupters." As with all environmental exposures, this is an extremely difficult issue to develop to firm conclusions. In some studies, DDT was particularly targeted as being linked with a rising incidence of breast cancer (Wolff, 1995). Much of the evidence has been reviewed in detail by Ahlborg et al. (1995), but is not supportive or is inconclusive. Unger et al. (1984) found no relationship between breast fat tissue DDT (and DDE) and the incidence of mammary cancer. Although in the studies of Mussalo-Rauhamaa et al. (1990) correlation with ,B-hexachlorocyclohexane was found, none was observed with DDT. In contrast, elevated levels ofDDT or DDE were reported in cancerous breast tissue fat compared with tissue from benign mammary disease (Falck et aI., 1992; Guttes et aI., 1998). Studies on the relationship between blood levels of DDT and/or
DDE and breast cancer in the United States have been mixed. Some have shown a relationship with either blood p,p'-DDE levels and mammary cancer occurrence (Wolff et aI., 1993) or p,p'-DDE levels and hormone-responsive breast cancer (Oewailly et aI., 1994a, b). It has been proposed that increased estrogen receptor level in breast tumors over two decades could be explained by organochlorine exposure (Dewailly et aI., 1997). In contrast, a number of studies have found no relationship between blood levels of DDE and risk of breast cancer (Hunter et aI., 1997; Kreiger et aI., 1994; Schecter et aI., 1997). The last study is of additional interest in that women in North Vietnam were examined who had generally high levels of DOT or OOE due to exposure from antimalarial use. Many of the studies up to 1994 were examined by meta-analysis and appeared to confirm the lack of association between DDE levels in tissues and breast cancer incidence (Key and Reeves, 1994). However, there are many in vitro investigations which purport to agree with the hypotheses (e.g., Ardies and Dees, 1998; Shekhar et aI., 1997) and further investigations to clarify the issue are required. Probably there will be a further concern around p,p-DDE acting as an antiandrogen (Kelce et aI., 1995, 1998) (see Section 60.2.3.12). There have been reports that pancreatic cancer might be associated with exposure to DDT and ethylan. In a nested case-control mortality study among 5586 workers at a chemical plant, interviews with next of kin and co-workers and examination of work records showed that DDT exposure appeared to be associated with pancreatic cancer as identified by pathology, clinical surgical, autopsy records, or death certificates (Garabrant et aI., 1992). Among subjects whose mean exposure to DDT was 47 months, the risk was 7.4 times that among subjects with no exposure. DDD and ethylan exposures were also correlated. Malats et al. (1993) subsequently questioned the validity of the classification of pancreatic cancer. They proposed that despite the difficulties it might strengthen the findings if comparisons were made between histologically confirmed cases. Interestingly, this indeed turned out be true (Garabrant et aI., 1993). For those cases in which pancreatic cancer classification was based only on death certificates no association with exposure to DOT or its analogues was found. This theme has been developed further by Fryzek et al. (1997) looking at 66 residents of Michigan (39-70 years of age) who had been diagnosed histologically as having pancreatic cancer. Exposure was assessed by self-reporting questionnaires. In this cohort from the general population, a significant increased risk for exposure to ethylan was observed. Cases were 10.7 times more likely to report exposure to ethylan compared to controls. Nonsignificantly increased odds-ratios were for exposure to OOT and chloropropylate. Epidemiological studies of this sort with small case numbers are very difficult to perform and interpret. However, given the importance of breast and prostate cancer incidences in the United States and Europe it would seem wise to keep an open mind that DDT analogs might play a role in certain circumstances. Presumably where possible, future studies will address this issue.
60.2 DDT
1329
Table 60.6 Summary of the Effects of the Accidental or Suicidal Ingestion of DDT Individual dose (mg), formulation, and number of persons
Results and reference
300--4500, in food, I man
onset in I hr; vomiting; restlessness; headache; heart weak and slow; recovery
Unknown dose, in tarts, 25 men
onset in 2-2.5 hr; all weak and giddy; 4 vomited; 2 hospitalized; I confused, incordinated, weak;
5000-6000, in pancakes, 3 men
onset 2-3 hr; throbbing headache; dizziness; incoordination; paresthesias of extremities;
next day (Muhlens, 1946) one with palpitations and numbness of hands; recovery in 24-48 hr (Mackeras and West, 1946) urge to defecate; wide nonreacting pupils; reduced vision; dysarthria; facial weakness; tremor; ataxic gait; reduced sensitivity to touch; reduced reflexes; positive Romberg; slightly low blood pressure and persistent irregular heart action; partial recovery in 2-3 days, but slight jaundice appeared 4-5 days after ingestion and lasted 3--4 days; all normal 19 days after poisoning except irregular heart action in one (Naevested, 1947) 2000, in pancakes, 2 men
no illness (Naevested, 1947)
Up to 20,000, in bread, 28 men
onset in 30-60 min in those most severely affected, men first seen 2-3 after ingestion; in spite of severe early vomiting that reduced the effective dose, severity of illness and especially intensity of numbness and paralysis of extremities proportional to amount of DDT ingested; all but 8 men recovered in 48 hr; 5 others fully recovered in 2 weeks, but 3 men still had some weakness and ataxia of their hands 5 weeks after ingestion (Garrett, 1947)
Unknown dose, in flour, about 100 women
onset about 3.5 hr after ingestion; total of about 85 cases of which 37 were hopitalized; symptoms mild and similar to those in earlier outbreaks except gastrointestinal disturbance in most severe cases included abdominal pain and diarrhoea as well as nausea; most fully recovered in 24 hr (Jude and Girard, 1949)
Unknown dose, 14 cases
symptoms in established cases similar to those reported earlier (Francone et aI., 1952) with the exception of one man who was already sick when he received a dosage of 6 mg/kg, poisoning did not occur at dosages of 5.1-10.3 mg/kg. Ingestion of 16.3-120.5 mg/kg produced excessive perspiration, nausea, vomiting, convulsions, headache, increased salivation, tremors, tachycardia, and cyanosis of the lips. Onset varied from 2 to 6 hr, depending on dosage. Recovery required as much as 2 days (Hsieh, 1954).
286-1716, in meatballs, 8 cases, 11 exposed Unknown dose, 1 case
death 13 hr after suicidal ingestion (Committee on Pesticides, 1951)
Unknown dose, 22 unrelated cases
22 separate cases, including IS attempted suicides; some complicated by solvents; 3 deaths (Committee on Pesticides, 1951)
60.2.4.7 Dosage Response
60.2.4.8 Storage in Fat
The clinical effects of different dosage levels of DDT in humans are summarized in Tables 7.24 in Hayes (1975) and in Tables 60.5 and 60.6, herein. The degree of storage determined by different dosage levels of DDT has been summarized in Fig. 7.4 in Levine (1991) and details regarding higher than normal dosage rates are given in Table 15.9 in Smith (1991). A clinically useful degree of induction of microsomal enzymes was obtained with a DDT dosage of 1.5 mg/kg/day for 6 months (see Section 60.2.4.4). As discussed in Section 60.2.4.6 workers who absorbed a dosage of about 0.25 mg/kg/day showed demonstrable but only slight induction (Poland et al., 1970). Workers with less exposure as indicated by lower serum levels of DDT showed no detectable induction.
The highest reported storage of DDT and related compounds remains that of a healthy worker whose fat contained DDT and DDE at concentrations of 648 and 483 ppm, respectively (Hayes et aI., 1956). Laws et al. (1967) reported considerably lower storage values among the most exposed persons in a DDT manufacturing plant (see Table 15.11 of Smith, 1991). An important point is that whereas almost all investigations of workers are said to have been carried out on "heavily exposed" populations (or words to that effect), some of the groups studied had absorbed little more DDT than is absorbed by the general population-especially the general population of some tropical countries. The first evidence that humans metabolize a part of the DDT they absorb to DDE was obtained from the analysis of fat from a worker (Mattson et aI., 1953). The accumulation of DDE
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CHAPTER 60 DDT and its Analogs
relative to total DDT-related compounds is best illustrated in humans. Of the total DDT stored in the fat of workers exposed to technical DDT (about 4% DDE) for 11-19 years, only 38% was in the form of DDE, and, of course, some of that DDE came from their diets including meat (Laws et aI., 1967). In India, where many people avoid meat but may consume milk, cheese, and eggs, 34-41 % of total DDT stored by people without special exposure was DDE (Dale et aI., 1965). In the United States, during a time when DDT residues in food were decreasing, the proportion of total DDT in the form of DDE increased from about 60% in 1955 to about 80% in 1970; during the same interval the concentration of total DDT in body fat decreased from about 15 ppm to less than 10 ppm as recorded in Table 7.10 in Hayes (1975). By 1980, DDE constituted 86.7% oftotal DDT in one population (Kreiss et aI., 1981). Thus, a low proportion of DDE indicates a relatively high intake of preformed DDT and relatively few years for metabolism of stored DDT to DDE. A number of factors, especially dosage, age, sex, race, and various disease states have been discussed in connection with the storage and excretion of DDT by people but only dosage has been shown to be of practical importance. DDT and related compounds are stored at much lower rates in the general population than in persons with occupational exposure. However, these relatively low levels of storage constitute one of the most important aspects of the measurable effects of pesticides on people (Dale and Quinby, 1963). Briefly, storage of total DDT in the body fat of ordinary people in the United States increased from 5.3 ppm in 1950 to about 15.6 ppm in 1955 and 1956 (see Table 15.12, Smith, 1991). Thereafter, the levels decreased gradually (Bums, 1974), albeit somewhat irregularly, to about 8 ppm in 1970 and 3 ppm in 1980. In fact, despite concerns about residues of DDT, levels have continued to fall. In annual surveys in the United States based on 898-1920 samples per year, the geometric mean levels for total DDT in adipose tissue on a lipid basis were 7.88, 7.95, 6.88, 5.89, and 5.02 ppm for fiscal years 1970, 1971, 1972, 1973, and 1974, respectively. For each year, the values were higher for older age groups and higher for black than for white people. During fiscal year 1974, the values for persons 0-14, 15-44, and 45 years old or more were 2.15, 4.91, and 6.55 ppm, respectively, for white people and 4.02, 9.18, and 11.91 ppm, respectively, for black people (Kutz et aI., 1977). The values would have been somewhat lower if they had been based on wet weight. It has been calculated that if exposure to DDT ceased it would take 10-20 years for DDT to disappear from a person but that DDE would persist throughout the life span (Morgan and Roan, 1977). 60.2.4.9 Storage in Blood and Other Organs No information seems to be available on blood levels ofDDT in persons poisoned by the compound. Concentrations measured in the blood or serum of workers in the past are shown in Table 15.11 of Smith (1991). The highest value for total DDT in serum reported from several countries was 2.2 ppm (with an average of 0.7371 ppm) based on gas chromatography (Laws
et aI., 1967). A different situation is indicated by a report by Genina et al. (1969), who used a total chloride method to analyze samples of blood from controls and from persons with occupational exposure to DDT, polychloropinene, and BHC. These authors reported chloroorganic compounds as high as 38.4 ppm in the blood of warehousemen. This concentration is about 20 times the highest value found by the same authors in their control group. The factor of 20 is not remarkable, but (especially in view of the fact that polychloropinene and BHC are excreted more readily than DDT and DDE) values as high as 9 ppm in the controls are completely unexpected. Whether the difference was based on massive exposure or the crude analytical methodology employed is unclear. The concentrations of DDT in the blood of ordinary people are shown in Table 15.13 in Smith (1991). It is of interest that although each person without special exposure to DDT has relatively constant serum levels of DDT and DDE, DDE values differ more than the DDT values from person to person (Apple et aI., 1970). Whether this reflects differences in metabolism or differences in past exposure is unclear. Kreiss et al. (1981) have shown that DDE in serum samples of a community exceptionally exposed to DDT increased with age of the individual. Levels of DDT and its metabolites in the serum from those aged 21 years rose over a 12 month period following application of the pesticide to their homes in KwaZulu, South Africa. In contrast, levels fell in the age group 3-20 years, showing the complexity of any phamacokinetic interpretations (Bouwman et aI., 1994). Surveys have demonstrated a gradual decline in the concentrations of DDT and related compounds in human fat. Presumably a similar decline has occurred in the levels of these compounds in human serum. Consumption of fish appeared to be a predictor of plasma DDE levels but most reliable were age and serum cholesterol (Laden et aI., 1999). When storage of DDT has been found to be greater in black people, the difference could be accounted for by greater exposure (D'Ercole et aI., 1976; Hayes, 1975). However, Sandifer (1974), who found that the concentrations of DDT in the sera of blacks was two to three times greater than those in whites, also found a significant correlation between total DDT and deficiency of glucose-6-phosphate dehydrogenase, a condition much more common in blacks than whites. Thus, a genetic factor in the storage of DDT appears possible, but much stronger evidence would be necessary to confirm it. Whether the high storage in blacks is strictly environmental or partly genetic, it is certain that as high or higher levels have been recorded among several groups of rural blacks in different parts of the southeastern United States (Arthur, 1976; D'Ercole et al., 1976; Keil et aI., 1972a, b, 1973) than were reported by Kreiss et al. (1981) among blacks in Triana, Alabama, who had mean values of 0.096 and 0.062 ppm for total DDT in the serum of males and females, respectively. Storage of DDT and related compounds in the organs of adults and fetuses in the general population was discussed and tabulated by Hayes (1975). Concentrations in the viscera of adults averaged 1.0 ppm, but concentrations in lymph nodes and
60.2 DDT
especially bone marrow (a fatty tissue) approached the level in adipose tissue (::::6.0 ppm). Concentrations in some viscera of stillborn infants were similar to those in adipose tissue of the same infants and also in adults, suggesting that there had been a mobilization of DDT from fat prior to death. Saxena et al. (1987) have reported that the levels of DDT in human leiomyomatous uterine tissue were significantly higher than those in normal tissue (means of 0.845 and 0.103 ppm, respectively). Whether this is related to any estrogenic actions of DDT is unknown. See also Section 60.2.4.6. 60.2.4.10 Secretion in Milk
No information is available on the secretion ofDDT in the milk of women who were occupationally exposed to the compound or who were made ill by it, regardless of circumstances. The concentrations of DDT in the milk of women in various general populations are shown in Table 15.14 of Smith (1991). Values reported from Guatemala and early values from the USSR were much higher than those from other countries, and yet there was no indication of illness among babies fed such milk. The significance of DDT in milk and the dosages that different concentrations of it determine were discussed by Hayes (1975), Jenson (1983), Spindler (1983), and Coulston (1985). Quinby et al. (1965a, b) noted that women apparently were in negative DDT balance during lactation, but no direct measurement of DDT intake of women participating in the study was made. Subsequently, the ingestion of DDT in food and the secretion of DDT in milk were measured in the same women, and the fact of negative balance was confirmed (Adamovic and Sokic, 1973; Adamovic et aI., 1978; Cocisiu et aI., 1976) and may be a significant factor in determining the lower levels of DDT found in women than men in the general population (Adamovic and Sokic, 1973). Jonsson et al. (1977) found significantly lower levels of DDT (mean of 0.008 ppm) and of DDE (mean of 0.035 ppm) than had been reported earlier for the milk of city dwellers. However, levels remained quite high (0.05-1.90 ppm) in some rural black people (Woodard et aI., 1976). Some evidence suggested that DDT levels are higher in milk from smokers than nonsmokers, although there may be an occupational explanation (Coulston, 1985). In areas of KwaZulu, human milk levels of DDT and metabolites were significantly higher in women whose houses had been treated with DDT to interrupt malaria transmision (Bouwman et aI., 1990a, b). Primiparous mothers had significantly more than mUltiparous mothers. Transfer from the mother's milk to the child's blood was clearly demonstrated (Bouwman et al., 1992). Overall, despite the presence of DDT in human milk and placenta, there seems little risk to neonates in many different populations. Most evidence has shown a continuing decline in DDT levels in humans since it was banned for use in many parts of the world [e.g., Stevens et al. (1993) have found a marked decline in levels since 1974]. Clearly, levels would have to be very high before any advice against breast feeding could be given.
1331
60.2.4.11 Excretion ofDDT-Related Compounds
Among workers whose DDT intake was estimated to be about 35 mg/day, Ortlee (1958) found that the concentration of DDA in urine ranged from 0.12 to 7.56 ppm and averaged 1.71 ppm. Among workers whose exposure was about half as great, Laws et al. (1967) found concentrations from 0.01 to 2.67 ppm with a mean of 0.97 ppm. Continuous sampling of a DDT-formu1ating plant worker's urine showed that excretion of DDA increased promptly when exposure began on each of five consecutive workdays but often continued after exposure, sometimes reached a peak about midnight, and then decreased rapidly. On day 6, when there was no occupational exposure to DDT, the excretion of DDA continued until a very low level was reached. The highest concentration of DDA reported in this study was 0.68 ppm (Wolfe and Armstrong, 1971). The urine of people in the general population contains not only DDA but also neutral compounds including p,p' -DDT and p,p' -DDE (Cueto and Biros, 1967). Men with heavy occupational exposure to DDT excreted much more DDA but showed only a statistically insignificant increase in excretion of DDT and DDE. The urinary excretion of DDT derived material is of such an order of magnitude that it may account for much of the excretion of absorbed DDT. The excretion of DDA by people with different kinds and degrees of exposure is presented in Table 15.15 of Smith (1991). DDT and DDE are also excreted in the bile and higher levels than normal were found in the bile of one pest-control operator (Paschal et aI., 1974). Further discussion of the storage and excretion of DDT can be found in Levine (1991). 60.2.4.12 Other Laboratory Findings
In the absence of occupational DDT poisoning, there has been no opportunity to explore (as has been done with the cyclodiene insecticides) the relationship between clinical and EEG findings. In fact, the only DDT workers studied in this regard were exposed also to benzene hexahydrochloride and benzilan, so the findings might have been related to one or more of the compounds or to their interaction. Electroencephalograms were obtained from 73 of these workers exposed for periods ranging from 7 months to 20 years (Israeli and Mayersdorf, 1973; Mayersdorf and Israeli, 1974). Just over 78% of the records were normal and 21.9% were abnormal. The most severe changes involved persons exposed to the three compounds for 1-2 years; less severe changes were seen with either shorter or longer exposure. The changes were not correlated with age; the range and mean of age for those judged abnormal were almost identical with these values for persons considered normal. Some of the records showed bitemporal sharp waves with shifting lateralization combined with low voltage theta activity. Other records showed spike complexes, paroxysmal discharges composed of slow and sharp waves most pronounced anteriorly, and low-voltage rhythmic spikes posteriorly. None of the persons examined showed any abnormal clinical neurological finding. The incidence of abnormal electroencephalograms in the general population is 9.0 or 9.2%, according to other investiga-
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CHAPTER 60 DDT and its Analogs
tors cited by Israeli and Mayersdorf. Czegledi-Janko and Avar (1970) considered that at that time nonspecific EEG abnormalities occurred in 10-20% of the general population, so there is some question of whether the results are meaningful. Clinical laboratory findings associated with DDT poisoning are not specific and it is difficult to diagnose that poisoning has occured from this agent rather than others.
were killed quickly by dermal applications at the rate of 400 mg/kg day; they were made severely ill but did not die when treated at the rate of 200 mg/kg/day for 90 days. In rats fed for 2 years, the lowest dietary level producing gross effects was 400 ppm and the lowest level fed (100 ppm, about 5 mg/kg/day) produced tissue damage. In the rat, pathology is indistiguishable from that caused by DDT (Lehman, 1951, 1952).
60.2.4.13 Treatment of Poisoning
60.3.2.1 Absorption, Distribution, Metabolism, and Excretion
No useful guidance regarding treatment has been gleaned from the very few cases of DDT poisoning that have occurred. Animal studies indicate that sedatives, ionic calcium, and glucose or another ready source of energy would be useful. On the basis of experience in treating people poisoned by different convulsive poisons, it seems likely that diazepam would be beneficial.
60.3 TDE 60.3.1 IDENTITY, PROPERTIES, AND USES TDE is 1, 1-dichloro-2,2-bis(4-chloro-phenyl)ethane (Table 60.1). The common name TDE(ISO) is an acronym for tetrachlorodiphenylethane. Except in France, it is a generally recognized name for the compound as a synthetic insecticide. For reasons that are obscure, the word DDD (an acronym for dichlorodiphenyldichloroethane) has been used more commonly for l,l-dichloro-2,2-bis (chlorophenyl)ethane when viewed as a metabolite of DDT or when used as a therapeutic drug, and this distinction has been retained here. Almost everything we know about the compound relevant to humans is associated with its use as a drug rather than its use as an insecticide. Nonproprietary names for the a,p' -isomer which is used as a drug include chlordithane (USSR) and mitotane (United States). A proprietary name for the insecticide has been Rhothane®. Code designations include D-3, ENT-4,225, ME1,700, and NSC-38,721 (for a,p' -isomer only). TDE has the empirical formula C14HlOCl4 and a molecular weight of 320.05. The pure material forms colorless crystals melting at 109-11 O°c. The technical material consisted mainly of the p,p' -isomer but also contained a substantial proportion of a,p' -isomer and lesser proportions of related compounds. The insecticidal properties of TDE were first described by Laiiger et al. (1944). The formulations have included the technical material; wettable powders, 5%; emulsion concentrates, 25%; and dusts, 5% and 10%. 60.3.2 TOXICITY TO LABORATORY ANIMALS The effects of TDE are similar to those of DDT, but TDE is much less toxic in the rat and in humans. Gaines (1969) found the oral LD 50 in both male and female rats to be greater than 4000 mg/kg, Lehman (1951, 1952) reported 3400 mg/kg as an oral LD 50 in rats and 1200 as a dermal value in rabbits. Rabbits
The metabolism of p,p' -DDD has been described earlier in this chapter. Regardless of dosage form, 75% or more of a,p' -DDD is absorbed from the gastrointestinal tract (Korpachev, 1972a). Following repeated doses, storage of a,p'-DDD reached its highest point in 10-20 days and then decreased somewhat in spite of continued intake. Elimination was rapid after treatment stopped but was detectable longest in the adrenals and adipose tissues (Korpachev, 1972b). The metabolism of a,p'-DDD in the rat has been investigated thoroughly by Reif and Sinsheimer (1975); their major results are summarized in Fig. 60.3 which also records the metabolites found in humans by Reif et al. (1974). More recent studies to explain the covalent binding of a,p'-DDD in lung and adrenals are also described Section 60.2. 60.3.2.2 Biochemical Effects The basis for the action of a,p'-DDD on the adrenal is not understood fully in connection with any species but it is clear that marked species differences exist (see Section 60.2). The mechanism that leads to prompt atrophy in the dog may be quite different from the mechanisms that limit the production or increase the breakdown of corticosteroids in species in which most or all of the adrenal cells stay alive. It is clear that a reduction of steroid production accompanies atrophy of the dog. Kupfer (1967) considered: (a) reduced steroid production in species other than the dog, including the possibility that such reduction is secondary to inhibition of glucose-6-phosphate dehydrogenase activity in the adrenals, and (b) blockage of steroid action by a steroid metabolite formed under the influence of DDD. However, the existence of these effects, much less their importance, remains obscure. Hart and Straw (1971a) showed that administration of a,p'DDD to dogs for only 2-48 hr completely blocked the normal increase in steroid production in response to ACTH in vitro but, paradoxically, produced a marked increase in the incorporation of labeled amino acids into protein of the slices. The same authors presented evidence that the site of action is the intramitochondrial conversion of cholesterol to pregnenolone (Hart and Straw, 1971b), specifically, ACT-activated conversion and not baseline steroid production (Hart and Straw, 1971d). A secondary site involves inhibition of intramitochondrial conversion of 11-deoxycortisol to cortisol (Hart and Straw, 1971d). Further evidence supporting the importance of the primary site was found by Komissarenko et al. (1972). a,p' -DDD inhibited ACTH-induced steroid production by 797% within 2 hr, and the active principle is either a,p' -DDD per se or
60.3 TDE
a derivative formed in the adrenal gland of the intact dog (Hart and Straw, 1971c). o,p'-DDD applied to liver slices in vitro is not effective in reducing ACTH-induced steroidogenesis in the slices. However, the compound did reduce the formation of corticosteroid from progesterone or deoxycorticosterone added to homogenates made from adrenal cortices from dogs, chickens, rats, and human fetuses. These results are consistent with the view that the action of o,p' -DDD is to block ll-flhydroxylation (Kravchenko, 1973). Furthermore, a concentration of 16 ppm produced this effect in a monolayer culture of human fetal adrenal cells (Komissarenko, 1971). Martz and Straw (1973) interpreted the decrease in adrenocortical heme and P-450 produced by o,p'-DDD in the dog as a suggestion that the compound is metabolized to a more active form, and this is supported by in vitro studies with isolated adrenal mitochondria (Martz and Straw, 1980; Pohland and Counsell, 1985). Whether these actions are related the covalent binding of 3-methylsulfonyl-p,p'-DDE in mouse adrenals by CYPIlB has yet to be investigated (see Sections 60.2 and 60.2.3.5). There is evidence for a peripheral action of o,p' -DDD on steroid transformation in humans also, although the site of action is different. This evidence was obtained by studying the excretion of metabolites of small injected doses of radioactive steroid before and during administration of the drug. It was concluded that 3f1-hydroxy-~5-steroid dehydrogenase was inhibited (Bradlow et aI., 1963). Further evidence that o,p' -DDD has some inhibitory effect on the synthesis of corticosteroids in humans was provided by in vitro tests on adrenal tissue removed surgically from patients, some of whom had been under treatment with the drug (Touito et aI., 1978). Total doses prior to surgery had varied from 324 to 2280 gm and had been given over periods of 1-12 months. Compounds whose synthesis (from radioactive precursors added to incubation flasks) was inhibited in tissue from treated patients were cortisol, corticosterone, 18-hydroxycorticosterone, and aldosterone. Direct addition of o,p'-DDD to human adrenal tissue in vitro was without effect on synthesis of corticosteroids. Following massive dosage (60 mg/kg, iv), all of the isomers of DDD inhibit ACTH-induced steroid production in the dog, but the inhibition reached 50% of control in only 27 min after dosing with the m,p'-isomer (Hart et aI., 1973). There was a marked temporal correlation between the percentage inhibition of ACTH-induced steroid production, the disruption of normal cellular structure of the innermost zones of the adrenal cortex, and the severity of the damage to mitochondria in these zones caused by the three isomers. The effectiveness of m,p'DDT for treating metastatic adrenocortical carcinoma had already been demonstrated (Nichols et aI., 1961). However, in humans and dogs m,p' -DDD is less effective than o,p' -DDD (deFossey et aI., 1968; Reznikov, 1973). Administration of o,p' -DDD to dogs is followed by a decrease in plasma albumin and an increase in globulins, especially
1333
Guinea pigs receiving o,p' -DDD intraperitoneally at a rate of 100, 200, or 300 mg/kg/day for 20 days showed decreases in ascorbic acid levels corresponding to dosage (Petrun' and Nikulina, 1970). It was speculated that this might interfere with synthesis of corticosteroids. Like other chlorinated hydrocarbon insecticides, o,p' -DDD stimulates hepatic microsomal oxygenation of both drugs and steroids and this may explain much of its action on corticoid metabolism in a wide range of species (Kupfer, 1967). Increased breakdown is evidenced by increased excretion of polar metabolites while nonpolar metabolites remain stable or even decrease-fainding encountered in human patients (Hellmann et aI., 1973). However, the demonstrated effect on corticoid metabolism fails to explain why o,p' - and m,p' -DDD are unique in their overall effects on the adrenal, including their ability to produce adrenocortical atrophy in the dog. Other powerful inducers of microsomal enzymes lack these effects. Furthermore, in some systems DDD is a relatively weak inducer compared, for example, to DDT and DDE (Gillett et aI., 1966). Whereas induction does occur in dogs, its interpretation is complex; for example, the induction caused by repeated doses can be suppressed by cortisol (Martz and Straw, 1972). Mikosha (1985) has proposed that inhibition of NADP reduction by malic enzyme in adrenals may play a role in o,p' -DDD action, perhaps by causing a decrease in steroid metabolism (Ojima et aI., 1985). 60.3.2.3 Effects on Organs and Tissues DDD is used to control different forms of adrenal overproduction of corticoids in humans. This therapy originally was based on the demonstration that DDD (Nelson and Woodard, 1948, 1949) and especially o,p'-DDD (Cueto and Brown, 1958; Komissarenko et aI., 1968) cause gross atrophy of the adrenals and degeneration of the cells of its inner cortex in dogs. This is true even though it was first reported (Komissarenko et aI., 1970; Nelson and Woodard, 1948, 1949; Zimmerman et aI., 1956) that DDD produces almost no detectable damage to the adrenals of a variety of species, including humans. In the dog, o,p' -DDT produces gross atrophy of the adrenals when administered at a dosage of only 4 mg/kg/day. The dosage of technical grade DDD required to produce the same effect is 50-200 mg/kg/day (Cueto and Brown, 1958). However, in spite of its exceptional susceptibility, there is a definite threshold below which the dog does not respond. About 15% of technical DDT was o,p' -isomer, much of which is gradually metabolized to o,p' -DDD. Yet dogs remained healthy and reproduced normally in a threegeneration study involving dosages of technical DDT as high as 10 mg/kg/day. DDD has been little used for Cushing's syndrome in dogs (Lubberink et aI., 1971), but it is effective at lower dosages than those used in humans, and side effects are less serious and less frequent (Schechter et al., 1973). It is an interesting fact that p,p'-DDE and the -OH analog of p,p'-DDD causes moderate hypertrophy of the dog adrenal and 2,2-bis(p-
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CHAPTER 60
DDT and its Analogs
choropheny l)ethane causes moderate hyperplasia (Larson et aI., 1955). The effect of DDD on thymolymphatic tissues is poorly understood. In one of the earliest studies of the compound, Lillie et al. (1947) reported that the spleen of all treated animals showed impressive siderosis. Much later Gawhary (1972) reported that, in rabbits, intramuscular injection of a commercial grade DDD (mainly p,p'-isomer) caused acute atrophy of the thymus and hypertrophy of the adrenal, although the m,p'isomer at a dosage of 100 mg/kg/day caused hypertrophy of the thymus and an increase in its choline acetylase activity. Decrease in the weight of the thymus and spleen as well as the adrenal glands of rats treated with o,p'-DDD was reported by Hamid et al. (1974). Furthermore, Cueto and Moran (1968) and Cueto (1970) showed that, at a dosage of 50 mg/kg/day for 14 days, o,p'DDD caused a gradually progressive hypotensive failure in dogs injected with epinephrine or norepinephrine, while leaving unchanged the cardioaccelerator and immediate pressor response of these drugs. The hypotensive failure was associated with weakening of the contractile force of the heart and with a reduction of plasma volume. The latter may have been caused by loss of fluid from the intravascular compartment and was not caused by release of histamine. The hypotensive state could be prevented to a significant degree by pretreatment with prednisolone. In mice, p,p'-TDE at a dietary level of 250 ppm moderately increased the incidence of lung tumors in both sexes (Tomatis et aI., 1974). DDD (p,p'-TDE ) is toxic to isolated rabbit Clara cells and human bronchial epithelial cells by what appears to be cytochrome P450 activation to the acyl chloride (Nichols et aI., 1995). The o,p'-isomer was protective in rats treated earlier with the established carcinogen dimethylbenz[a]anthracene (DMBA) (Kravt'sova et aI., 1971). Leydig cell tumors were reported in the testis of rats receiving o,p'-DDD at the rate of 0.6 mg/kg/day for 285348 days (Lacassagne, 1971). This report is inconsistent with other studies (Lehman, 1951, 1952), and this may indicate that a contaminant was involved or a strain variation in the response. In an NCI study (1978a) there was a possible effect of TDE in causing an increased incidence of follicular cell carcinoma or follicu1ar cell adennoma of the thyroid in male Osborne-Mendel rats but no effects in female B6C3F1 mice. TDE was found not to be mutagenic in Drosophila (Voge1, 1972). It was found mutagenic in two of three indicator organisms in host-medicated tests but not in direct tests, suggesting that a metabolite was the active agent. However, in the same series of studies, both DDT and DDA were negative (Buse1maier et aI., 1973). In addition to atrophy of the zona fasciculata and zona reticularis in the dog, o,p'-DDD changes the ultrastructure of most cell types of the anterior pituitary of that species. The most striking feature is an increase in corticotrophocytes such as is seen following adrenalectomy, and the increase in cells is presumably associated with increased production of ACTH.
The hypothalamus also is involved (Gordienko and Kozyritsii, 1970; Gordienko and Kozyritskii, 1973). In spite of their severe nature, the changes produced in the dog adrenal are at least partially reversible (Komissarenko et aI., 1972). Dosageresponse relationships of mitochondrial swelling and of some other details of pathology in the dog adrenal have been explored by Gordienko and Kozyritskii (1973) and by Powers et al. (1974), who also investigated regeneration of the gland. Hypertrophy of the thyroid in dogs receiving 25 mg/kg and its inhibition in those receiving 50 mg/kg had been reported (Gordienko et aI., 1972). 60.3.3 TOXICITY TO HUMANS 60.3.3.1 Therapeutic Use
Following the demonstration that DDD caused atrophy of a part of the adrenal cortex of dogs, the compound has been used in humans in the hope of controlling excessive cortical secretion or of reducing the size of adrenal tumours. The underlying condition may be hyperplasia or adrenocortical carcinoma. Early attempts using mixed isomers and/or dosages less than 100 mg/kg/day often were ineffective, although side effects might be produced (Sheehan et aI., 1953). The dosage of o,p'-DDD has varied from 7 to 285 mg/kg/day, but a dosage of approximately 40 or more often 100 mg/kg/day for many weeks has been necessary to produce any benefit in humans (Bergenstal et aI., 1960; Bledsoe et aI., 1964; Gallagher et aI., 1962; Gutierrez and Crooke, 1980; Komissarenko et aI., 1970; Southern et aI., 1966a, b; Verdon et aI., 1962; Wallace et aI., 1961). The effects of idiopathic hyperplasia may be controlled; in fact, a state of adrenal insufficiency may be produced (Canlorbe et aI., 1971; Helson et aI., 1971; Korthe-Schutz, 1977) or of adrenocortical activity secondary to a tumor that produces ACTH (Carey et aI., 1973). Very early attempts to use DDD for treating Cushing's syndrome often failed because the o,p'-isomer was not used and sometimes because the dosage was small but this was not true for what apparently was the first therapeutic use (Sheehan et aI., 1953). Using the o,p'isomer, a favorable response is produced in about one-fourth to one-half of patients with inoperable adrenocortical carcinoma (Canlorbe et aI., 1971; Gutierrez and Crooke, 1980; Hoffman and Mattox, 1972; Hutter and Kahoe, 1966; Lubitz et aI., 1973; Montgomery and Struck, 1973). In fact an occasional cure, involving complete regression of metastases, is produced by chemotherapy including o,p'-DDD (Hart et aI., 1973; Pellerin et aI., 1975; Perevodchikova et aI., 1972; Rappalport et aI., 1978; Shick et aI., 1973). Other patients have lived for several years (Bricaire and Luton, 1977; McKierman et aI., 1978). More commonly, symptoms are relieved and life is prolonged only about 7-8 months or a little longer (Can1orbe et aI., 1971; Hoffman and Mattox, 1972; Hutter and Kahoe, 1966; Lubitz et aI., 1973), or even less (Hajjar et aI., 1975). The success of treatment often is indicated early by a reduction of steroid excretion (Hoffman and Mattox, 1972; Lubitz et aI., 1973), but
60.3 TDE
steroid excretion may increase, decrease, or remain unchanged (Fukushima et aI., 1971). Removal of the tumor and o,p'-DDD treatment may be combined (Levy et aI., 1985). The success of treatment is greater in Cushing's syndrome due to adrenal hyperplasia (Weisenfeld and Goldner, 1962). An early example of what appeared to be complete cure was reported by Bar-Hay et al. (1964). Ten of 17 patients with this condition experienced cure or remission for 12-32 months after the drug had been withdrawn (Luton et aI., 1973). The large dosage of o,p'-DDD necessary to produce clinical benefit often produces general lassitude, anorexia, nausea, vomiting, diarrhea, and/or dermatitis (Bochner et aI., 1969; Danowski et aI., 1964; Halmi and Lascari, 1971; Hoffman and Mattox, 1972; Hutter and Kahoe, 1966; Naruse et aI., 1970; Southern et aI., 1961; Weisenfeld and Goldner, 1962). Gynecomastia, hematuria, leukopenia, and thrombocytopenia have been reported more rarely (Luton et aI., 1972; Perevodchikova et aI., 1972). The symptoms disappear soon after administration of the drug is stopped or the dosage is reduced. Furthermore, some patients do not develop toxicity. A lO-year-old girl received 7500 mg/day for a total of 9 kg without discernible side effects (Helson et aI., 1971). Even large, therapeutic doses of o,p'-DDD cause no histological alterations of the adrenals in humans (Wallace et aI., 1961). However, electron microscopy revealed degenerative changes in the mitochondria of the zona fasciculata of a patient who had received o,p'-DDD at the rate of about 3000 mg/day for 1 month (Temple et aI., 1969). Dosages in the therapeutic range (specifically those between 110 and 140 mg/kg/day) produced no detectable injury to the liver, kidney, or bone marrow even though the patients exhibited the reversible symptoms listed earlier (Bergenstal et aI., 1960). Kupfer (1967) reviewed the extensive literature indicating that the effect in humans and other species, except the dog, is caused by stimulation of corticoid metabolism by massive doses of o,p'-DDD and not to any direct reaffect on the adrenal. Southern et al. (1966a, b) agreed that the effect was predominantly extra-adrenal in humans when the drug was first given but offered evidence that adrenal secretion of cortisol eventually was reduced. Even though therapeutic doses eventually have a direct effect on the adrenal, doses encountered by workers exposed to technical DDT do not (Clifford and Weil, 1972; Morgan and Roan, 1973). Somewhat encouraging results were reported in the use of p,p'-DDD for treating diabetics with hyaline vasular changes and hyperpolysacharidemia (Tornblom, 1959). Apparently, there has been no attempt to use o,p' -DDD for this condition. o,p'-DDD has been used, in a much lower dosage, for treating spanomenorrhea associated with hypertrichosis (Klotz et al., 1971). Menstruation was restored in 13 of 15 women with these conditions, and normal pregnancies occurred in five of them during the treatment period. The babies were normal. There was some improvement in hypertrichosis in nine and no improvement in six. At least part of the action of o,p'-DDT in controlling excessive androgens involves its action on their metabolism. It
1335
was found in a study of three patients with metastatic adrenal carcinoma and one with penicious anemia that the compound decreased the conversion of labeled androgens to androsterone by about 76% and to etiocholanolone by about 80%. The main effect on androgen metabolism was consistent with induction of microsomal oxidase activity by the drug (Hellmann et aI., 1973). When uptake of radioactive iodine is used for diagnosis of Cushing's syndrome, [131 I]19-iodocholesterol is the compound usually employed. DDD labeled with 131 1 has been used for the same purpose (Skromme-Kadlubik, 1972, 1973a, b, 1974). No comparative study of the duration of storage of the two compounds appears to have been made. However, it is clear that it is possible to introduce enough radiation via 131 1labeled DDD either to kill rodents or to cause atrophy of their adrenal glands, depending on the schedule of administration (Skromme-Kadlubik et aI., 1974). This was viewed as an indication that 1311-labeled DDD might be useful for treating human adrenal carcinoma.
60.3.3.2 Analytical Findings Studies associated with what apparently was the first attempt to use p,p'-DDD in treating Cushing's syndrome established that the compound is concentrated in the adrenal gland. Eleven weeks after the last course of DDD, when the concentration in adipose tissue was less than half what it had been earlier, the concentration in an adrenal biopsy was 50 ppm, wet weight. On a lipid basis, the concentrations in fat and adrenal were almost identical (Sheehan et aI., 1953). A patient who had received o,p'-DDD at the rate of 4000 mg/day for 58 days had a blood level of 6 ppm and excreted 8.3 mg of free and 39.7 mg of conjugated DDA in a 24-hr urine sample (Sinsheimer et aI., 1972). There is evidence for two plasma pools of o,p'-DDD (Slooten et aI., 1982). Normal volunteers excreted increased concentrations of DDA within 24 hr of receiving p,p'-DDD at a rate of 5 mg/day and continued to excrete DDA at greater than predose levels for over 4 months after dosing was stopped after 81 days (Roan et aI., 1971). Treatment of appropriate cases with o,p'-DDT usually results in a decrease in urinary steroid excretion (Gutierrez and Crooke, 1980). An unusually detailed study of the individual compounds is that of Hartwig et al. (1968). In long-term administration of o,p'-DDD (2grn/day for 1-3 months) to patients with adrenal carcinoma or Cushing's syndrome, Ojima et al. (1984) found that plasma levels of pregnenolone, progesterone, cortisol, corticosterone, and some other C2l steroids were progressively decreased, as well as urinary excretion of 17-ketosteroids and 17-hydroxycorticosteroids. Touito et al. (1985), however, were unable to demonstrate any correlation between concentrations of o,p'-DDD in adrenals removed from patients preoperatively treated with the drug for Cushing's syndrome and inhibition of some steroid biosynthesis enzymes measured in vitro. There is a suggestion that o,p'-DDD suppresses ACTH-secreting cells in the pituitary
1336
CHAPTER 60 DDT and its Analogs
as well as depressing steroid hormone secretion (Takamatsu et aI., 1981). In some patients either p,p' -DDD (Tornblom, 1959) or o,p'DDD (Molnar et aI., 1961) caused an increase in plasma cholesterol, but the opposite also may occur (Danowski et aI., 1964). Oddly enough, such patients are refractory to the therapeutic effects of the drug (Molnar et aI., 1961). An interesting finding is that o,p'-DDD has a hypouricemic effect apparently by increasing the renal clearance of uric acid (Reach et aI., 1978; Zumoff, 1979). The induction of microsomal enzymes by o,p' -DDD has been identified following their isolation from the urine of treated patients (Reif et aI., 1974). Persons suffering from Cushing's syndrome can be distinguished from normal by their increased adrenal uptake of [131]19-iodocholesterol. In one patient under treatment with o,p'-DDD, uptake of radioactive iodine was reduced but not to the normal range (Morita et aI., 1972). DDD is commonly found in blood and tissues from the general population. For levels of it in blood; see Table 15.13 in Smith (1991).
60.4 ETHYLAN 60.4.1 IDENTITY, PROPERTIES, AND USES Ethylan is 1,1-dichloro-2,2-bis(4-ethylphenyl)ethane, that is, the p,p'-diethy1 analog of DDD (Table 60.1). Ethylan is the only common name in use, but apparently it has been approved officially only in the USSR. The trade name Perthane® often is used. Code designations for ethylan include B-63,138 and Q-137. The CAS registry number is 72-56-0. Ethylan has the empirical formula CIsH20Cl2 and a molecular weight of 307.27. The pure compound is a crystalline solid with a melting point of 56-57°C. The technical product is a waxy solid with a melting point not below 40°C and with some decomposition above 52°C. The insecticide is practically insoluble in water but soluble in acetone, kerosene, and other organic solvents. Ethylan was introduced in 1950 and has been used to control pear psylla, leaf hoppers, various larvae on vegetable, and moths and carpet beetle on textiles. 60.4.2 TOXICITY TO LABORATORY
ANIMALS The oral LD 50 values for ethy lan were 8170 and 9340 mg/kg in rats and mice, respectively. However, the corresponding intravenous values were only 73 and 173 mg/kg in the same species. No dermal LD 50 value could be measured; all rabbits that received a 30% solution at the rate of 3 ml/kg/day for 13 weeks survived (Finnegan et al., 1955). Gaines (1960) agreed that the oral toxicity was very low (LD 50 > 4000 mg/kg). Minimal and infrequent changes were seen in the livers of rats fed dietary levels of 2500 and 5000 ppm for 2 years. There was no effect on survival, and differences in growth rate did
not correspond to dosage. Thus a dietary level of 1000 ppm might be considered a no-effect level. In contrast, the same investigators found that a dietary level of 5000 ppm was lethal to dogs within 22 weeks. Levels of 100 or 1000 ppm did not interfere with survival or growth when fed for 1 year, although the 1000 ppm level led to some atrophy of the adrenals (Finnegan et aI., 1955). Cortisone given at the same time as ethylan tended to block the effect of the latter on the adrenal (Bleiberg and Larson, 1957). Reznikov (1973) considered the action of p,p'ethylan similar to that of p,p' -DDD. 60.4.2.1 Absorption, Distribution, Metabolism,
and Excretion Rats fed ethylan at a concentration of 50 ppm (about 2.5 mg/kg/ day) for 6 weeks stored the compound in their fat at a concentration of 19 ppm (Finnegan et aI., 1955). Four generations of rats were fed a standard synthetic diet containing 20% fat to which several pesticides were added. The diets of the seven groups studied differed only in the kinds of fat (cotton seed oil, lard, etc.) they contained. The average concentration of added ethylan found by analysis in different samples of dietary fat varied from 2.01 to 2.71 ppm. No ethylan was detectable in the body fat or other tissues of the rats (Adams et aI., 1974). After a single dose of rt4C] ethylan, rats excreted 90% of the radioactivity in their feces and 5% in their urine in 2 weeks (Bleiberg and Larson, 1957). 60.4.2.2 Biochemical Effects Ethylan reduced excretion of 17-hydroxycorticosteroids and caused adrenal atrophy in dogs (Cobey et aI., 1956). Dogs that had received ethylan for 10 or 14 days at a rate of 200 mg/kg/day slept 12-14 hr following anesthesia with sodium pentobarbital compared to only 6-7 hr following the same dosage of barbiturate before receiving ethylan. Similar studies with different DDD formulations revealed that increased sleeping time did not depend on the presence of adrenal atrophy, but they did not exclude the possibility that it depended on altered function of the adrenal. The increased sleeping time did not depend on reduced clearance of the barbiturate from the blood, which was not influenced by ethylan or DDD. Thus the cause remained obscure. Whatever the cause, the increase in sleeping time was peculiar to dogs and did not occur in rats treated with DDD (Nichols et aI., 1958). In dogs, ethylan (50 mg/kg/day for 10 days) significantly increased the glutathione reductase of the adrenal cortex but not of the liver (Komissarenko et aI., 1968). 60.4.2.3 Effects on Organs and Tissues Ethylan produces adrenal cortical atrophy in the dog (Finnegan et aI., 1955; Larson et aI., 1955). No such effect was noted in the rat. Presumably the effect is virtually specific for the dog, as is true of o,p'-DDD, which has been more extensively studied. Dogs killed or rendered moribund by dietary level of 5000 ppm (about 105 mg/kg/day) showed marked atrophy of the adrenal cortex, the medulla was unaffected. The capsule was wrinkled,
60.5 Methoxychlor the zona glomerulosa contained cells with granular and diminished cytoplasm, and there was a focal loss of these cells. The zona fasciculata was greatly narrowed, and there was extreme vacuolization among cells in the medial two-thirds. However, fat-staining material was deficient throughout the fasiculata, especially the inner part. The zona reticularis had practically disappeared, leaving only a few cells containing lipochrome pigment. There were a few focal concentrations of lymphocytes. Atrophy was present but less severe in two of three dogs that received 1000 ppm (about 21 mg/kg/day). On the contrary, severe atrophy was produced in less than 3 weeks by an oral dosage of 200 mg/kg/day. When ethylan was administered to mice at the highest tolerated rate for about 18 months, the results for tumorigenicity were equivocal (Innes et aI., 1969). Some evidence for hepatic tumor formation in female mice but not males or rats of either sex has been reported (NCI, 1979).
60.4.3 TOXICITY TO HUMANS Ethylan was administered to nine men with metastatic carcinoma of the prostrate and to five women with metastatic carcinoma of the breast because there had been reports of a favorab1e effect of surgical adrenalectomy on the clinical course of some patients with these diseases and because the compound had been shown to cause adrenocortical atrophy in the dog (Taliaferro and Leone, 1957). All the patients also received ACTH either intermittently or continuously. With one exception, the dosage of ethylan varied from 50 to 150 mg/kglday, the latter for a total of 189,000 mg within 21 days. The most intensive treatment was 200-300 mg/kg/day for a total of 96,000 mg in 6 days. The smallest dosage produced diarrhea, vomiting, and especially nausea in some patients and required cessation of treatment. In contrast, other patients, especially those who were less sick to begin with, tolerated the higher dosages, including 200-300 mg/kg/day, with no symptoms whatever. Marked thrombocytopenia and 1eukopenia were noted in one patient just after a 14-day course of treatment. These changes, which were attributed to ethylan, resolved promptly when treatment was stopped. There was no other evidence of hematopoietic toxicity and no evidence of hepatic, renal, neural, or other toxicity. It was not considered that the relatively brief treatments influenced the clinical course in any of the cases. However, among patients receiving 150-300 mg/kg/day in divided doses, ethy1an caused a marked depression of plasma 17hydroxycorticosteroid levels, but never below the normal range. Lower dosages had no consistent effects. No distinct benefit but nausea, vomiting, and skin rash were seen in four other patients with carcinoma who received ethylan at dosages of 1800-8000 mg/day for periods of 4-54 days (Weisenfeld and Goldner, 1962). In surveys of workers and the general population who have developed pancreatic cancer association with ethylan exposure has been reported (Garabrant et aI., 1992, 1993). However, those studies involved self-reporting exposure. See also Section 60.2.4.
1337
60.5 METHOXYCHLOR 60.5.1 IDENTITY, PROPERTIES, AND USES Methoxychlor is 1,1, 1-trichloro-2,2-bis(4-methoxyphenyl) ethane or 2,2-bis(p-methoxyphenyl)-1, 1, 1-trichloroethane, that is, the p,p'-dimethoxy analog of p,p'-DDT. The structure is shown in Table 60.1. The common name, methoxychlor (BSI, ICPC, ISO), is in general use. Other nonproprietary names have included dianisy1 trichloroethane, dimethoxy-DT, DMDT (an acronym for dimethoxydiphenyltrichloroethane), and methoxy DDT. One trade name is Maralate®. Code designations include OMS-466. The CAS registry number is 72-43-5. Methoxychlor has the empirical formula C16HlSCl302 and a molecular weight of 345.65. The pure material forms colorless crystals melting at 89°C. Technical methoxychlor is a gray flaky powder containing about 88% of the p,p'-isomer, the remainder being mainly o,p' -isomer, although up to 50 other contaminants have been detected (Lamoureux and Feil, 1980; West et aI., 1982). The density is 1.41 at 25°C. Methoxychlor is stable to heat and ultraviolet light and resistant to oxidation; it is dehydrochlorinated by alkalies and by heavy metal catalysts. The solubilities of methoxychlor are approximately the same as those of DDT. It is readily soluble in most aromatic solvents, moderately soluble in alcohols and petroleum oils, and essentially insoluble in water. Methoxychlor was first described by Laiiger et al. (1944), and it was introduced about 1945. Formulations include wettable powder (25 and 50%), emulsifiable concentrate (24%), dusts (4-10%), and aerosols. Methoxychlor is effective against a wide range of insects affecting fruits, vegetables, forage crops, and livestock. The low toxicity of methoxychlor and its short biological half-life were largely responsible for its greatly expanded use following the ban on DDT in many countries.
60.5.2 TOXICITY TO LABORATORY ANIMALS 60.5.2.1 Basic Findings In spite of its low toxicity, methoxychlor in sufficient dosage is capable of causing convulsions in the dog (Tegeris et aI., 1966). However, in the rat, the compound causes depression of the central nervous system (Lehman, 1951, 1952). Tremors have been noted but they are not a prominent symptom. In rabbits killed by a few doses, the only signs noted were diarrhea and anorexia (Smith et aI., 1946). The acute toxicity of methoxychlor is very low; oral LD 50 values of 5000 mg/kg (Hodge et aI., 1950) and 6000 mg/kg (Lehman, 1951, 1952) have been reported for the rat and values of 1850 mg/kg (Domenjoz, 1946) for the mouse and 2000 mg/kg for the hamster (Cabral et aI., 1979a). Rats fed a dietary level of 30,000 ppm suffered a severe reduction in growth, and most of them died in less than 45 days. Those fed a dietary level of 10,000 ppm survived but gained almost no weight; paired feeding studies showed that failure of growth was due
1338
CHAPTER 60
DDT and its Analogs
entirely to food refusal. A dietary level of 1600 ppm for 2 years caused measurable reduction of growth but produced no reduction in life span and no histological change in the tissues. Dosages of 20 mg/kg/day for 1 year or a dietary level of 200 ppm (about 10 mg/kg/day) for 2 years were both no-effect levels (Hodge et aI., 1950, 1952). Rabbits are relatively susceptible to methoxychlor; a dosage of 200 mg/kg/day was fatal within 15 days (Smith et aI., 1946). Dogs fed the compound in their diet in such a way that they received 1000 mg/kg/day for 6 months lost weight, and many of those that received 2000, 2500, or 4000 mg/kg/day began having convulsions within 6 weeks and died within 3 additional weeks. Strangely enough, dogs that received 2500 mg/kg/day administered by gastric tube as a suspension in 1% gum tragacanth for 5 months showed no indication of injury although the amount of absorption was not clear (Tegeris et al., 1966). In an earlier study lasting 1 year, the highest dosage fed to dogs (300 mg/kg/day) lost weight for a month, but later those on the two lower dosages regained their original weight (Tegeris et aI., 1966). A dietary level of 2500 ppm for 8-16 weeks produced no ill effects in chickens (Lillie et aI., 1973). 60.5.2.2 Absorption, Distribution, Metabolism, and Excretion Mixtures containing methoxychlor appeared to pass directly from the luminal border of the rat jejunum to the intercellular spaces immediately, in addition to the usual transport through the endoplasmic reticulum toward the Golgi apparatus. Furthermore, absorption of methoxychlor was accompanied by distention of vesicles and intracellular spaces (Imai and Coulston, 1968). Storage of methoxychlor was minimal in the fat of rats that had received it for 18 weeks. No storage could be measured at a dietary level of 25 ppm, and storage decreased after the ninth week at levels of 100 and 500 ppm in spite of continued intake. Two weeks after dosing was discontinued no methoxychlor was detected (Kunze et aI., 1950). Storage of methoxychlor in sheep reaches a steady state in 6-8 weeks or, at certain dosage levels, actually declines after that interval in spite of continued intake perhaps due to induced metabolism. Storage loss is prompt after dosing is stopped (Reyno1ds et aI., 1975,1976). Following oral administration of radioactive methoxychlor to mice, 98.3% was recovered from the excreta within 24 hr. The compound was metabolized to 2-(p-hydroxyphenyl)-2'(methoxyphenyl)-l, 1, 1-trichloroethane and 2,2-bis(p-hydroxyphenyl)-l,l-trichloroethane, which were eliminated largely in conjugated form (Kapoor et aI., 1970). Detailed studies have also been performed with goats (Davison et aI., 1982, 1983). The normally low rate of storage of methoxychlor was not influenced by simultaneous feeding of DDT or dieldrin (Street and Blau, 1966). Rat liver microsomes form both the mono and dihydroxy products as well as more polar compounds (Bulger et aI., 1985; Ousterhaut et aI., 1981). With (R)and (S)-[monomethyl-2H3]methoxychlors, intramolecular deuterium effects and entantiotopic differentiation have been
observed and the demethylations appear due to CYP1B/2 and CYP2C6 isoforms (Ichinose and Kurihara, 1987; Kishimoto et aI., 1995; Kishimoto and Kurihara, 1996). Much of the metabolism work has been done by Kupfer and colleagues using microsomes, purified cytochrome P450 isoforms, or insect cells containing human cytochrome P450 species (Dehal and Kupfer, 1994; Kupfer and Bulger, 1987; Stresser and Kupfer, 1997, 1998a, b). In the rat, demethylation seems to occur with CYP2B isoforms which also actively hydroxylate in the ortho positions of methoxychlor and the mono and dimethoxy analogues. In humans, CYP2C19 and CYP1A2 seem to be responsible for demethylation. Ortho hydroxylation of the monodemethylated products is predominantly catalyzed by CYP3A4. On the whole bidemethylation and ortho hydroxylation seem to occur less readily with human samples than with rat liver. Another route of metabolism is an initial dechlorination before replacement of the methoxy groups (Davision et aI., 1983; Kupfer and Bulger, 1987) but this has not been greatly investigated. Phenolic products from both routes appear to be converted to activated intermediates, which may bind covalently to macromolecules (Bulger et aI., 1983; Kupfer et aI., 1986; Kupfer and Bulger, 1987). This might be related to the formation of the catechols formed by ortho hydroxylation (Bulger and Kupfer, 1989; Kupfer et aI., 1990). A scheme of methoxychlor metabolism is shown in Fig. 60.3. 60.5.2.3 Biochemical Effects Methoxychlor induced liver microsomal enzymes in sheep, but even at a dietary level of 2500 ppm the degree of induction was less than that caused by DDT at a dietary level of 250 ppm. Methoxychlor caused no change in food consumption, body weight, liver weight, uterine weight, or estrous cycle (Cecil et aI., 1975). In rats, methoxychlor induced hepatic CYP2B1 and CYP3A enzymes levels, but in regimes with multiple treatment and less efficiently than DDT (Li et aI., 1995). Some increases in hepatic glucose-6-phospatase activity have been observed in rats given methoxychlor without showing any histological damage (Morgan and Hickenbottom, 1979), and significant decreases in lactate levels occurred even at doses < 1% of the oral LD 50. Neonatal administration of methoxychlor to rats resulted in elevated levels of sex-monoamine oxidase activities in adult rats, implying changes in the brain hormone environment during development which did not become apparent until adulthood (Lamartiniere et aI., 1982). Whereas massive doses of methoxychlor have an estrogenic effect in swine (Tegeris et aI., 1966) and perhaps other species, no such effect is detectable in chickens at a dietary level of 10 ppm (Foster, 1973) or in sheep at a dietary level of 2500 ppm (Cecil et al., 1975). Even 10 ppm is a far greater residue than humans or livestock are likely to encounter. There was no evidence that methoxychlor induced microsomal enzymes in heifers when administered at the rate of 112 mg/kg/day for 9 days, as judged by recovery of radioactivity derived from 16a, 17a-rI 4 C]dihydroxyprogestrone acetophenide (DHPA) used to synchronize estrus (Rumsey and
60.5 Methoxychlor
/
1339
Possible oxidative route following initial monodechlorination
'\ Estrogenic metabolites
Potential covalent binding via semiquinone radical
CCI 3
HO,~ HoN Figure 60.3
/'
isomer
~ OH
Metabolism of methoxychlor. The exact mechanisms of covalent binding are not known.
Schreiber, 1969). However, because metabolites ofDHPA were not distinguished, the study cannot be considered a critical test of induction. Microsomal metabolism of methoxychlor in rat liver has been reported to result in binding to iodothyronine 5' -monodeiodinase at cysteine or lysine residues with resulting depression of iodinase activity in vivo. The significance of these findings on thyroid hormone metabolism and action is unknown (Zhou et aI., 1995).
60.5.2.4 Effects on Organs and Tissues Of several polycyclic aromatic compounds said to be impurities in commercial methoxychlor, only one was mutagenic and in only one strain (Grant et aI., 1976). Mixed mutagenicity results have been more recently reported for methoxychlor (Oberly et aI., 1993). Some tumors were found in rats fed for 2 years at dietary concentrations as high as 1600 ppm, but the kind and incidence did not differ from those in controls. No tumors were found in dogs that had received dietary levels up to 10,000 ppm (Hodge et aI., 1952). Negative results also were found in mice that received a single subcutaneous injection (10 mg/mouse) or in others given weekly skin applications (0.1 or 10 mg). However, it was concluded from identical study of other compounds regarded as carcinogens that the skin tests were not an adequate substitute for tests by other routes (Hodge et aI., 1966). When two strains of mice were fed technical methoxychlor at a dietary level of 750 ppm for as much as 2 years, the incidence
and malignancy of carcinoma of the testis were increased in one strain but not the other. It was suggested that the carcinogenicity was related to the estrogenic activity of methoxychlor (Reuber, 1979a). The occurrence of neoplasms of all sorts showed a very rough dosage response in male rats fed technical methoxychlor at dietary levels of 100 ppm or more and in females fed 10 ppm or more; tumors identified as carcinomas were found in both sexes but only in rats fed 2000 ppm for 2 years (Reuber, 1979b). The conclusion that methoxychlor is carcinogenic for the liver in C3H and BALB/c mice and Osbome-Mendel rats was proposed as being related to the covalent binding of activated metabolites (Kupfer and Bulger, 1987). However, anyone interested should consult the original reports, including NCI (l978b) which shows only poor evidence for carcinogenicity of methoxychlor. At a dietary level of 1000 ppm, the tumorigenic property of methoxychlor was less than additive when it was fed in combination with Aramite, DDT, and thiourea or Aramite, DDT, and aldrin (Deichmann et aI., 1967). Large doses of methoxychlor (1000,2000, or 4000 mg/kg/ day) produce dosage-dependent chronic nephritis and hypertrophy of the kidneys, mammary glands, and uteri of swine (Tegeris et aI., 1966). In rats, Hodge et al. (1950) could not find characteristic liver cell changes at any dosage level. They did find striking testicular atrophy at dietary levels of 10,000 ppm or greater and this was not present in parfed controls. Similar findings were reported by Tullner and Edgcomb (1962). No atrophy was present in a dog that had received methoxychlor
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at 100 mg/kg/day for a month. Under certain conditions, continued massive dosage of methoxychlor produced cystic tubular nephropathy in rats (Tu lIner and Edgcomb, 1962). Wistar rats given methoxychlor (100 or 200 mg/kg/day) for 760 days (males) or 14 days (females) showed inhibition of spermatogenesis and folliculogenesis (Bal, 1984). There were degenerative changes in Sertoli cells and in spermatogonia and spermatocytes, with some transformed to poly nucleate cells. Cytoplasmic vacuolations were observed in the epithelium of the ductus epididymis. Atresia of the ovarian follicles was evident with pyknosis and karyorrhexis of the granulosa cells. In rats, Lehman (1951, 1952) found that the lowest dietary level producing tissue damage was 550 ppm (about 25 mg/kg/day); the effects, which were confined to the liver, consisted mainly of a slight increase in incidence of hepatic cell adenomas. In rats and monkeys the early induction of liver enzymes and the concomitant increase in hepatic endoplasmic reticulum may be temporary, disappearing in animals treated with methoxychlor for a prolonged time (Serrone et aI., 1965). In dogs, a dosage of 2500 mg/kg/day caused grossly visible congestion of the intestinal mucosa. It has also caused progressive degeneration of the mitochondria of mucosal cells of the small intestine marked in the early stages by matrical swelling and disruption of the cristae and later by disappearance of cristae and appearance of small myelin bodies. The mitochondria of these cells showed some recovery in a dog that had been returned to uncontaminated food for only 3 weeks after 12 weeks of the high dosage of methoychlor (Tegeris et aI., 1968). For other effects on organs, see the two following sections on reproduction. 60.5.2.5 Effects on Reproduction
Large doses of methoxychlor have estrogenic effects (Tullner, 1961). Methoxychlor levels of 2500 and 5000 ppm reduced mating, and only one litter was produced. However, when the same rats were returned to an uncontaminated diet, they reproduced normally. A dietary level of 1000 ppm started before mating and continued throughout lactation had no effect on reproduction in that generation, but the female pups had early vaginal opening and reduced reproduction when mature, and the reproductive behavior of mature male pups was also defective (Harris et aI., 1974). The potency appeared to be 1110000 of diethylstilboestrol. In a more recent study, male and female weanling rats were dosed with methoxychlor at 100 mg or 200 mg/day through puberty and gestation until day 15 of lactation in females (Gray et aI., 1989). Although various parameters of reproductive potential were altered in both sexes, fertility was only reduced in females when mated with untreated animals. No implantations were observed in another study when female rats received the same doses of methoxychlor for 14 days before mating and during pregnancy (Bal, 1984). Preimplantation effects of methoxychlor seemed more important than postimplantation in ensuring success (Cummings and Gray, 1989). In vitro tests showed that pure methoxychlor itself is not estrogenic, although the commercial product had some activity
and bis(4-hydroxyphenyl)trichloroethane was quite active (Bulger and Kupfer, 1977; Bulger et aI., 1978a, b). The estrogenic activity of impure methoxychlor in inducing uterus growth, uterine orinithine decarboxylase and epidermal growth factor receptor, creatine kinase and peroxidase (Bulger et aI., 1978b; Cummings and Metcalf, 1994; 1995; Metcalf et aI., 1995) appears to be caused by the demethy1ated analogs (Bu1ger et aI., 1985; Ousterhaut et aI., 1981) which are also metabolites (see Section 60.5.2.2). By both in vitro and in vivo criteria, 1,1dichloro bis(4-hydroxypheny1)ethene is the most potent agent (Bulger et aI., 1985; Cummings, 1997; Kupfer and Bulger, 1987). The estrogenic effects of methoxychlor are not restricted to those on uterine physiology and function. Both running wheel activity (estrogen controlled) and sex behavior in rats and hamsters were induced by 400 mg/kg/day (Gray et aI., 1988). The actions of methoxychlor were not, however, completely identical to those of estradiol. Exposure of pregnant mice to methoxychlor has been reported to cause changes in behavior of male offspring (vom Saal et aI., 1995). Methoxychlor affects the dicidual cell response of the rat uterus (Cummings and Gray, 1987, 1989), a technique mimicing the growth and development of the endometrium during pregnancy, by a mechanism that apparently occurs by interaction directly with the uterus. Other mechanisms such as embryo transport rate might also be involved (Cummings and Perreault, 1990). Some evidence suggests that reproductive effects of methoxychlor metabolites in male rats (Bal, 1984; Tullner and Edgcomb, 1962) may be mediated, in part, by elevation of prolactin concentration and release, which in turn influences hypothalamic levels of gonadotropin-releasing hormone (Goldman et aI., 1986). In studies of the effect of methoxychlor on reproductive tract development following neonatal exposure of mice, precocious vaginal opening, cornification and increased tract size, and ovarian atrophy were observed in females and reduced serum testosterone, testicular DNA content, seminal vesicles, and prostrate in males (Cooke and Eroschenko, 1990; Eroschenko and Cooke, 1990; Eroshenko et aI., 1995). Changes in females were not however, completely identical to those observed with 17 ,B-estradiol (Eroshenko, 1991). On the other hand uterine luminal proteins were identical following technical methoxychlor and 17,B-estradiol administration to ovariectomised mice (Rourke et aI., 1991) as were morphometric parameters (Swartz et aI., 1994) although some toxicity was observed. When rats received methoxychlor intragastrically on days 6-15 of pregnancy, dosages of 50-400 mg/kg/day reduced maternal weight gain (Khera et aI., 1978). At 200 mg/kg/day, the compound decreased the number and weight of fetuses and caused delayed ossification leading to wavy ribs and other bent bones. No real teratogenesis was observed and no effects were observed in vitro or human and rat testicular cells when examined for single strand DNA breaks (Bjorge et aI., 1996; Khera et aI., 1978). Methoxychlor prevented ovariectomy-induced bone loss in the rat (Dodge et al., 1996). Chapin et al. (1997) have dosed rats before and following birth and looked for immune and re-
60.6 Chlorobenzilate
productive changes at doses of 0, 5, 50, or 150 mg/kg/day in a large study. Primary adult effects were reproductive and 5 mg/kg/day was not a NOAEL. A predictable result of the rapid metabolism and excretion of methoxychlor is the fact that very little of it is excreted in the milk and is thus of low risk to humans. When the compound was fed to cows at a dietary concentration of 7000 ppm, the concentration in the milk reached slightly over 2 ppm in 91 days and remained essentially constant until feeding was stopped on day 112 (Gannon et al., 1959). After dosing was stopped, the concentration in milk fell to <0.1 ppm in a week. In mice, methoxychlor fed to lactating dams did affect the reproductive tract of suckling females (Appel and Eroschenko, 1992). In chickens, dietary levels as high as 2500 ppm (about 145 mg/kg/day) had no effect on the health of hens or their production of hatchable eggs or on the ability of cockerels to fertilize the eggs. The hens were tested for 16 weeks and the cockerels for 8 weeks (Lillie et al., 1973). In summary, methoxychlor possesses many of the estrogenic properties of 17,B-estradiol probably after in vivo demethylation. Its fast metabolism and low potency (>1/10,000 of 17,B-estradiol) do pose questions as to any risk to humans and whether all its reproductive toxicity properties are identical to estrogen requires further study. Hall et al. (1997) suggested that in mice methoxychlor acts as an estrogen agonist in the uterus but as antagonist in the ovary. 60.5.3 TOXICITY TO HUMANS 60.5.3.1 Experimental Exposure Groups of volunteers were given methoxychlor at rates of 0, 0.5, 1, and 2 mg/kg/day for 8 weeks. Even the highest dosage was without detectable effect on health, clinical chemistry, or the morphology of blood, bone marrow, liver, small intestine, or testis (Stein et al., 1965; Stein, 1970). The highest dosage administered by Stein was similar to 1.4 mg/kg/day, which is considered safe for occupational intake, as reflected in the threshold limit value of 10 mg/m 3 . The low sensitizing property of methoxychlor has been noted (Szarmach and Poniecka, 1973).
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medical help 24 hr after exposure and was admitted to hospital about 36 hr after exposure. He was then dehydrated and suffering severe abdominal cramps and continuing diarrhea. On day 3 after exposure, jaundice was noted and continuous bilateral tinnitus began. On day 4, the patient was completely deaf and slightly dizzy. On day 5, rapidly progressive renal failure required hemodialysis. Soon afterward, peripheral sensory and motor neuropathy appeared, including hypoesthesias, paresthesias, persistent leg and foot pain, bilateral footdrop, and difficulty in moving the extremities. A generalized rash appeared. In spite of marked recovery, the patient still had profound, bilateral, sensorineural hearing loss, tinnitus, and moderate neuropathy of the legs and arms when he was reevaluated over 6 years later (Harell et al., 1978). Clearly, the delay in onset and the character of the illness were not consistent with poisoning by methoxychlor, malathion, or a combination of them, regardless of dosage. No thought seems to have been given to other possible causes, whether toxic or otherwise. A 49-year-old man who was exposed to a dust of methoxychlor and captan developed aplastic anemia a few weeks later and died within 6 months. He had also had light exposure to methoxychlor during the previous 2 years without symptoms (Ziem, 1982). 60.5.3.3 Laboratory Findings Most investigators have not found methoxychlor in human tissue. Apparently, the first exceptions were Griffith and Blanke (1975) who reported finding the compound infrequently and in unstated concentrations in blood taken at autopsy under the medical examiner system of Virginia. Under the circumstances of collection, the possibility of occupational exposure of the deceased could not be excluded. The reported persistence of methoxychlor for at least 7 days on the hands of a worker (Kazen et al., 1974) is interesting in view of the rapid metabolism of the compound once it is absorbed. However, it must be said that with modem analytical sensitivities, methoxychlor or metabolite residues might be found more frequently. 60.5.3.4 Treatment of Poisoning In the unlikely event that treatment is required, it must be symptomatic.
60.5.3.2 Use Experience There apparently has been no confirmed case of poisoning, occupational or otherwise, involving methoxychlor. However, atypical cases associated with methoxychlor have been reported. A 21-year-old man first noticed symptoms 8-9 hr after spraying several fruit trees with a formulation diluted from a mixture containing methoxychlor (15%) and malathion (7.5%) (Harell et al., 1978). The entire task took only 15-20 min, and afterward a shower was taken. The first symptoms were blurring of vision and gradual onset of nausea. Next morning, the man began vomiting and developed severe diarrhea. He sought
60.6 CHLOROBENZILATE 60.6.1 IDENTITY, PROPERTIES, AND USES The IUPAC name for chlorobenzilate (BSI, ISO, IMAF) is ethyl 4,4' -dichlorobenzilate. Other names are 4,4'-chlorobenzilic acid ethyl ester, ethyl 2-hydroxy-2,2-bis(3-chlorophenyl) acetate, and ethyl 4,4'-dichlorodiphenyl glycollate. For the structure see Table 60.l. Among many proprietary names have been G23992, Acaraben®, Benz O-chlor®, Benzilan®, and
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Kop-Mite®. The CAS registry number is 510-15-6. Ch1orobenzilate has the empirical formula C16H14Ch03 and a molecular weight of 325.20. It is a colorless solid melting at 37-37°C. It is very soluble in acetone and hexane but virtually insoluble in water. Impurities in the technical product, which is about 95% pure, can be dichlorobenzophenon, the ethyl ether of chlorobenzilate, and 4,4-dichlorobenzil. Chlorobenzilate was introduced as a technical product in 1952. It has been used mainly as a miticide on citrus crops or to control mites in beehives.
60.6.2 TOXICITY TO LABORATORY ANIMALS The acute oral LD 50 to mice, rats (Horn et aI., 1955), and hamsters (Cabral et aI., 1979b) is about 700 mg/kg. Symptoms in rats and mice include depressed motor activity and rapid wheezing respiration. Dogs tolerated daily oral doses of 64 mg/kg for 35 weeks and rats 500 ppm in the diet for 2 years (Horn et aI., 1955). After daily chlorobenzilate doses of 12.8 mg/kg to dogs, 5 days/week for 35 weeks, approximately 40% of the total dose was excreted unchanged or a urinary metabolites. No significant storage in fat of dogs or rats was reported (Horn et aI., 1955). Knowles and Ahmad (1971) described the conversion of chlorobenzilate by rat liver homogenates to p,p'-dichlorobenzilic acid, p,p'-dichlorbenzophenone, p,p'dichlorobenzyhydrol, and p-chlorobenzoic acid. In carcinogenicity studies chlorobenzilate induced hepatocellular carcinomas in mice, but the evidence in rats is uncertain (NCI, 1978c). Some testicular atrophy was observed in rats.
60.6.3 TOXICITY TO HUMANS A case of a pesticide sprayer poisoned by chlorobenzilate has been described (Ravindran, 1978). Symptoms included ataxia, delirium, fever, and muscle pains. Chlorobenzilate was detected in the urine of some workers employed in Florida citrus groves. Exposed workers had levels ranging from 0.07 to 6.2 mg/liter. It should be noted that the methodology employed involved oxidation to p,p'dichlorobenzophenone and would not distinguish between the parent chemical and some of its metabolites. (Levy et aI., 1981).
60.7 DICOFOL 60.7.1 IDENTITY, PROPERTIES, AND USES The IUPAC name is 2,2,2-trichloro-1, 1-bis (4-chlorophenyl)2,2,2-trichloroethanol, or 1, I-bis(p-chlorophenyl)-2,2,2-trichloroethanol, or 4,4-dichloro-a-trichloro-methylbenzhydrol. For the structure see Table 60.1 Dicofol (BSI, ISO) is also called Kelthane (JMAF). Proprietary names include Acarin®, Decofol®, Hifol®, Kelthane®, and Mitigan®. Dicofol has the empirical formula C14H9ClSO and a molecular weight of 370.50. The pure substance is colorless and melts at
78.5-70.5°C. It is soluble in most organic solvents but practically insoluble in water. Dicofol was introduced as a commercial chemical in 1955. Like chlorobenzilate, dicofol is used mainly as a miticide for citrus fruits, nuts, cotton, and beans. It still appears to be used in some countries. The technical product is a brown viscous oil with a d 2S of 1.45. The active compounds are 80% 1, 1-bis(4-ch1oropheny 1)-2,2,2-trich1oroethano1 and 20% 1-(2-chloropheny1)-1-(4-chloropenyl)-2,2,2-trichloroethanol (the o,p'-isomer). The other major impurity is 1,1,1,2tetrachloro-2,2-bis(4-chlorophenyl)ethane (Baum et aI., 1976). Dicofol can be produced as water-dispersab1e powders, as emulsions, and in a dust.
60.7.2 TOXICITY TO LABORATORY ANIMALS
In rats and rabbits the acute oral LD 50 for technical grade dicofol seems to range from 575 to 2000 mg/kg (Ben-Dyke et aI., 1970; Brown et aI., 1969; Smith et al., 1959). Dogs seem to be much less sensitive (Smith et aI., 1959). Rats fed dicofol for up to 2 years showed no effects on survival at levels below 1000 ppm but growth was impaired (Smith et aI., 1959). The maximum tolerable dose for mice in a subchronic study was 500 ppm (Sato et aI., 1987). Dogs fed 300 ppm showed no effect after 1 year, but some deaths occurred at 900 ppm. Dicofol seems to be metabolized in rats to 4,4' -dichlorobenzophenone, which is stored in fat and muscle as well as being excreted in the feces (Brown et aI., 1969). DDE was also found, but there is doubt as to whether this was due to metabolism of dicofol or to contamination of the technical product employed. Water-soluble metabolites have been detected in the urine of mice give radiolabelled dicofol. Nearly 50% of the administered doses was excreted in the urine within 24 hr. Part may be glucuronides of 4,4'-dichlorobenzhydro1 (Tabata et aI., 1979). Brown and Casida (1987) showed that in vivo mice convert dicofol to dichlorobenzophenone and dich1orobenzhydrol and that DDE originates from the impurity a-CI-DDT. There is little published work on the specific toxic effects of dicofol in experimental animals. Some small adverse effects associated with reproduction in rats and mice have been reported (Brown, 1972; Trifonova and Gladenko, 1980). In a comparative study, 98% dicofol, the technical product Kelthane, and DDT were given to male rats in equimolar amounts. Dicofol produced dosage-related increases in microsomal protein, cytochrome P-450, and the specific activities of cytochrome reductase, ethoxycoumarin o-deethylase, aminopyrine N-demethy1ase, and glutathione S-transferase at a potency equivalent to that of Kelthane, DDT, or phenobarbital (Narloch et aI., 1987). Some evidence has been obtained for its hepatocarcinogenicity in male B6C3F1 mice but not in rats (NCI, 1978d).
References
60.7.3 TOXICITY TO HUMANS Only one case of possible human poisoning by dicofol seems to have been reported, and this was in combination with trichlorfon (Zolotnikova and Somov, 1978). Greenhouse workers reportedly suffered frequently from allergic dermatitis. A detailed study of the protection of workers in Florida citrus groves from contamination by dicofol has been reported (Nigg et at., 1986). Dicofol has cytokinetic and cytogenetic effects on human lymphoid cells in vitro (Sobti et at., 1983).
REFERENCES Abou-Donia, M. B., and Menzel, D. B. (1968). The metabolism in vivo of 1,1, I-trichloro-2,2-bis(p-chloropheny I)ethane (DOT), I, I-dichloro-2,2bis(p-chlorophenyl)ethane (DOE) in the chick by embryonic injection and dietary ingestion. Biochem. Pharmacol. 17,2143-2146. Adamovic, V. M., and Sokic, B. (1973). Lower level phenomena of DOT cumulation in female abdominal fatty tissue. Arh. Hig. Rada Toksicol. 24, 303-306 (in Russian). Adamovic, V. M., Sokic, B., and Jonanovic-Similganski, M. (1978). Some observation concerning the ratio of the intake of organochlorine insecticides through food and amounts excreted in the milk of breast-feeding mothers. Bull. Environ. Contam. Toxicol. 20,280-285. Adams, M., Coon, F. B., and Poling, C. E. (1974). Insecticides in the tissues of four generations or rats fed different dietary fats containing a mixture of chlorinated hydrocarbon insecticides. J. Agric. Food Chem. 22,69-75. Agthe, c., Garcida, H., Shubik, P., Tomatis, L., and Wenyon, E. (1970). Study of the potential carcinogenicity of DOT in the Syrian golden hamster. Proc. Soc. Exp. BioI. Med. 134, 113-116. Ahlborg, U. K., Lipworth, L., Titus-Ernstoff, L., Hsieh, c.-C., Hanberg, A., Baron, J., Trichopoulos, D., and Adami, R.-O. (1995) Organochlorine compounds in relation to breast cancer, endometrial cancer and endometriosis: An assessment of the biological and epidemiological evidence. Crit. Rev. Toxicol. 25,463-531. Ahmed, F. E., Hart, R. W, and Lewis, N. J. (1977). Pesticide induced damage and its repair in cultured human cells. Mutat. Res. 42, 161-173. Alary, J. G., Guay, P., and Brodeur, J. (1971). Effect of phenobarbital on the metabolisms of DOT in the rat and in the bovine. Toxicol. Appl. Pharmacol. 18, 457-468. Aldridge, W N., Clothier, B., Forshaw, P., Johnson, M. K., Parker, V. H., Price, R. J., Skilleter, D. N., Verschoyle, R. D., and Stevens, C. (1978). The effect of DOT and pyrethroids cismethrin and decamethrin on the acetyl choline and cyclic nucleotide content of rat brain. Biochem. Pharmacol. 27, 17031706. Aleksieva, T., Vasilev, G., and Spasovski, M. (1959). Study of the toxic effects of DOT. J. Hyg., Epidemiol., Microbiol. Immunol. 5,8-15 (in Russian). Aim, H., Tiemann, U., and Torner, H. (1996). Influence of organochlorine pesticides on development of mouse embryos in vitro. Reprod. Toxicol. 10, 321-326. Ando, M. (1982). Dose-dependent excretion of DOE (I,I-dichloro-2,2-bis(pchlorophenyl)ethylene) in rats. Arch. Toxico!. 49, 139-147. Antunes-Madeira, M. C., Almeida, L. M., and Madeira, V. M. (1993). Depthdependent effects of DOT and lindane on the fluidity of native membranes and extracted lipids. Implications for mechanisms of toxicity. Bull. Environ. Contam. Toxicol. 51,787-794. Antunes-Madeira, M. c., and Madeira, V. M. (1993). Effects of DOE on the fluidity of model and native membranes: implications for the mechanisms of toxicity. Biochim. Biophys. Acta. 1149, 86-92. Appel, R. J., and Eroschenko, V. P. (1992). Passage of methoxychlor on milk and reproductive organs of nursing female mice. 1. Light and scanning elctron microscopic observations. Reprod. Toxicol. 6, 223. Apple G., Morgan, D. P., and Roan, C. C. (1970). Determiants of serum DOT and DOE concentrations. Bull. Environ. Contam. Toxicol. 5, 16-23.
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Ardies, C. M., and Dees, C. (1998). Xenoestrogens significantly enhance risk of breast cancer during growth and adolescence. Med. Hypotheses. 50,457464. Arthur, R. D. (1976). "The Prevalence of and Types of Pesticides in the Air of the Mississippi Delta and the Blood Serum of the General Population of Mississippi." Final Report (E-32) from the Department of Biochemistry, Mississippi State University to the Epidemiological Studies Program, Technical Services Division, U.S. Environ. Prot. Agency, Washington, DC. Askari, E. M., and Gablicks (1973). DOT and immunological responses. n. Altered histamine levels and anaphylactic shock in guinea pigs. Arch. Environ. Health 26, 309-312. Attaran, A., Roberts, D. R., Curtis, C. F., and Kilama, W L. (2000). Balancing risks on the backs of the poor. Nature Medicine 6, 729-731. Austin, H., Keil, E., and Cole, P. (1989). A prospective follow-up study of cancer mortality in relation to serum DOT. Am. J. Public Health 79, 43-46. Baker, M. T., and Van Dyke, R. A. (1984). Metabolism-dependent binding of the chlorinated insecticide DOT and its metabolite ODD to microsomal protein and lipid. Biochem. Pharmacol. 33, 255-260. Bal, H. S. (1984). Effect of methoxychlor on reproductive systems of the rat. Proc. Soc. Exp. BioI. Med. 176, 187-196. BaneIjee, B. D., Ray, A., and Pasha, S. T. (1996). A comparative evaluation of immunotoxicity of DOT and its metabolites in rats. Indian J. Exp. BioI. 34, 517-522. Bar-Hay, J., Benderly, A., and Rumney, G. (1964). Treatment of a case of non tumourous Cushing's syndrome with o,p'-DDD. Pediatrics 33, 239-244. Baum, H., Black, R. F., and Kurtz, C. P. (1976). Dicofol: Collaborative study of the hydrolysable chlorine method. 1. Assoc. Off. Anal. Chem. 59, 11091112. Ben-Dyke, R., Sanderson, D. M., and Noakes, D. N. (1970). Acute toxicity data for pesticides. World Rev. Pestic. Control 9, 119-127. Bergenstal, D. M., Hertz, R., Lipsett, M. B., and Moy, R. H. (1960). Chemotherapy of adrenocortical cancer with o,p'-DDD. Ann. Intern. Med. 53,672682. Bishara, R. H., Born, G. S., and Christian, J. E. (1972). Radiotracer distribution and excretion study of chlorophenothane in rats. J. Pharm. Sci. 61, 19121916. Bjorge, c., Brunborg, G., Wiger, R., Holme, J. A., Scholz, T., Dybing, E., and Soderlund, E. J. (1996). A comparative study of chemically induced DNA damage in isolated human and rat testicular cells. Reprod. Toxicol. 10, 509519. Bledsoe, T., Roland, D. P., Hey, R. L., and Liddle, G. W (1964). An effect of o,p' -ODD on extra-adrenal metabolism of cortisol in man. J. Clin. Endocrinol. Metab. 24, 1303-1311. Bleiberg, M. J., and Larson, P. S. (1957). Studies on the adrenocortical effects and metabolism of 2,2-bis-(p-ethyphenyl)-I,I-dichloroethane (Perthane). J. Pharmacol. Exp. Ther. 119, 133-134. Bochner, F., Lloyd, H. M., Roeser, H. P., and Thomas, M. J. (1969). Effects of o,p' -ODD and aminogluthehimide on metastatic adrenocortical carcinoma. Med. J. Aust. 1,809-812. Bouwman, H., Cooppan, R. M., Reinecke, A. J., and Becker, P. J. (l990a). Levels of DOT and metabolites in breast milk from Kwa-Zulu mothers after DOT application for malarial control. Bull. WHO, 68, 761-768. Bouwman, H., Reinecke, A. J., Cooppan, R. M., and Becker, P. J. (l990b). Factors affecting levels of DOT and metabolites in human breast milk from KwaZulu. J. Toxicol. Environ. Health. 13,93-115. Bouwman, H., Cooppan, R. M., Botha, M. J., and Becker, P. J. (199Ia). Serum levels of DOT and liver function of malaria control personnel. S. Air. Med. J. 79,326-329. Bouwman, J., Cooppan, R. M., Becker, P. J., and Ngxongo, S. (l991b). Malaria control and levels of DOT in serum of two populations in Kwazulu. 1. Toxicol. Environ. Health 33, 141-155. Bouwman, H., Becker, P. J., Cooppan, R. M., and Reinecke, A. J. (1992). Transfer of DOT used in malaria control to infants via breast milk. Bull. WHO 70,241-250. Bouwman, H., Becker, P. J., and Schute, C. H. J. (1994). Malaria control and longitudinal changes in levels of DOT and its metabolites in human serum from KwaZulu. Bull. WHO 72, 921-930.
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CHAPTER 60
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CHAPTER 60
DDT and its Analogs
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CHAPTER
61 Inorganic and Organometal Pesticides Thomas W. Clarkson University of Rochester
There are at least 18 elements that characterize one or more inorganic pesticides. Of these elements, 10 (chromium, copper, zinc, phosphorus, sulfur, tin, arsenic, selenium, fluorine, and chlorine) have been shown to be essential for normal growth. In these instances, the toxic effects clearly do not depend on the element per se but on the specific properties of one form of the element or one of its compounds, or merely on an inordinately high dosage. The other eight elements (barium, cadmium, mercury, thallium, lead, bismuth, antimony, and boron) have not been shown to be essential to growth of animals, although there is evidence that some may be. In any event, experience has shown that toxicity is not an argument against essentiality. Some highly toxic elements such as iron, selenium, arsenic, and fluorine certainly are essential to normal development. In the following sections, representative inorganic pesticides are arranged with reference to the periodic classification of the elements. In some instances this has involved sequential consideration of the members of a periodic group, such as the halides. In other instances, a series of elements, such as the heavy metals, have been considered in the order of their atomic numbers. This arrangement of the elements helps to explain the chemistry and toxicology of. their compounds. A final section deals with boron because the compounds involved are not important either as pesticides or as toxic ants for mammals, or for both reasons. The organometals and organometalloids are described along with the corresponding inorganic compounds. Such organic forms involve a stable covalent bond between the metal and carbon atoms, such as in methyl mercury compounds, and should be distinguished from compounds with ionic linkages to the organic moiety, such as in mercuric acetate, which are classified as inorganic compounds of the metal or metalloid. The organic species generally differ from the inorganic species in terms of kinetics, absorption, distribution, and excretion. The nature of their toxic effects may also differ markedly except for those organic species that are rapidly metabolized to the inorganic form in the body. Handbook of Pesticide Toxicology Volume 2. Agents
61.1 BARIUM Barium is an alkaline earth metal in the same group as magnesium, calcium, strontium, and radium. Its valence is 2. All water- and acid-soluble compounds of this element are poisonous. The intravenous LD 50 values of three of these compounds expressed as Ba2+ in two strains of mice ranged from 8.12 to 23.31 mg/kg, the toxicity being approximately the same as that of magnesium and greater than that of strontium (Syed and Hosain, 1972). 61.1.1 BARIUM CARBONATE 61.1.1.1 Identity, Properties, and Uses Chemical Name Structure
Barium carbonate.
BaC03.
Synonyms Barium carbonate occurs in nature as the mineral witherite. A code designation for commercial barium carbonate is C.L-77,099. The CAS registry no. is 513-77-9. Physical and Chemical Properties Barium carbonate has the empirical formula CBa03 and a molecular weight of 197.37. It is a tasteless, odorless, heavy white powder with a density of 4.2865. At about 1300°C it decomposes into BaO and C02. Its vapor pressure is negligible. Barium carbonate is almost insoluble in water. It is slightly soluble (I: 1000) in water saturated with carbon dioxide, soluble in dilute hydrochloric or nitric acid or in acetic acid, and soluble in solutions of ammonium chloride nitrate. Formulations and Uses Barium carbonate is a rat poison. It also is used in ceramics, paints, enamels, rubber, and certain plastics. The technical product is 98-99% pure. Rodenticidal baits contain 20-25% of the compound.
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CHAPTER 61
Inorganic and Organometa1 Pesticides
61.1.1.2 Toxicity to Laboratory Animals Basic Findings The oral LD 50 of a suspension of barium carbonate for wild Norway rats is 1480 mg/kg. The animals survive 1-8.5 days (Dieke and Richter, 1946). Strain differences in LD 50 values have been reported in mice (Syed and Hosain, 1972). Apparently no study has been made of the effects of repeated doses of barium carbonate. However, the much more soluble and somewhat more toxic barium chloride has been studied. Rabbits given subcutaneous doses at rates as high as 10 mg/kg per day showed no clinical effects. Rabbits given two daily doses at the rate of 10 mg/kg developed weakness, urination, defecation, difficult breathing, bradycardia, extrasystoles, and other signs lasting 2--4 hr following each dose (Fazekas, 1968). Absorption, Distribution, Metabolism, and Excretion Barium carbonate is highly insoluble in water. It is partially solubilized by acid in the stomach. The only real danger of the rodenticide is through ingestion. Various barium compounds can cause pneumoconiosis, but no form of poisoning by barium has been of much significance in industry. Information is lacking on the gastrointestinal absorption of barium carbonate. Barium given orally as barium chloride is absorbed about 10 times (63-84%) more in younger rats than in mature animals (7%) (Taylor et aI., 1962). Mature dogs also absorb about 7% (Cuddihy and Griffith, 1972). Several animal studies suggest that barium is incorporated into the bone matrix in a fashion similar to calcium (Bauer et aI., 1956; Bligh and Taylor, 1963; Dencker et aI., 1976; Taylor et aI., 1962). The uptake into bone decreases with age of the animal. Study of rats injected intraperitoneally with 140Ba indicated that excretion was most rapid during the first 4 hr and reached 7% in the urine and 20% in the feces in 24 hr (Bauer et aI., 1956).
Experiments in dogs showed that barium can be actively reabsorbed by the kidney tubules. Its clearance was correlated with calcium clearance. Protein binding of barium averaged 54% and was of the same order of magnitude as that of other alkaline earths. The data did not exclude a common transport mechanism with calcium (Rahill and Walser, 1965). Mode of Action Barium stimulates striated, cardiac, and smooth muscle, regardless of innervation. It is antagonistic to all muscle depressants, no matter whether they act primarily on nerve or muscle. Initial stimulation of contraction leads to vasoconstriction through direct action on arterial muscle, peristalsis through action on smooth muscle, tremors and cramps through action on the skeletal muscles, and various arrhythmias through action on the heart. If the dose is sufficient, stimulation is followed by weakness and eventually by paralysis of the different kinds of muscle. Some effects such as hypertension, violent tremors, and convulsions are uncommon following ingestion of barium carbonate. They are more likely to follow absorption of more soluble barium compounds. If death does
occur, it is caused by failure of muscular contraction leading to respiratory failure or cardiovascular collapse (Sollmann, 1957). Several studies on animals exposed to barium compounds by parenteral routes indicate that barium decreases serum potassium concentrations. These experiments support findings from a human case study that hypokalemia is an important effect of acute barium toxicity [Agency for Toxic Substances and Disease Registry (ATSDR), 1992a]. 61.1.1.3 Toxicity to Humans Experimental Exposure A solution of 140BaCb was injected intramuscularly into five children and intravenously into two adults, all with normal skeletal metabolism. Three of the children and one adult also received 45CaCb in the same injection. The pattern of excretion of calcium was similar, but both the initial and final rates were slower. Bone took up barium more rapidly than calcium under the same conditions, but, because of the more rapid excretion of barium, less of it was available to the bone. However, skeletal metabolism of the two elements was closely similar. Therapeutic Use
Barium carbonate has no therapeutic uses.
Accidental Poisoning Apparently the major accident involving barium carbonate was one in which 85 British soldiers were poisoned by eating pastry made from flour accidentally contaminated by the compound intended for use as rat poison. The clinical picture was basically the same in all patients. In spite of some individual variation, three poorly defined stages could be recognized: (l) an acute gastroenteritis with mild sensory disturbance, (2) loss of deep reflexes and the onset of muscle paralysis, and (3) progressive muscular paralysis. Often improvement in all affected systems began 3--4 hr after onset, and recovery was complete within 24-36 hr. Only a few of the most severely affected patients proceeded to the third stage. In these few, general muscle paralysis began on the second day of the illness and lasted for a further 24 hr. There was complete muscle paralysis of arms and legs, and in one case the paralysis affected the muscles of respiration, but function of the diaphragm was sufficiently spared that the patient survived. Even these dangerously ill patients remained mentally clear, and recovery was surprisingly rapid; by the fourth day all affected muscles had regained their full power. There were no deaths. Many of the patients with severe early diarrhea recovered much more rapidly than those with delayed bowel action (Morton, 1945). The clinical findings in other outbreaks were generally similar. Additional signs and symptoms included thirst, sweating, blurred vision, a desire to urinate, and a moderately increased blood pressure (Dean, 1950; Lewi and Bar-Khayim, 1964). Electrocardiographic changes considered characteristic of hypokalemia have also been reported (Diengott et aI., 1964). Chronic effects in humans have not been well studied. All epidemiological studies have suffered from numerous limitations
61.2 Chromium
61.2 CHROMIUM
(Brenniman and Levy, 1985; Brenniman et aI., 1979; Elwood et al., 1974; Schroeder and Kraemer, 1974). Dosage Response When barium carbonate was mistakenly used as X-ray contrast medium, six patients survived 133 g each (about 1900 mg/kg) but another died after only 53 g. The author cited earlier reports indicating that as little as 4 g (about 57 mg/kg) has proved fatal (Dean, 1950). However, such incidents must be rare, for in one large outbreak with no deaths among 85 cases, it was estimated, on the basis of analysis of the contaminated food, that the most severe cases received about 15 g of barium carbonate (about 214 mg/kg). This rate is less than the LD 50 for the rat but sufficient to kill some rats. It seems likely that vomiting is important in protecting people who ingest barium carbonate. Although the calculation must be viewed with caution, one might (by taking the ratio of the LD 50 values into account) conclude that a dosage of 28 mg/kg/day for barium carbonate is equivalent to what many years ago was a usual but often toxic dosage of 1.7 mg/kg/day for barium chloride. The permissible occupational intake of soluble barium salts (as Ba2+) is 0.07 mg/kg/day. Laboratory Findings Hypokalemia may occur in severe poisoning. Apparently there is no report of the concentration of barium in a case of poisoning. Following a single substantial dose, barium is deposited in bone (Bauer et aI., 1957). However, under practical conditions it does not act as a bone seeker. The concentration in normal human bone ash is only 7 ppm compared to average concentrations of 2.3-28 ppm in the ash of different organs (Snowden and Stitch, 1957; Snowden, 1958). Tipton and Cook (1963) reported that (except for skin, lung, and intestine, which may be environmentally exposed) the median value for barium in tissue ash did not exceed 7 ppm in the United States. Much higher concentrations of barium usually are found in Africa, the Near East, and the Far East (Tipton et al., 1965). On a wet weight basis, the concentration of barium in normal plasma does not exceed 0.44 ppm (Gofman et aI., 1964). Treatment of Poisoning Emptying of the stomach by vomiting or gastric lavage should be followed by sodium or magnesium sulfate (30 g). These compounds act not only as purgatives but as detoxifying antidotes because they precipitate the toxic barium ion as insoluble barium sulfate. On the basis of both experimental studies and clinical experience, Lydtin et al. (1965) recommended that magnesium sulfate and calcium chloride be administered intravenously as early as possible in barium intoxication. They found this treatment effective even when administered several hours after intake of the poison. The fluid and salt balance should be followed with particular attention to potassium and replacement therapy carried out as required. Be sure to have a respirator available. It may be lifesaving.
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Chromium is a metal somewhat like iron and separated from it in the periodic table only by manganese. Chromium has valences of2, 3, and 6, but only hexavalentchromium compounds (chromates) are important as pesticides. They are also the most toxic. Chromium in its trivalent state is an essential element. It is thought to be a necessary part of the glucose tolerance factor of the liver and required for certain other aspects of carbohydrate metabolism (Mertz, 1967). 61.2.1 SODIUM DICHROMATE 61.2.1.1 Identity, Properties, and Uses Chemical Name
Sodium dichromate.
Synonyms Other names for the compound are bichromate of soda and sodium bichromate. The CAS registry no. is 1058801-9. Physical and Chemical Properties Anhydrous sodium dichromate has the empirical formula Cr2Na207 and a molecular weight of 261.96. The dihydrate forms reddish to bright orange elongated prismatic crystals with a density of 2.348 at 25°C. The anhydrous salt melts at 356.7°C and begins to decompose at about 400°C. It is very soluble in water. Use Sodium dichromate is used as a defoliant of cotton and other plants and as a wood preservative. 61.2.1.2 Toxicity to Laboratory Animals Basic Findings The lethal intravenous dosages of sodium dichromate for several species are in the range 37-417 mg/kg. The toxicity of the potassium salt is of the same order of magnitude and its lethal oral dosage for the dog is 2829 mg/kg (Spector, 1955). Although one report (Schroeder, 1973) indicated that prolonged administration of 5 ppm hexavalent chromium in drinking water produced a slight decrease in growth, other studies indicate that concentrations up to 10,000 ppm may be given without ill effect (Gross and HelIer, 1946; MacKenzie et al., 1958). The National Toxicology Testing Program (NTP, 1996a, b) found no adverse effects in rats given potassium dichromate in feed for 9 weeks at a dosage rate of about 2 mg Cr/kg/day. An evaluation of animal toxicity data (ATSDR, 2000) indicates few adverse effects from chromium VI compounds have been detected at oral dosage rates below 1 mg Cr/kg/day even for long-term exposures. Absorption, Distribution, Metabolism, and Excretion Little information is available on the kinetics of disposition of
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chromium compounds in experimental animals. Chromate is absorbed by the lung (Baetjer et al., 1959), gastrointestinal tract (MacKenzie et al., 1959; Ogawa, 1976), and skin (Wahlberg, 1965). It can also be absorbed from skin injured by it, and when the burn covers as much as 10% of the body surface the absorption may be sufficient to cause fatal poisoning. Highest tissues levels are found in kidney and liver (MacKenzie et aI., 1958). In rats and hamsters fed chromium compounds, fecal excretion of chromium was in the range 97-99% of the administered dose. Urinary excretion of chromium varied from 0.6 to 1.4% of the dose whether given as chromium III or VI (Donaldson and Barrcras, 1966; Henderson et aI., 1979; Sayato et aI., 1980). Effects on Organs and Tissues Acute poisoning may produce death rapidly through shock or after several days through renal tubular damage and uremia. In most cases, the primary effect of acute exposure is kidney failure (Evan and Dail, 1974; Mathur et aI., 1977). Berry et al. (1978) noted that chromium was localized within the proximal renal tubules specifically within the lysosomes. It was retained throughout the study period of 8 months and was eliminated only when the entire cytoplasm underwent necrosis. Kirschbaum et al. (1981) have postulated that the initial effect is on specific elements of the microfilamentous system that is responsible for directing intracellular flow of reabsorbed solutes. The liver and other organs may be involved but generally to a lesser degree. No evidence of carcinogenicity was found in mice given potassium chromate (mg Cr/kg/day) in drinking water for three generations (Borneff et aI., 1968). 61.2.1.3 Toxicity to Humans Use Experience Significant injury to the liver, gastrointestinal system, and blood has been reported rarely in connection with industrial exposure. However, many cases of industrial injury to the skin and to the nasal and respiratory mucous membranes have been produced by hexavalent chromium compounds (Browning, 1969). A small part of this involved chromates intended for use in pesticides. Contact dermatitis was reported in men who handled timber treated with chromium (Behrbohm, 1957). Occupational exposure to a variety of chromium VI compounds has been associated with an increased risk of bronchiogenic and nasal cancers (for a review, see ATSDR, 2000). There is now sufficient evidence to classify chromium VI as a human carcinogen according to the definitions of the International Agency for Research on Cancer (IARC, 1980). There is no evidence that the use of chromates as wood preservatives offers any hazard of cancer. Dosage Response Judging from urinary excretion, absorption of chromium at a rate of not less than 0.02 mg/kg/day led to serious illness; twice that rate of absorption also produced liver injury in other workers, although no symptoms appeared. A ceiling concentration in the working atmosphere of 100 I-lg
Cr(VI) per cubic meter has been recommended [Occupational Safety and Health Administration (OSHA), 1998]. Laboratory Findings Chromium can be detected in almost every sample of normal tissue. For most organs, the median value for concentration is close to the limits of detection. For example, the median concentration in the liver ash is 0.7 ppm or about 0.009 ppm, wet weight. The concentration in plasma does not exceed 0.04 ppm (Gofman et aI., 1964). The lung and skin, however, have median concentrations in the order of 0.17 and 0.22 ppm, wet weight, presumably because of their direct exposure to contamination. Most investigators have found similar results in the United States (Tipton and Cook, 1963). Investigations by Guthrie et al. (1978), Kayne et al. (1978), and Andersen (1981) indicate that measurements of chromium by atomic absorption in normal blood and urine before 1978 were probably too high due to inadequate background correction. Mean or median normal levels in the general population with ranges in parentheses are as follows (ATSDR, 2000): serum 0.006 I-lg/l (0.1-0.17 I-lg/l); urine 0.4 I-lg/I (0.24-1.8); hair 0.234 mg/g (not available); and breast milk 0.30 I-lg/l (0.061.56). Treatment of Poisoning Chromium is chelated by dimercaprol and by calcium disodium edetate, of which the latter is preferred (Hayes, 1975). Otherwise treatment is symptomatic (Hayes, 1975). Chelating agents are also valuable for treating chrome ulcers of the skin. A 10% calcium disodium edetate ointment was effective in treating all of 54 chrome skin ulcers. In about 90% of cases, the complex chrome salts adherent to the base of the ulcer could be removed painlessly after the ointment had been applied for 24 hr. In the remaining cases, removal of the adherent chrome could be carried out after 48 or at most 72 hr. Once its base had been cleared of chrome salts, the ulcer healed promptly (Maloof, 1955).
61.3 COPPER Copper compounds are not an important source of poisoning [for a general review, see U.S. Environmental Protection Agency (U.S. EPA, 1987)]. With few exceptions, those used as pesticides owe their mammalian toxicity to a massive overdose of copper ions, especially the cupric ion. Because many of the compounds do not dissolve readily, their toxicity is low. However, copper sulfate is a soluble, ionizable compound. It has been the cause of the majority of cases of poisoning involving copper compounds, and its effects are typical of those involving an excess of copper ions. The toxicities of copper acetoarsenite and copper arsenate are related to their arsenic content (see Section 61.12). Copper is an essential element. It is closely associated with the absorption and metabolism of iron. Zinc and molybdenum also interact with copper. Copper is necessary for the formation of hemoglobin, although it is not part of the molecule. When copper is deficient, hemoglobin is not formed at the normal rate
61.3 Copper
even though there is a reserve of iron in the liver. Copper is also required for bone formation. The element is also essential for carbohydrate metabolism, catecholamine biosynthesis, and the cross-linking of collagen, elastin, and hair keratin. Copper is an integral part of more than 12 specific proteins, including cytochrome oxidase, tyrosinase, ascorbic acid oxidase, uricase, catalase, and peroxidase. Deficiency of copper occurs occasionally in domestic animals and is easily produced experimentally. Signs of copper deficiency include anemia, gastrointestinal disturbances, depressed growth, dystrophy of bone, depigmentation of hair or wool, impaired reproduction, and heart failure (Underwood, 1977). Copper deficiency has not been described in humans, probably because adequate copper occurs in such a wide variety of human food. Many vegetables, cereals, and meats contain between 1 and 10 ppm of copper, but liver contains about 24 ppm and oysters about 36 ppm. The normal dietary intake of copper usually does not exceed 5 mg/day but usually does exceed 2 mg/day (Davies and Bennett, 1983). Kehoe et al. (1940) reported an average intake of 2.32 mg/day. Additional intake such as that from occupational exposure leads to little increase in the retention of copper. Drinking water is not an important source of copper except in rare cases where soft water has been contained in copper piping for long periods (Piscator, 1979). The existence of a hereditary disease (hepatolenticular degeneration or Wilson's disease) characterized by abnormal absorption and retention of copper from a normal diet calls attention to the precision with which copper metabolism is regulated in healthy persons. Some information on the concentration of copper in different normal tissues is given in Section 61.3.1.3. 61.3.1 COPPER SULFATE 61.3.1.1 Identity, Properties, and Uses Chemical Name Structure
Cupric sulfate.
CUS04.
Synonyms Copper sulfate also is known as blue copperas, bluestone, blue vitrol, Roman vitrol, and Salzburg vitrol. It occurs in nature as the mineral hydrocyanite. The CAS registry no. is 7758-98-7. Physical and Chemical Properties Copper sulfate occurs as the anhydride, a monohydrate, and, the pentahydrate, which is the form used as a fungicide or algicide. The molecular weight for the anhydride is 159.61, and that for the pentahydrate is 267.6. The pentahydrate forms a blue, crystalline, odorless solid with a metallic taste. It occurs in nature as the mineral chalcanthite. Its density at 15.6°C is 2.286. Most formulations of copper sulfate contain 98-99% pure salt. The compound is soluble in water (316 g/l at O°C) but insoluble in ethanol and most organic solvents. Copper sulfate solutions are strongly corrosive
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to iron and galvanized iron. The crystals are slightly efflorescent in air. History, Formulations, and Uses The fungicidal activity of soluble copper salts was discovered in 1807 by B. Prevost. The algicidal properties of the salts came into practical use about 1895. Copper sulfate is used as a fungicide for control of downy mildew, blights, leaf spots, apple scab, bitter rot, and peachleaf curl. It also is used as an herbicide for the control of algae in water. Copper compounds are also used as algicides and insecticides. Products containing copper compounds are frequently used with other chemicals. Bordeaux mixture, formed from CUS04 and Ca(OHh, is used as a fungicide and seed treatment. 61.3.1.2 Toxicity to Laboratory Animals Basic Findings The oral LD 50 of copper sulfate for the rat is 960 mg/kg (Stokinger, 1981) and in mice is 87 mg/kg (Jones et al., 1980). The reason for large species differences is not known. Poisoned animals show violent retching, muscular spasms, and collapse. The onset is within a few minutes of dosing, and many rats die within an hour. However, some survive the gastrointestinal irritation only to die several days later of systemic effects (Lehman, 1951, 1952). Rats fed copper sulfate for 4 weeks at a dietary level of 500 ppm as copper (about 25 mg/kg/day) showed slightly decreased food intake and a slight decrease in growth rate but appeared entirely normal. Higher dietary levels of copper led to progressively greater food refusal. Rats fed 4000 ppm ate less than one-fifth the normal amount of food, lost weight, and died within a week. Part of their trouble was starvation, for they ingested less copper (7.6 mg/rat/day) than rats that survived 4 weeks when fed 2000 ppm (9.8-11.8 mg/rat/day) (Boyden et aI., 1938). Pigs maintained on diets supplemented with copper sulfate (250-425 ppm) for 48-79 days exhibited a gradual development of anemia, jaundice, hepatic necrosis, gastrointestinal hemorrhage, and decreased weight gain (Suttle and Mills, 1966a, b). Pigs exposed to 100-500 ppm of copper ulfate in the diet for 54-88 days experienced reductions in hemoglobin and hematocrit and reduced weight gain (Kline et al., 1971). Lifetime exposure of mice to copper gluconate in drinking water (42.4 mg Cu/kg/day) resulted in a 13% decrease in maximal life span from 986 to 874 days (Massie and Aiella, 1984). Two inhalation studies (lohansson et aI., 1984; Lundborg and Camner, 1984) found no toxic effects in the lungs of rabbits exposed to 0.6 mg/m 3 of copper chloride 6 hr/day, 5 days/week for up to 6 weeks. Pimentel and Marques (1969) exposed 12 guinea pigs to a saturated atmosphere of Bordeaux mixture for 6.5 months, three times daily (the duration of each exposure was not reported), and found micronodular lesions and hystiocytic granulomas. In a brief report, Eckert and lerochin (1982) claim that copper sulfate is the principal toxic agent in Bordeaux mixtures.
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Absorption, Distribution, Metabolism, and Excretion Copper absorption after an oral dose occurs in the upper gastrointestinal tract in mammals (Evans, 1973). Two mechanisms are involved. One is an energy-dependent process involving copper-amino acid complexes (Kirchgessner et al., 1967) and the other involves an inducible carrier protein (Evans and Johnson, 1978). The gastrointestinal absorption of copper is affected by several factors: (1) competition with other metals; (2) the amount of copper ingested; (3) certain dietary components such as phytates and fiber; and (4) the chemical form of copper (for details, see ATSDR, 1990). Absorbed copper is predominantly bound to albumin and is transported to the liver, which is the main storage organ. Copper is incorporated into a number of enzymes. It is secreted in bile and also incorporated into ceruloplasm, an alpha globulin that accounts for about 90% of all copper in plasma. Ceruloplasmin is a major regulator of copper retention and storage. The major route of excretion is the feces via secretion in bile. Urinary excretion is a minor route (Underwood, 1977).
death in 3, hepatic failure in 1, and methemoglobinemia in another (Chugh et al., 1977). Katoh et al. (1977) reported a case that was typical in showing the usual initial signs followed by severe anemia, icterus, and kidney failure but very unusual in that the patient died after surviving 40 days. A disease known as Indian childhood cirrhosis has been associated with high intakes of copper (Sharda, 1984). The disease is characterized by hepatic necrosis, Mallory's hyaline inclusions in many hepatocytes, intralobular fibrosis, and very high copper content in liver (Pundit and Bhave, 1983). It is generally believed that water and milk stored in copper or brass vessels lead to high intakes in children.
Use Experience Ordinary occupational exposure may lead to an itching, papulovesicular eczema. Contact with the eye by copper su1fate dust or even strong solutions may produce conjunctivitis or ulceration of the cornea. Two cases of pneumoconiosis were described in sprayers who had applied Bordeaux mixture to vineyards for several years. Symptoms included shortness of breath, weakness, loss Mode of Action The corrosive effect of large doses of copper of weight, and cough productive of a thick, yellow sputum. sulfate in the gastrointestinal tract leads to shock, which may be Although no tubercle bacilli were found, tuberculosis was dithe cause of death. Damage to the erythrocytes, liver, and kid- agnosed on the basis of diffusely abnormal chest X-rays. In ney may combine to kill patients who survive the initial effects. hospital, both patients improved slowly, both clinically and The biochemical mechanism of excessive doses of copper is not radiologic ally. However, recovery was incomplete, and sympwell understood. toms recurred as soon as the men were reexposed to Bordeaux mixture spray (Pimentel and Marques, 1969). The account in61.3.1.3 Toxicity to Humans dicated that the condition responded to cessation of exposure Therapeutic Use Copper sulfate (300 mg in 100 ml of wa- and may have been improved by rest, but there was no eviter) has been used as an emetic in cases of known or suspected dence that drugs intended for tuberculosis are beneficial against poisoning by other compounds. If vomiting did not occur, in- copper pneumoconiosis. Later study showed that some patients testinal colic, diarrhea, and systemic symptoms often appeared presented not with characteristic respiratory symptoms but with chills, fever, joint and muscular pain, weakness, and/or anorexia (Sollmann, 1957). that became severe enough to cause hospitalization only 2 or Accidental and Intentional Poisoning At least until re- 3 weeks after onset. Although no pathological organism was cently, practically all systemic illness attributed to copper sul- identified, it was assumed that persons suffering from vineyard fate was the result of accidental or suicidal ingestion. Although sprayer's lung were unusually susceptible to infection. Regardsuch acute poisoning is uncommon in the United States, it is less of the exact mechanism, the condition certainly was the cause of death in many in whom it was recognized (Pimente1 common in some countries (Chawla and Mehta, 1973). The corrosive action of the copper may produce a character- and Menezes, 1975, 1976). Apparently no evaluation of the istic stain of the mucous membranes of the mouth and pharynx hazard of using Bordeaux mixture sprays has been made. One and will certainly produce severe painful gastrointestinal irrita- paper (Pimentel and Menezes, 1976) was based on 30 autoption, nausea, and diarrhea. The repeated vomiting of blue-green sies, but there was no indication of the size of the population masses is common. The stools are profuse and watery at first of workers with similar exposure from whom this sample was and later contain blood. Patients often die in shock 2 or 3 hours drawn. Be that as it may, it is astonishing that so serious a disafter ingestion of the poison. If they survive, the absorption of ease was recognized so recently in connection with a kind of copper produces severe hemolysis so that hemoglobinuria and occupational exposure that began in 1882. In fact the recent anemia are present in 5-6 hours, and icterus appears soon after. recognition of the condition and the restricted geographical disIf the patient survives a few days, signs of liver damage and re- tribution of cases recognized so far forces one to consider the nal tubular damage may appear. A fatal case was described in possibility that some unrecognized factor may be critical to dedetail by Chugh et al. (1975). In one series of cases, 11 of 29 velopment of the condition. Epidemiological study certainly is patients developed acute renal failure. Intravascular hemolysis needed. appeared to be the chief factor responsible for the renal lesions. Although uremia was controlled adequately by dialysis, only Dosage Response The fatal dose tor humans is difficult to es6 of the 11 patients recovered. Septicemia was responsible for timate because of vomiting, but it is about 109 for adults or
61.4 Zinc
140 mg/kg. This estimate of the fatal ingested dosage is not inconsistent with the observed fatal retained dosage mentioned earlier. An 18-month-old boy narrowly survived a dosage estimated at 262 mg/kg that was reduced to an unknown degree by vomiting and gastric lavage (Walsh et aI., 1977). The usual emetic dose of copper sulfate is 4.3 mg/kg or slightly more. However, an accidental dosage probably no more than half as large as regards copper ion may cause vomiting, diarrhea, and abdominal cramps (Pennypacker et aI., 1975). The average normal intake of copper ion is about 0.03 mg/kg/day. The occupational health saftey limit (threshold limit value) is 0.2 mg/m 3 for copper fumes and 1.0 mg/m3 for copper dusts and mists [American Conference of Governmental Industrial Hygienists (ACGIH),1988]. Laboratory Findings The peak blood copper level in a fatal case was 82.67 ppm (Chugh et al., 1975). A child who narrawly survived had an initial serum level of 16.5 ppm (Walsh et aI., 1977). The milk of female vineyard workers, who were exposed to copper sulfate and a variety of other pesticides, contained 6.2 times as much copper as the milk of milkmaids who did equally hard work but were not exposed to pesticides. Placentas from the vineyard workers contained 4.7 times more copper than those from milkmaids (Nikitina, 1974). Copper is essential to life and occurs in all tissues. Kehoe et al. (1940) reported that the concentrations in blood, muscle, and most of the viscera range from 0.85 to 1.90 ppm. Values for plasma lie in the narrow range 1.16-1.42 ppm (Gofman et al., 1964). The concentrations in the brain (2.2-6.8 ppm), liver (7.1 ppm in adults and 24 ppm in infants), and bone (3.74.7 ppm in rib and 6.8 ppm in long bones) are higher. The concentration in erythrocytes is slightly greater than that in the plasma. The values reported by Tipton and Cook (1963) and by Liebscher and Smith (1968) are basically similar but expressed on a different basis. The organs of Orientals frequently contain more copper than those of Americans (Tipton et aI., 1965). The concentration of copper in the urine varies from 0.01 to 0.08 and averages about 0.034 ppm (Kehoe et aI., 1940). Fecal excretion is greater than urinary excretion and averages 1.96 ing/day according to Kehoe et al. (1940). Pathology Persons who die soon after ingestion of copper sulfate show on autopsy a characteristic staining of the lining of the digestive tract and fatty degeneration of the liver, kidney, and to some degree other organs. Persons who develop acute renal failure as part of acute poisoning by copper sulfate mayor may not show well established acute tubular necrosis. In those who survive long enough, granulomatous lesions of the kidney may develop (Chugh et al., 1977). Biopsy revealed nodules containing copper in the lungs of men who developed pneumoconiosis following years of exposure to Bordeaux mixture (Pimentel and Marques, 1969). Later study of biopsy and autopsy material from workers exposed to Bordeaux mixture sprays for 3 to 35 years revealed that many rural workers who developed pulmonary granulomas
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developed liver granulomas also, and in some the liver was enlarged. The liver granulomas varied, with all transitional forms, from those consisting entirely of histiocytic cells to those consisting of clusters of epithelioid cells perfectly organized in sarcoid-type follicles. No giant cells or lymphocytic borders were observed. A finely granular material in the granuloma cells and in prominent Kupffer cells did not stain for ferric iron but was positive for copper when stained with rubeanic acid. No copper was demonstrable in the hepatocytes. Thus the distribution of copper in the liver was entirely different from that in any previously described granulomatosis of whatever origin. Some of the patients also had micronodular cirrhosis or fatty change of the liver, but it was thought that this might be due to alcoholism. Angiosarcoma and idiopathic portal hypertension also were seen (Pimentel and Menezes, 1975, 1976). Treatment of Poisoning Treatment should include a prompt effort to prevent absorption and later the use of a chelating agent. The protein of milk or egg white combines with copper to form an insoluble copper proteinate. However, the product must be removed by vomiting or lavage before it is digested and the copper released. Copper may also be precipitated by potassium ferrocyanide given in a dose of 600 mg in a glass of water. British anti-lewisite (BAL), dicalcium EDTA, and penicillamine are all effective in binding copper. Penicillamine is definitely the drug of choice in Wilson's disease. There is less experience on which to base a choice in the treatment of acute poisoning.
61.4 ZINC Zinc follows copper in the periodic table, and, like copper, it is an essential element. More than 20 metalloenzymes containing zinc have been identified. These include alcohol dehydrogenase, alkaline phosphatase, carbonic anhydrase, and DNA polymerase [National Research Council (NRC), 1979]. Zinc forms a necessary part of carbonic anhydrase, carboxypeptidase, alcohol dehydrogenase, lactic acid dehydrogenase, glutamic dehydrogenase, and alkaline dehydrogenase molecules (Vallee, 1959, 1962). Zinc also plays an essential role in maintaining the structure of nucleic acids in genes, such as the zinc finger structures (ATSDR, 1994). It has been known for some time that a deficiency of zinc leads to testicular atrophy and failure of body growth in animals. Zinc deficiency, complicated only by iron deficiency, has been found in patients from villages in the Middle East who exhibited severe growth retardation and sexual hypofunction. In adolescent patients, zinc supplementation alone resulted in body growth, gonadal development, and the appearance of secondary sexual characteristics. Some of the dwarfs who received reagent grade iron but no zinc failed to develop sexually, and their growth rate was less than that following zinc (Prasad, 1966). Dermatitis also is seen in zinc deficiency. Animal experiments indicate that simultaneous administration of cadmium
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will exacerbate the effects of zinc deficiency (Petering et aI., 1971). Poor intestinal absorption of zinc is believed to be a factor in the familial disease acrodermatitis enterohepatica (Moynahan, 1974). Zinc is a normal constituent of food, and only rarely is it possible to avoid an adequate intake. The average daily intake of zinc in different areas of the world ranges from 5 to 22 mg in adults (Halsted et al., 1974). Protein-rich foods, especially marine organisms sach as oysters, have the highest levels (10-50 mg/kg wet weight), whereas concentrations in grains, vegetables, and fruits are relatively low, usually less than 5 mg/kg fresh weight (Great Britain Ministry of Agriculture, Fisheries, and Food, 1981). The concentration in drinking water is usually the same as in freshwater and seawater (1-10 J.l.g/I) but may be much higher if the water is passed through zinc-coated pipes (Sharrett et al., 1982). Information on the concentration of zinc in normal human tissue is given in Section 61.4.1.3. Pesticides containing zinc may be divided into four classes: (a) Most inorganic and certain organic compounds of zinc that are toxic because sufficient doses supply an excess of zinc ions. Examples include zinc chloride and zinc acetate. These compounds are discussed in the following section. (b) Complexes of zinc with other elements, the greater toxicity of which predominates in the effect of the complex regardless of the exact chemical combination. An example is zinc arsenate. (c) Zinc phosphide, which owes its toxicity to the phosphine (PH3) it produces (see Section 61.4.2 and see Section 14.6.2 of the first edition of this Handbook). (d) Certain organic compounds of zinc, the toxicity of which is not essentially different from those of similar compounds that are salts of other metals. An example is zineb (see Section 21.12.5 of the first edition of this Handbook). 61.4.1 ZINC CHLORIDE 61.4.1.1 Identity, Properties, and Uses Chemical Name Structure
Zinc chloride.
ZnCh.
Synonyms Other names include butter of zinc and zinc dichloride. The CAS registry no. is 7646-85-7. Physical and Chemical Properties Zinc chloride has the empirical formula Cl2Zn and a molecular weight of 136.29. It forms white, odorless, very deliquescent granules with a density of 2.907. It has a melting point of approximately 290°C and a boiling point of 732°C. One gram of zinc chloride dissolves in 0.5 ml water, 1.3 ml ethanol, or 2 ml glycerol. The compound is freely soluble in acetone. Zinc oxychloride is formed in the presence of water. Formulations and Uses Technical grade is at least 95% pure, the remainder being mostly water and oxychloride. Zinc chloride may be used with copper and chromium compounds as a
wood preservative. Zinc chloride is also used as an herbicide [Hazardous Substances Data Bank (HSDB), 1993]. 61.4.1.2 Toxicity to Laboratory Animals Basic Findings Zinc chloride may be taken as typical of those compounds that owe their toxicity to the zinc ion. However, to gain a reasonably complete picture of the toxicity of the ion it is necessary to refer to other compounds also. The small differences in toxicity between most zinc compounds are explained by differences in their solubility and degree of ionization. The acute (less than 14 days oral intake) and the intermediate (14 days to 1 year) LD 50s are in the range 100-1000 mg Znlkg/day in rats and mice (ATSDR, 1994) for zinc chloride and similar compounds of zinc. Oral administration of zinc compounds can depress hemoglobin levels. The lowest reported observed effects level in rats was 12 mg Znlkg/day as zinc chloride given over 4 weeks in the drinking water (Zaporowska and Wasilewski, 1992). However, most studies report lowest observed effects levels above 100 mg/kg/day (ATSDR, 1994). Absorption, Distribution, Metabolism, and Excretion Early studies (Drinker et aI., 1927a, b) indicate that animals can regulate zinc absorption such that large oral doses produce minimal change in tissue levels. Thus intestinal absorption is highly variable (10-90%), depending on zinc status and the magnitude of the oral dose. Also, calcium and phytates in food interfere with zinc absorption (Becker and Hoekstra, 1971). Metallothionein may play a role in regulation of zinc absorption (Richards and Cousins, 1976). Another protein, also rich in cysteine residues, also plays a role (Hempe and Cousins, 1991). In the plasma, albumin is the primary carrier of zinc, which represents the metabolic ally active pool (ATSDR, 1994). Zinc is initially concentrated in the liver after ingestion, and is subsequently delivered throughout the body. The liver, pancreas, bone, kidneys, and muscle are the major tissue storage sites. Zinc is secreted in bile (Barrowman et aI., 1973) and fecal excretion is considerably greater than urinary excretion in both animals (Schryver et al., 1980) and humans (Aamodt et aI., 1979). Biliary secretion may be a glutation-dependent process (Alexander et aI., 1981). Mode of Action The mechanism of toxicity of zinc in general is not well understood. The lowered hemoglobin levels are believed to be due to competition of zinc with copper to produce a relative copper deficiency (ATSDR, 1994). Treatment of Poisoning in Animals By using 65Zn injected intravenously as the chloride, it was shown that the intramuscular injection of BAL doubled the concentration of zinc in the red cells, increased urinary excretion as much as 20 times (from a level of 4% or less of the administered dose), and decreased fecal excretion (from a level of 31 % to 42% of the administered dose in rats that received no BAL). Total excretion of zinc was increased. Under the conditions used, the injection of BAL completely eliminated acute toxicity but did not prevent death
61.4 Zinc
72-96 hours after injection. In fact, BAL seemed to accentuate the renal damage caused by sufficient zinc (Bruner, 1950). The study throws no light on the question of whether BAL would divert a harmful amount of zinc through the kidneys under practical conditions of accidental poisoning. 61.4.1.3 Toxicity to Humans Therapeutic Use Zinc oxide and zinc carbonate, which are highly insoluble, are applied liberally to inflamed skin in the form of powder, calamine lotion, and the like. Zinc chloride, sulfate, and acetate have been used for their antiseptic, astringent, or caustic properties. Their use is limited by difficulty of local control, not by systemic effects. Zinc sulfate has been used as an emetic at an oral dose of 1000 or 2000 mg in a glass of water (Sollmann, 1957). This dose intentionally produces a concentration in the water well above the threshold (675-2280 ppm) that may lead to vomiting if taken on an empty stomach, as sometimes happens accidentally when fruit juice or other acid drinks are stored in galvanized vessels. Accidents and Use Experience Zinc chloride is caustic. Stokinger (1963) recorded without detail or reference that it has caused dermatitis in men working with railroad crossties treated with the compound as a fungicide. More direct contact with zinc chloride, as in its use as a soldering flux, may cause ulceration of the fingers, hands, and forearms. Respiratory exposure to a sufficient concentration of zinc chloride smoke is highly irritating to the mucous membranes of the nasopharynx, trachea, and bronchi and may be fatal (Evans, 1945). Schaidt et al. (1979) have proposed that the high toxicity of ZnCb in the lung is due to the formation of hydrochloric acid. Zinc chloride intended for use as a pesticide apparently has not led to injury other than dermatitis. Ingestion undoubtedly would produce illness similar to that caused by copper sulfate. In several instances, the preparation or storage of an acid food in a galvanized or zinc-plated vessel has led to severe vomiting and to headache and discomfort in the chest (Hegsted et aI., 1945). Dosage Response The lethal dose in humans is unknown but has been estimated at 3000-5000 mg (43-71 mg/kg). The emetic dosage is 14-28 mg/kg. Zinc intake from food and water is usually at the rate of about 0.14-0.21 mg/kg/day but in some communities it may reach 0.75 mg/kg/day without injury. To prevent respiratory irritation and other effects of fumes from ZnCb, the American Conference of Governmental Industrial Hygienists (ACGIH, 1981) have set a threshold limit value of 1 mg/m3 and a short-term exposure limit of 2 mg/m 3. Laboratory Findings There is only moderate variation between the normal concentrations of zinc in different tissues; in most organs it is about 20-30 ppm; in liver, bone, and voluntary muscle it is from 60 to 180 ppm. Higher concentrations are found in the prostate (860 ppm) and the retina (500-1000 ppm), whereas blood contains 6.6-8.8 ppm, of which the greater part
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is in the red cells. The plasma contains only 0.93-1.03 ppm (Gofman et al., 1964), but whole blood contains 0.60-19.87 ppm with a mean of 5.30 ppm (Kubota et aI., 1968). Excretion of zinc is chiefly via the feces (Vallee, 1959). The daily urinary output is about 0.5 mg (Elinder et al., 1978; Halsted et aI., 1974). In studies of normal unexposed humans, the biological halflife of zinc is in the range 160-500 days (Aamodt et aI., 1975). Bone and muscle, which contain the major amount of the body's zinc, have longer biological half-lives than other tissues, for example, liver (NRC, 1979). 61.4.2 ZINC PHOSPHIDE 61.4.2.1 Identity, Properties, and Uses Chemical Name
Zinc phosphide.
Synonyms As a rodenticide, zinc phosphide has been manufactured under tradenames Fasco Field rat powder®, Kilrat®, Mouse-Con®, and Rumetan®. The CAS registry no. is 131484-7. Physical and Chemical Properties Zinc phosphide has the empirical formula P2Zn3 and a molecular weight of 258.09. It forms a gray-black crystalline powder with a faint garlic odor and taste. It has a density of 4.55, a melting point above 420°C, and a boiling point of 1l00°C. It is insoluble in water and alcohol. Zinc phosphide is stable when dry but spontaneously flammable on contact with acids. It is decomposed by acids into phosphine and is slightly corrosive to metals. Formulations and Uses Technical zinc phosphide is 80-90% pure. Rodenticidal baits contain 0.5 or 1.0% of the compound; pastes contain 5-10%. As reviewed by von Oettingen (1947), zinc phosphide has been used not only as a rodenticide but also as an insecticide for control of mole crickets. 61.4.2.2 Toxicity to Laboratory Animals Basic Findings The oral LD 50 of zinc phosphide for rats is 40.5 mg/kg (Dieke and Richter, 1946). Because the complete reaction of 40.5 mg of zinc phosphide produces 10.6 mg of phosphine and because the fatal dosage of phosphine for rats exposed to vapor concentrations ranging from 564 to 7.5 mg/m 3 varies from only 13.5 to 8.9 mg/kg (see Section 14.6.2.2 in the first edition of this Handbook), the toxicity of zinc phosphide is fully accounted for by the toxicity of the phosphine it produces when hydrolyzed by the acid of the stomach. The symptoms produced by the two compounds are similar except that respiratory exposure to the gas may have a slightly greater tendency to produce pulmonary edema and associated symptoms. However, the emetic action of its zinc moiety reduces the toxicity of zinc phosphide to humans and other animals that
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can vomit. The usual emetic doses of hydrated zinc sulfate in humans (1000-2000 mg) are equivalent to zinc intakes of 3.26.4 mg/kg, whereas the oral LD 50 of zinc phosphide for rats is equivalent to a zinc uptake of 30.8 mg/kg. Thus, a dangerous dose has a powerful emetic action in humans. At the same time, little injury from a dangerous dose of zinc phosphide can be attributed to the zinc moiety. Absorption, Distribution, Metabolism, and Excretion According to Curry et al. (1959), both phosphide and phosphine were demonstrated in the liver of rats poisoned by zinc phosphide, but were only demonstrable in those that had survived long enough to excrete most of the poison. Although this conclusion probably is correct, the data do not support the authors' assumption that phosphide is absorbed in paniculate form. It is reported that the main urinary excretion product of zinc phosphide in rats and guinea pigs is hypophosphite (Curry et aI., 1959). 61.4.2.3 Toxicity to Humans Accidental and Intentional Poisoning One of the bestdescribed cases of poisoning by zinc phosphide involved a 37year-old woman who, with suicidal intent, drank a mixture of 180 g of the rodenticide with water. Vomiting began 1 hr after ingestion and was frequent and violent. She was discovered in a state of shock after about 5 hr. On admission to hospital, her skin was cold and blue; the heart was inaudible, no limb pulses were palpable, and blood pressure was unobtainable. The carotid pulse was 80 per minute. The breath smelled of phosphine. Rectal temperature was 33°C. Treatment was entirely appropriate and almost certainly prolonged the patient's life. Besides very thorough gastric lavage and the use of detoxifying antidotes, the important features of treatment included rehydration and efforts to combat severe metabolic acidosis. Within 8 hr, 1200 mEq of sodium bicarbonate was administered. An attempt to induce diuresis was promptly but only briefly successful. Peripheral limb pulses returned, blood pressure reached 90/60 mm Hg, and by 21 hr after ingestion the patient was conscious, rational, and oriented, and she gave a lucid account of the events leading to her condition. However, complications had developed already and others appeared later. By 16 hr after ingestion, serum bilirubin had risen to 2.2 mg/lOO ml. Hepatic tenderness appeared, and other tests of liver function became abnormal. Blood pressure and urine output decreased, and blood urea reached 30 mg/l 00 ml. A thrombotest showed 28% of normal coagulation activity. Variable tetany appeared; although blood calcium was low (3.3 mEqll), the condition did not respond to injection of calcium gluconate. Abdominal pain became severe. The extremities again became pulseless and icy cold. Fever and rapid breathing preceded a rapidly developing cortfusional state. For a brief period, the patient seemed to suffer terrifying hallucinations, and toward the end she screeched repeatedly. Unexpected cardiac arrest occurred 41 hr after ingestion (Stephenson, 1967). In a drunken state, a 19-year-old woman in her 30th or 31st week of pregnancy ingested an unknown amount of zinc phos-
phide intended as a rodent bait. She soon lost consciousness and was cyanotic when brought to hospital. However, her pulse was detectable and her blood pressure normal. Treatment included gastric lavage and supportive measures. Recovery was complete by the third day. A baby girl later was born at term by normal labor; she was normal at birth and in subsequent development (Kuptsov and Aslanov, 1970). Symptoms are basically the same regardless of route of administration whether oral, vaginal (Santini, 1955), respiratory (Elbel and Holsten, 1936), or directly into the subcutaneous and muscular tissues (Blisnakov and Iskrov, 1961). In a review of cases where there was enough information to reach a conclusion, it was found that there was no mortality following three industrial accidents with zinc phosphide. However, mortality was 66% in 3 domestic accidents, 69% in 26 suicides, and 38% in 8 attempted murders. Ten patients who died did so within 7-58 hr after ingestion, with an average of 24.6 hr. Patients who survived for 3 days were out of danger, although some suffered liver and/or kidney injury for days before full recovery (Stephenson, 1967). Apparently no exception to this rule regarding survival has been observed even though Rimalis and Bochkamikov (1978) reported a case in which hepatorenal insufficiency was noted 5 days after ingestion, and the patient was not discharged until 36 days after ingestion. Use Experience Zinc phosphide has given rise only rarely to occupational poisoning, although accidental poisoning of children by the baits is a real possibility. In one case, inhalation of dust from grain coated with the compound was followed several hours later by vomiting, diarrhea, cyanosis, rales, tachycardia, meteorism, restlessness, fever, albuminuria, and eventual recovery (Elbel and Holsten, 1936). In another instance reviewed by von Oettingen (1947) it apparently was not the applicator but workers in a fish-processing plant who were affected by phosphine generated from zinc phosphide after the rodenticide had come in contact with acid brine used for curing fish. Dosage Response Adults have been killed by doses of 4000 or 5000 mg (Frketic et al., 1957; Gili, 1948). However, others have survived as much as 25,000 mg (Paszko, 1961), 50,000 mg (Simonovic, 1954), or even 100,000 mg (Rimalis and Bochkamikov, 1978). Early vomiting improves the prognosis. In fact, one of two young women who had ingested similar amounts of zinc phosphide in a suicide pact survived with only transient symptoms because she had been induced to vomit; the other, who would not vomit, died in spite of gastric lavage 1 hr after ingestion (Mannaioni, 1960). Laboratory Findings Phosphine may be detected most readily by odor. The concentration of zinc in the tissues is increased; in one case the serum level was between 5.9 and 6.1 ppm (Stephenson, 1967). For zinc in normal tissues, see Section 61.4.1.3. Other findings may include metabolic acidosis, reducing agents in the urine, increased serum bilirubin and other abnormal tests of liver function, thrombocytopenia, methemoglobin, and electrocardiogram (ECG) abnormalities.
61.5 Cadmium The reducing substances in the urine include glucose but may induce hypophosphite and dissolved phosphine (Stephenson, 1967). Pathology When death is rapid, abnormal findings may be restricted to pulmonary edema (Frketic et al., 1957; Mannaioni, 1960; Paszko, 1961; Puccini, 1961; Stephenson, 1967) or to this and cerebral edema (Mannaioni, 1960; Puccini, 1961). Other findings that may be present especially in persons who survive longer include centrilobular necrosis of the liver, tubular necrosis of the kidneys, mucosal hemorrhage of the stomach, and bloody pleural, peritoneal, or pericardial fluid (Stephenson, 1967). Treatment of Poisoning (Hayes, 1975).
Treatment is entirely symptomatic
61.5 CADMIUM Cadmium is a metal in the same periodic group as zinc and mercury. The toxicity of cadmium is especially evident when there is a deficiency of zinc, and within limits cadmium toxicity may be counteracted by supplementing the diet with zinc. A wide range of cadmium concentrations are in human food-0.005-0.1 mg/kg wet weight. Kidney and oysters may have concentrations exceeding 1 mg/kg. Drinking water usually has concentrations below 5 !-lgll, but higher levels may be found due to cadmium impurities in the zinc of galvanized pipes and cadmium-containing solders in pipe fittings. More details are available in reviews of cadmium toxicity (Foulkes, 1974; Friberg et aI., 1986). Several cadmium salts showed some promise as insecticides, but their use never became extensive. A number of cadmium salts have been used as fungicides on turf. 61.5.1 CADMIUM CHLORIDE 61.5.1.1 Identity, Properties, and Uses Chemical Name Structure
Cadmium chloride.
CdCh.
Synonyms The CAS registry no. for cadmium chloride is 10108-64-2. Physical and Chemical Properties The empirical formula for cadmium chloride is CdCh and the molecular weight is 183.32. The compound forms colorless, odorless crystals with a density of 4.05, a melting point of 568°C, and a boiling point of 960°C. It is freely soluble in water and acetone, slightly soluble in methanol and ethanol, and practically insoluble in ether. Use
Cadmium chloride is a fungicide for turf.
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61.5.1.2 Toxicity to Laboratory Animals Basic Findings Acute exposures (less than 14 days) to aerosols of cadmium chloride or similar soluble compounds result in a lethal outcome in mice, rats, and other small mammals at air concentrations generally above 10 mg Cd/m3 ; intermediate exposures (2 weeks to 1 year), at levels above 0.1 mg Cd/m 3 ; and longer exposures, at levels above 0.05 mg Cd/m3 (ATSDR, 1997). Acute exposures (less than 14 days) to orally administered cadmium chloride and chemically similar cadmium compounds are lethal to rats at intakes above 10 mg Cd/kg/day, and intermediate (2 weeks to 1 year) exposures at intakes above I mg Cd/kg/day (ATSDR, 1997). Cadmium compounds can produce a number of serious systemic effects, including anemia, kidney damage (the main target organ after long-term exposures), and skeletal disturbances such as osteoporosis. The latter may be due, at least in part, to alterations in renal metabolism of vitamin D. In rats and mice, acute oral exposure at near lethal doses can produce testicular atrophy and necrosis (Andersen et al., 1988; Bombard et al., 1987). Cadmium inhalation can cause lung cancer but only in rats (Oldiges et aI., 1989; Takenaka et al., 1983). Absorption, Distribution, Metabolism, and Excretion Absorption from the gastrointestinal tract usually is limited by rapid and violent vomiting. Absorption may occur from the respiratory tract following exposure to dusts and aerosols. Dermal absorption is not significant. Once absorbed, cadmium is tenaciously stored and only slowly excreted. Cadmium is believed to be carried to the liver attached to serum albumin. In the liver, it induces metallothionein, a protein of molecular weight 6000 that avidly binds cadmium. It is lost from the liver as a cadmium-metallothionein complex, filtered as the glomerulus, and take up by the proximal tubular cells of the kidney. Once inside the kidney cell, the metallothionein-cadmium complex is broken down by lysosomes to release free cadmium. The latter can induce metallothionein synthesis in the kidney cells. However, if the rate of release of cadmium exceeds the ability of the cell to produce metallothionein, the cadmium will attach to other "sensitive sites" and damage the cell. The ability of cadmium to induce metallothionein and to bind avidly to this protein is believed to account for its long-term storage in the body (Friberg et al., 1986). Storage is mainly in the pancreas, liver, and kidney (Friberg, 1956) and, following respiratory exposure, in the lung. Biochemical Effects Nonfatal doses of cadmium induce the synthesis of metallothionein, as just discussed, and may cause an increased tolerance for the ion. When groups of mice were given four-tenths of the LD 50 dosage of cadmium chloride previously determined for the strain and then challenged 48 hr later, the intraperitoneal LD 50 was increased from 5.2 mg/kg in controls to 6.7 mg/kg. The difference was statistically significant. A significant increase was obtained with cadmium acetate also (Jones et al., 1979).
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61.5.1.3 Toxicity to Humans Accidental and Intentional Poisoning Apparently, the limited use of cadmium chloride as a pesticide has not led to poisoning of anyone. However, the evidence that has accumulated about the high toxicity of cadmium and its very tenacious storage leave no doubt that cadmium in any form should not be used as a pesticide. Dramatic but not fatal poisoning used to occur fairly frequently when fruit juices or other acid foods were held for a time in cadmium-plated vessels. The citric or other organic salts formed on standing were fully ionized in the presence of hydrochloric acid in the stomach. Signs and symptoms included salivation, nausea, persistent vomiting, mild diarrhea, abdominal pain, and tenesmus. Illness often began suddenly 0.25-5 hr after ingestion; the interval was usually less than 2 hr unless food was eaten at the same time. Recovery usually was well advanced in 1-2 hr (Cangelosi, 1941; Frant and Kleeman, 1941). The amount of cadmium necessary to produce these effects was not great. Schwartze and Alsberg (1923) cited a case of voluntary ingestion of cadmium sulfate, equivalent to about 18 mg of metallic cadmium, which produced vomiting. Mild illness is produced by concentrations as low as 15 ppm.
Use Experience and Environmental Exposure The few episodes of acute occupational poisoning by cadmium have all been associated with inhalation of dust or fumes. The cases were characterized by severe pulmonary irritation, and persons who died showed marked hyperplasia of the lung epithelium and thickening and edema of the alveolar septa. Signs of severe gastrointestinal irritation were present in a high proportion of cases. Less severe but more prolonged occupational exposure to cadmium may produce a distinctive form of emphysema leading to dyspnea and often accompanied by a low-grade anemia. The most typical feature of chronic cadmium poisoning is kidney damage. The first sign is the excretion of low-molecularweight proteins. This condition is known as tubular proteinuria because these proteins are normally reabsorbed by the tubular cells in the kidney (Friberg, 1950; Piscator, 1966). At high exposure, more severe effects in kidney function occur, such as aminoaciduria (Clarkson and Kench, 1956), glucosuria, and phosphaturia (Piscator, 1966). By far the largest number of cases of cadmium poisoning have occurred in Japan as a result of environmental exposure. These cases differed greatly from those caused by cadmium under other circumstances, being characterized by osteomalacia, skeletal deformity, and very severe pain. In fact, the condition took its name "itai-itai disease" from the predominance of pain. The entire matter has been reviewed critically by Friberg et al. (1974), who concluded that there was no doubt that cadmium was the cause but that there was reason to believe that high intake of cadmium and deficient consumption of certain essential food items, especially calcium and vitamin D, had been contributing factors. In spite of the very different clinical manifestations seen in itai-itai disease, it is thought that the basic
injury is to the kidney as in other forms of chronic poisoning by cadmium and that the osteomalacia is secondary. The World Health Organization (WHO) has proposed a guideline for drinking water of 5 J.Lg Cd/I (WHO, 1984a) and a provisional tolerable weekly intake of 004-0.5 mg (WHO, 1984b). The D.S. Occupational Safety and Health Administration recommends a permissible exposure limit for all cadmium compounds of 5 J.Lg Cd/m3 in the occupational setting (OSHA, 1992). Dosage Response Apparently, no information is available on the minimal dosage of a cadmium salt that has been fatal in humans. According to reports cited by Schwartze and Alsberg (1923) doses of 250-1000 mg of cadmium bromide produced illness, and a 3-mg dose of cadmium sulfate produced vomiting. The World Health Organization (WHO) has proposed a guideline for drinking water of 5 J.Lg Cd/I (WHO, 1984a) and a provisional tolerable weekly intake of 004-0.5 mg (WHO, 1984b). The D.S. Occupational Safety and Health Administration recommends a permissible exposure limit for all cadmium compounds of 5 J.Lg Cd/m3 in the occupational setting (OSHA, 1992). Laboratory Findings Friberg et al. (1974) have given reasons for thinking that some of the early reports of concentrations of cadmium in environmental samples and human tissues were incorrectly high because the chemical methods used formerly were not entirely specific. This is an unusual situation because improvement in chemical methods often leads to more complete recovery and recognition of the material analyzed and thus to higher values. Considering only results based on dependable chemical methods, it now seems likely that blood levels in normal adults usually are <0.01 ppm. The renal cortex usually contains cadmium at concentrations of 20-50 ppm, higher than the concentrations in any other tissue. The highest concentration is reached at about 50 years of age, and levels then decline gradually. The kidneys contain about one-third of the total body burden, and the kidneys and liver together contain about half the body burden. Normal urinary excretion is about 0.002 mg/person/day or less, but levels increase with age. Intake is about 0.05 mg/person/day from food, with much less from water, air, and even cigarette smoking. The amount of cadmium in ordinary food is one-seventh to one-fifth of the intake thought to be required to produce the storage of 200 ppm in the kidney and consequent kidney damage. Absorption of cadmium ingested in food and water ranges from 4.7 to 7.0% but may be higher if calcium and protein are deficient. Retention of cadmium from smoke or fumes may be 25-50%. The placenta is an efficient barrier to small concentrations of cadmium; the newborn contains less than 0.001 mg. The biological half-life for cadmium in the whole body lies between 20 and 30 years. Thus cadmium has an unusually great tendency to accumulate in the body (Friberg et aI., 1974). In workers, cadmium in the blood is mainly in the cells; the concentration in whole blood usually is between 0.01 and 0.1 ppm. Concentrations in the kidney cortex may be around 300 ppm. The liver
61.6 Mercury
may contain proportionately more than would be expected in persons without special exposure. Cadmium is transported in the circulation at least in part bound to metallothionein. When workers begin to excrete protein in their urine, their excretion of cadmium increases, sometimes dramatically, and their blood levels fall correspondingly. Based on results in animals and on autopsy findings in workers, a concentration of 200 ppm in the kidney is the threshold for kidney injury as reflected by kidney function tests and the urinary excretion of protein. More recent data are available based on direct in vivo measurement of the kidney content of cadmium in occupationally exposed workers using a neutron activation technique (Ellis et aI., 1984). This has allowed the development of more sophisticated dosageresponse models (Kjellstrom, 1985) that estimate, for example, that a daily intake of about 200 J.!g Cd via food for 45 years will give a 10% chance of mild kidney effects. The equivalent occupational exposure would be 10 years at about 50 J.!g Cd/m3 . Besides the proteinuria, other laboratory findings may include mild hypochronic anemia, mild jaundice, and the presence of microscopic blood in the urine. Persons mildly poisoned by cadmium may show blood levels as high as 6.2 ppm and urinary levels as high as 2.2 ppm (Cotter, 1958). Treatment of Poisoning Cotter (1958) treated three patients whose major exposure was to cadmium, including at least one without any other exposure. The patients suffered one or more of the following: mental disturbance, cough and other respiratory disturbance, mild jaundice, mild anemia, and abnormal urinary findings. Treatment consisted of 500 mg of calcium disodium EDTA by mouth every 2 hr while the patient was awake and for a period of 1 week. All the patients showed marked clinical improvement. There was no evidence of kidney damage from the treatment, and in fact the proteinuria present before treatment gradually cleared completely. Icterus disappeared, and the red cell count and hemoglobin level improved. Before treatment, blood cadmium levels in the three patients were 0.022-6.2 ppm, and urinary cadmium was 0.036-2.2 ppm. The concentrations fell to undetectable levels following treatment. A patient who had marked irregularity of the heartbeat and other ECG abnormalities 3 hr after ingesting cadmium showed marked improvement following intravenous injection of calcium gluconate and magnesium sulfate, and she was almost normal within 48 hr after ingestion (Lydtin et al., 1965). However, evidence from animals indicates that some chelating agents such as EDTA can cause renal accumulation by redistribution from other tissues, and therefore application of chelating agents to cases of human poisoning should be viewed with great caution.
61.6 MERCURY There is no evidence that any quantity of mercury is beneficial to any form of life. However, the element is widely distributed
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in the environment, and traces of it occur in food, water, and tissues even in the absence of occupational exposure. Mercury is toxic no matter what its chemical combination. However, different forms of mercury have different absorption, distribution, and excretion characteristics; at the same rate of intake, they reach and maintain different concentrations in different tissues. For this and perhaps other reasons they have distinguishable toxic effects. In the late twentieth century the only mercury pesticides used in significant amounts were organic compounds; some of them constitute a real hazard. It is said that mercury vapor formerly was used in India for the disinfection of stored grain. Several inorganic mercury salts are still listed as fungicides. There is no evidence that the use of mercury vapor or inorganic mercury compounds as pesticides constitutes a significant hazard at this time, largely because such use is limited. Even so, some information on the toxicity of these forms of mercury is necessary for understanding the toxicology of the organic mercury fungicides. Finally, among the organic mercurials, the short-chain alkyl compounds are far more toxic than the phenyl or alkoxyalkyl compounds as discussed in Section 61.6.3. Valuable reviews (Berlin, 1986; Berlin et aI., 1969c; Bidstrup, 1964; Clarkson, 1972; WHO, 1976) of the toxicology of mercury and its compounds are available. 61.6.1 ELEMENTAL MERCURY 61.6.1.1 Identity, Properties, and Uses Synonyms Synonyms for elemental mercury (Hg) include mercure (French), mercurio (Italian), kwik (Dutch), Quecksilber (German), quicksilver, and RTEC (Polish). The CAS registry no. is 7439-97-6. Physical and Chemical Properties Elemental mercury is a heavy, silver-white metal which is liquid at room temperature. Its atomic number is 80, and its atomic weight is 200.59. Mercury has a melting point of -38.9°C, a boiling point of 356.9°C, and a density at O°C of 13.5955. The vapor pressure is 2 x 10- 3 mm at 25°C; air saturated at this temperature contains 19.5 mg of mercury (Giese, 1940). Mercury is insoluble in water, alkalies, and most common solvents. Mercury is not attacked by dilute hydrochloride and sulfuric acids but will dissolve in dilute nitric acid and hot, concentrated sulfuric acid. It oxidizes slowly. Use
See Section 61.6.
61.6.1.2 Toxicity to Laboratory Animals Basic Findings The toxicity of elemental mercury is essentially limited to the vapor, and, therefore, it is not convenient to determine dosage on a milligram-per-kilogram basis. Exposure of rats or rabbits to concentrations of elemental mercury vapor at air concentrations of about 30 mg/m 3 for periods of a few hours produced death or severe tissue damage.
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Exposures at air concentrations of 3-6 mg Hg/m3 produced pathological changes in kidney and brain as well as behavioral abnormalities (ATSDR, 1999). Both rabbits and dogs tolerated an exposure of 7 hr/day, 5 days/week for over 70 weeks at 0.1 mg Hg/m 3 without functional or microscopic injury (Ashe et aI., 1953). Rats apparently are less susceptible. When they were exposed for only 2 hr each day for 30 days at a concentration of 17 mg/m 3, they developed only a delay in escape response plus an increase in the duration and severity of reflexive fighting behavior and of actual fighting (Beliles et al., 1968). Absorption, Distribution, Metabolism, and Excretion Inhaled mercury vapor diffuses across the alveolar regions of the lung into the bloodstream (Berlin et al., 1969a). Mercury vapor is a monatomic gas which is highly diffusible and lipid soluble (Hursh, 1985). Once in the bloodstream, mercury vapor enters the red blood cells, where it is oxidized to divalent inorganic mercury under the influence of catalase (Halbach and Clarkson, 1978; Magos et aI., 1978). The oxidation can be inhibited by alcohol, thereby decreasing the retention of inhaled vapor (Nielsen-Kudsk, 1965, 1969). Despite oxidation in the red blood cells, dissolved mercury vapor persists in the plasma for sufficient time for it to be transported to other tissues. This explains why 10 times more mercury is retained in the brain after exposure to mercury vapor than after an equivalent intravenous dose of mercuric mercury in both mice (Berlin and Johansson, 1964) and primates (Berlin et al., 1969b). Placental transport of mercury is also greater after exposure to the vapor (Clarkson et aI., 1972). Oxidation by catalase also takes place in fetal tissues (Dencker et al., 1983). Except for these differences, the pattern of organ distribution of mercury after exposure to vapor or mercuric salts is generally simitar, with the highest concentrations always found in the kidney cortex. Excretion is mainly by urine and feces at roughly similar rates but there is a small loss of mercury in expired breath (Clarkson and Rothstein, 1964; Rothstein and Hayes, 1960, 1964). 61.6.1.3 Toxicity to Humans Experimental Exposure Using a group of four volunteers in each experiment, the concentrations of radioactive mercury vapor in the inspired and expired air were compared. The inspired air contained from 0.050 to 0.350 mg/m 3. The dead space for the vapor corresponded to the physiological dead space, indicating that all mercury vapor that reached the alveoli was absorbed (Nielsen-Kudsk, 1965, 1969; Teisinger and FiserovaBergerova, 1965). A later study in five volunteers confirmed that retention occurs almost entirely in the alveoli. Overall retention was 74%. Examination of the subjects in a whole-body counter yielded average half-lives for mercury as follows: lung, 1.7 days; head, 21 days; kidney region, 64 days; and standard chair position, 58 days (Hursh et aI., 1976). Use Experience Acute poisoning by metallic mercury is rare. Inhalation of vapor by laboratory workers in a closed space led
to bronchial irritation, violent coughing, and severe headache, followed in a few hours by fever, dyspnea, and nausea. Stomatitis appeared in 3 days. Dyspnea and fatigue lasted several months (Christensen et aI., 1937). Both renal and nervous involvement are unusual in acute poisoning (Browning, 1969). Neal et al. (1937, 1941a) gave a very thorough account of chronic mercurialism in the fur-cutting and felt-hat industries. Although mercuric nitrate was the material used to treat fur from which felt was made, the mercury was gradually released from the fur and felt in the form of metallic mercury vapor. Thus the workers had a mixed exposure to dust of mercury compounds (especially the nitrate) and to vapor ofthe element. The clinical picture of poisoning was similar to that observed among smaller groups whose occupational exposure involved metallic mercury only (Browning, 1969; West and Lim, 1968). The incidence of chronic mercurialism is roughly proportional to the concentration of mercury in the air at concentrations of 0.1 mg/m3 and upward, and within each range of concentration the incidence increases with duration of employment (Neal et aI., 1941a). Patients often gave a history of gastrointestinal disturbances and sore mouth. Salivation, gingivitis, and the loss of teeth were common. However, the major signs included fine tremors and psychic disturbances often called erethism. The tremor usually was noticed first in the hands and later ia the tongue, eyelids, face, and legs. It became worse during intentional movement or after the slightest emotional strain. The psychic disturbance took the form of irritability, excitability, timidity, irascibility, and difficulty in getting along with people. These signs were noted more frequently by an examiner than they were reported by the patient. They tended to be replaced by depression or despondency. Headache, drowsiness, insomnia, weakness, slurred speech, excessive sweating, dermographia, and vasomotor disorders all pointed to disorder of some part of the nervous system. Ataxia was seen occasionally and hemiplegia more rarely. In advanced cases there might be hallucinations, loss of memory, and intellectual deterioration. Death might follow symptoms resembling cerebral pachymeningitis (Neal et al., 1941a). More recent studies have indicated that a number of nonspecific symptoms, such as insomnia, introversion, and anxiety, may be produced at concentrations somewhat below 0.1 mg/m 3 (for review, see ATSDR, 1999). Laboratory Findings Effects on renal function may be severe (the nephrotic syndrome) at high exposure (Friberg et al., 1953; Kazantzis et aI., 1962). Air concentrations in the occupational setting in the range of 0.05-0.1 mg Hg/m3 have been associated with increased urinary excretion of biochemical biomarkers of kidney cellular changes such as renal enzymes and antigens without compromised kidney function (ATSDR, 1999). Treatment of Poisoning Conflicting reports have been published about the value of both BAL and calcium disodium EDTA for treating chronic poisoning by mercury vapor (Browning, 1969). Sunderman (1978) found that BAL (2,3-dimercapto-
61.6 Mercury I-propanol) was more effective in alleviating the signs and symptoms of mercury vapor poisoning than D-penicillamine or sodium diethyldithiocarbamate. However, treatment with complexing agents may not be necessary as the prognosis is good with virtually complete regression if exposure ceases (Berlin, 1986). 61.6.2 MERCURIC CHLORIDE 61.6.2.1 Identity, Properties, and Uses Chemical Name Structure
Mercuric chloride.
HgCh.
Synonyms Mercuric chloride also is known as corrosive sublimate, mercury bichloride, and mercury perch10ride. The CAS registry no. is 7487-94-7. Physical and Chemical Properties Mercuric chloride has the empirical formula ChHg and a molecular weight of 271.S2. It is an odorless, white, crystalline powder with a metallic taste. At 2SoC, it has a density of S.4.1t melts at 277°C, and sublimes unchanged at about 300°C. Its vapor pressure is 1.4 x 10-4 torr at 3SoC. Mercuric chloride is soluble in ethanol, methanol, acetone, ethyl acetate, and diethyl ester. It is slightly soluble in acetic acid, pyridine, and carbon disulfide. Its solubility in water at 20°C is 69 g/l. Mercuric chloride is unstable in the presence of alkalies and is decomposed to metallic mercury by sunlight in the presence of organic matter. It is readily reduced to mercurous chloride and elemental mercury. History, Formulations, and Uses Mercuric chloride first was used for crop protection in 1891. It formerly was used as a fungicide in soil application to control potato scab and clubroot of bras sic as and as an insecticide against root maggots of crucifers. It now is used in seed treatment of potatoes and as a wood preservative. Formulations include wettable powders, dusts, and solutions for injection into tree trunks. 61.6.2.2 Toxicity to Laboratory Animals Basic Findings Oral doses of mercuric chloride in the range IS-60 mg Hg/kg/day given daily for 2 weeks or less to rats or mice produce death through kidney failure. Daily oral exposure from 2 to 26 weeks at 1-28 mg Hg/kg produces mainly kidney damage and weight loss (for a review, see ATSDR, 1999). Exposure of rats for 2 years, S days/week at 1.9 mg Hg/kg produced pathological changes in kidney morphology (NTP, 1993). Absorption, Distribution, Metabolism, and Excretion Gastrointestinal absorption of oral doses of mercury chloride in rats or mice ranges from 1 to 40% of the dose. Several factors influence the degree of absorption including diet and age of the animal. Suckling animals have higher absorption rates. The
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kidneys are the main site of deposition with considerably less mercury crossing the blood-brain barrier or placenta compared to inhaled mercury vapor or methyl mercury compounds. Inorganic mercury is excreted in both urine and feces. Studies on dogs revealed the inorganic mercury in urine originates from inorganic mercury in renal tissues (Hursh et al., 1985). Inorganic mercury is readily eliminated in milk according to experiments in guinea pigs (Yoshida et aI., 1992). Effect on Organs and Tissues Mercuric ion in sufficient concentration causes a reversible precipitation of protein. Concentrations of the ion high enough to cause this effect may be reached as a result of direct contact with the skin, the mucous membranes of the eye, or the mucosa of the gastrointestinal tract. However, the mercuric ion produces systemic toxicity at much lower concentrations through inhibition of many enzymes, especially those containing the SH group. After a large dose, death is due to shock and circulatory collapse; arrhythmias may be present, and ventricular fibrillation may be the terminal event. After a somewhat smaller dose, death is due to renal damage leading to anuria. The necrosis of the kidney tubule cells of rats following subcutaneous injection of mercuric chloride at dosage of 0.1-4.0 mg/kg/day for several days was reflected not only by characteristic histology but by the shedding of cells in the urine and by an increase of urinary glutamic-oxaloacetic transaminase activity in the intact animal. The urinary output of cells reached a peak in a few days and then declined in spite of continued administration of the compound. The peak came earlier at higher dosage levels. A transient increase in transaminase corresponded with the peak of renal cell exfoliation, but the enzyme activity was a less sensitive test. Histological examination showed that the decline in cellular exfoliation was associated with tubular regeneration. The new cells were relatively immune to the effects of mercuric ion, but the basis of the tolerance is unknown (Prescott and Ansari, 1969). Mercuric chloride at a concentration of 3.S x 10-5 M acted as a mitogen in cultures of lymphoid cells from rats, guinea pigs, and rabbits (Pauly et al., 1969). 61.6.2.3 Toxicity to Humans Accidental and Intentional Poisoning Ingestion of mercuric chloride leads immediately to a burning metallic taste, thirst, and soreness of the throat followed soon by salivation, severe gastric pain, bloody diarrhea, and often vomiting. If the patient does not die promptly in shock following a dose of about 1 g, the onset is delayed for a few hours. Damage to the capillaries leads to stomatitis, loosening of the teeth, progressive renal damage, and continuing diarrhea. If the blood pressure is maintained, there may be an initial diuresis resulting from suppression of the renal tubular resorption. Progressive injury leads to a diminished flow of urine and eventually to anuria. Inorganic mercury is excreted so efficiently that persons who survive one or a few doses usually recover completely. Laboratory Findings
See Section 61.6.3.4.
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Pathology Local corrosive effects may be the only indication of injury of persons who die rapidly of shock following a large dose of mercuric chloride. In case of longer survival, renal tubular necrosis is the major finding.
Table 61.1 Saturated Vapor Concentration of Mercury and Certain Groups of Its Compounds at 20°C Concentrations Group
Treatment of Poisoning Both BAL and penicillamine are effective in treating poisoning by inorganic mercury compounds (Hayes, 1975). 61.6.3 ORGANIC MERCURY COMPOUNDS: SIMILARITIES AND DISTINCTIONS
(mg/m3)
Metallic mercury Dialkyl compounds
14 10,000
Methyl compounds
0.3-94
Ethyl compounds
0.05-9.0
Phenyl compounds
0.001-D.017
Methoxyethyl compounds
0.002-2.6
61.6.3.1 Uses The organic mercury compounds are used as seed dressings for the prevention of seed-borne disease of grains, vegetables, cotton, peanuts, soy beans, sugar beets, and ornamentals. They may be used for the control of fungus diseases of turf, fruits, cereals, and vegetables but not under conditions that will leave any measurable residue in the food of humans or animals. At least in other countries, organic mercury compounds have been used for the preservation of wood (Ahlmark, 1948) and in the paper, plastics, and fabric industries (Lundgren and Swensson, 1960). About 1915, phenylmercury fungicides began to replace mercuric chloride, which had been in use for seed treatment before the turn of the century. Alkyloxyalkyl mercury compounds were introduced next, and ethyl mercury was tested as early as 1929. Alkyl mercury fungicides were used extensively in the 1940s and thereafter. The following discussion is necessarily incomplete. Additional information may be found in a book by Bidstrup (1964) and in a health criteria document issued by the WHO (1976) The first of these references is notable for a tabulated description of cases caused by different forms of mercury.
61.6.3.2 Compounds and Their Characteristics Mercury is bivalent in the organic compounds used as fungicides. Most of these compounds fall into three major classes depending on whether the organomercury cation involves an (a) alkyl (e.g., methyl or ethyl), (b) phenyl, or (c) alkoxyalkyl (e.g., methoxyethyl) group. Compounds of each group may involve a wide range of inorganic or organic anions, including chloride, bromide, iodide, nitrate, hydroxide, acetate, dicyandiamide, toluene-sulfonate, benzoate, methanedinaphthyldisulfonate, and others. There are other bivalent organomercury compounds not used as fungicides, for example, dialkyl mercury compounds, certain local antiseptics, and the mercurial difuretics, but they are not discussed further here. Vapor pressure is an important factor determining the availability of organic mercury compounds for absorption. Table 61.1, derived in part from a detailed review by Swensson and Ulfvarson (1963), shows the concentrations of mercury and some of its compounds in saturated atmospheres. This helps to explain the observed extreme hazard of methyl mercury compounds in workplaces. Only a lack of use limits the hazards of
the very highly volatile dialkyl compounds. The high reported range of volatility of some alkoxyalkyl compounds depends at least in part on the fact that they often contain metallic mercury as a contaminant (Lindstrom, 1961). In fact, commercial formulations of all organic mercury compounds are likely to contain related compounds and metallic mercury resulting from the manufacturing process or from slow decompensation catalyzed by light.
61.6.3.3 Absorption, Distribution, Metabolism, and Excretion Absorption These compounds are absorbed slowly by the skin and more efficiently by the respiratory and gastrointestinal tracts. Some of the alkyl compounds are highly volatile thus increasing the hazard of their inhalation. There is no indication of any significant difference in the rate of absorption of different compounds at the same rate of dosage. Distribution and Storage Alkyl and aryl mercury compounds, like metallic mercury vapor, are transported mainly in association with the erythrocytes. This is in contrast to the inorganic mercury ion, which is bound mainly to plasma protein. Methoxyethyl compounds are evenly distributed between red cells and plasma. There are striking differences in the distribution and storage of different classes of organic mercury compounds, and this appears to be the basis for the differences in their toxic effects when given in repeated doses. The mercury content of different organs of rats following repeated doses of typical compounds is shown in Table 61.2. It may be seen that mercury from methoxyethyl mercuric hydroxide reaches about the same concentrations in the blood and vital organs as those from mercuric nitrate. However, the distribution of mercury derived from methylmercuric hydroxide is entirely different, being about 15 times as high in the brain, 100 times as high in the blood, but only one-tenth as high in the kidney. The behavior of mercury derived from phenylmercuric hydroxide is intermediate between that of alkyl mercury and inorganic mercury but much more nearly like the latter (Ulfvarson, 1962). Similar results were obtained in chickens (Swensson and Ulfvarson, 1968a).
61.6 Mercury Table 61.2 Average Mercury Content of Fresh Tissue of Ratsa,b Blood
Liver
Kidney
Brain
Compound
(ppm)
(ppm)
(ppm)
(ppm)
Hg( N03)z
0.028
0.372
20.1
0.024
Methyl Hg OH
3.04
0.676
2.9
0.155
Phenyl Hg OH
0.313
0.566
26.2
0.008
Methoxyethyl Hg OH
0.033
0.248
26.9
0.009
aFrom data of U1fvarson (1962).
bTreatment of rats is 0.1 mg Hg/day every other day for 2 weeks.
Metabolism Differences in distribution and storage of organic mercury pesticides are generally assumed to be due to differences in mobility across cell membranes, affinities for tissue ligands, and rates of metabolism to inorganic mercury. The latter is illustrated by the greater stability of alkyl compounds as compared to the more rapid degradation of methoxyethyl and phenol compounds (Daniel et aI., 1971; Gage, 1964; Norseth and Clarkson, 1970). Methyl mercury is readily transported across the placenta in rodents (Childs, 1973). In humans, cord blood levels closely parallel and are about 20% higher than the corresponding levels in maternal blood (Bakir et al., 1973; Skerfving, 1974). Species vary considerably in the relative distribution of alkyl mercury between brain and blood (Berlin et aI., 1969b). Table 61.2 shows that somewhat more mercury was stored in the liver and kidney when a phenyl compound was injected for 2 weeks than when inorganic mercury was injected at an equivalent rate. The same thing was seen in much exaggerated form when phenylmercuric acetate and mercuric acetate were fed at equivalent doses for a year or more; there was 10 to 20 or more times greater storage of the phenyl compound depending on dosage. The difference was attributed at least in part to greater absorption of the phenyl compound (Fitzhugh et al., 1950). Excretion Following injection of equivalent amounts of mercury, urinary excretion of phenyl compounds is almost twice that of inorganic mercury and more than 10 times greater than that of methyl mercury (Friberg, 1959; Swensson and Ulfvarson, 1959). That is, the same level of excretion is achieved only at high blood and tissue levels of alkyl compounds. In rats, following a single dose, the excretion of various forms of mercury was more rapid at first and then slower. For both mercuric nitrate and phenylmercuric hydroxide the "halflives" were 5 days during the first 9 days and 10 days during the subsequent 30 days. The half-lives were 16 and 26 days for methylmercuric hydroxide during the same intervals (Swensson and Ulfvarson, 1968b). Half-lives may be different in different species. The half-life of methyl mercury in the mouse is about 7 days compared to an average 70 days in humans and may be as high as 700 days in certain marine mammals (for review, see Clarkson, 1972). The half-life of methyl mercury in blood in humans averages about 50 days, whereas that of inorganic mercury is about
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30 days (for review, see Clarkson et aI., 1988). Information on phenyl mercury compounds is not available. Although dimethyl mercury apparently has not been used as a pesticide, it is of interest because of its formation in aquatic environments. Most ofthe dimethyl mercury injected into mice was rapidly exhaled. The remaining part was metabolized to methyl mercury, on which the entire toxic effect depended (Oestlund, 1969). 61.6.3.4 Laboratory Findings in Humans The record of mercury levels in human tissues and excreta is complicated by the fact that early analytical methods lacked the necessary sensitivity and made no distinction between inorganic and organic mercury. In fact, this distinction has not yet been made for many different occupational situations or for a reasonably complete spectrum of dietary patterns. As a result, the mercury concentrations in human tissues discussed next are expressed in terms of total mercury. Blood Goldwater and Hoover (1967) have reported blood levels of mercury in 812 people from 15 countries with no known special exposure to mercury. Approximately 75% of the samples were less than 0.5 ).l.g/l and about 90% were less than 20 ).l.g/l. Mercury in blood is greatly influenced by dietary fish consumption in otherwise nonexposed people. Methyl mercury in fish can make a substantial and dominant contribution to blood levels. Certain populations that depend on fish for their main source of protein can develop levels in excess of 299 ).l.g/l (for review, see WHO, 1976). In groups occupationally exposed to mercury vapor, blood levels are proportional to time-weighted average air concentrations for long-term exposures (1 year or more) and on a group basis (Smith et aI., 1970). In general, time-weighted air concentrations in the range 50-100).l.g Hg/m3_values corresponding to maximum allowable limits--correspond to blood levels in the range 35-70 ).l.g/l. Values of 500 to 1000 or 1300 ).l.g/l have been reported in patients (Berlin et aI., 1969c; Birke et aI., 1967). Urine A study of urine samples from 1107 people in 15 different countries (Goldwater and Hoover, 1967) revealed that approximately 80% of the samples had mercury concentrations below 0.5 ).l.g/l (the detection limit of the analytical method), 90% were below 10 ).l.g/l, and 95% were below 20 ).l.g. Studies in other populations are consistent with these figures (Buchet et al., 1980; Gotelli et al., 1985). Smith et al. (1970) have reported that urine concentrations of mercury are proportional to the time-weighted average air concentrations of mercury vapor in groups of exposed workers. In general, average air concentrations in the range 50100 ).l.g Hg/m3 (the maximum values expected in industrial atmosphere) would correspond to urine concentrations in the range 100-200 ).l.g/l (WHO, 1980). Exposure to alkyl mercury compounds results in elevated urinary concentrations of mercury (Bakir et al., 1973) but the
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icrease is small and urinary concentrations are not a viable biological indicator (for discussion, see Clarkson et al., 1988). Urinary concentrations may be markedly increased after occupational (Goldwater, 1973) or accidental (Gotelli et al., 1985) exposure to aryl mercury compounds. However, the quantitative relations between exposure and urinary excretion have not been described. Hair In a study of 559 samples in 13 countries, Airey (1983) noted a correlation between hair concentrations and average fish consumption. Thus he found the mean hair concentration to be 1.4 ppm in individuals consuming fish once or less per month, 1.9 ppm for consumption once every 2 weeks, 2.5 ppm for consumption once a week, and 11.6 ppm for consumption once or more per day. This correlation is probably due to intake of methyl mercury in fish. A close correlation has been noted in both population (Amin-Zaki et aI., 1976; Birke et al., 1972; Phelps et al., 1980) and experimental (Hislop et aI., 1963; Kershaw et al., 1980) studies between mercury concentration in hair next to the scalp and the simultaneous blood concentration after exposure to methyl mercury. The average ratio of hair to blood concentrations is about 250 : 1. Studies in volunteers who received measured amounts of methyl mercury in fish indicate that there is a 20-day delay between the blood concentration and the appearance of mercury in hair next to the scalp (Hislop et aI., 1963). Methyl mercury appears to be incorporated into the hair during its formation and remains stable. Because there is little variation in the rate of growth of head hair (1.15 cm/month) it is possible to determine the sequence of past exposures of persons to methyl mercury (AI-Shahristani and Shihab, 1976; Amin-Zaki et aI., 1976). Hair concentrations after exposure to other forms of mercury (mercury vapor, inorganic and aryl compounds) have not been reported in any detail. In the case of occupational exposure, the probability of external contamination should be considered. Factors affecting hair concentration and the use of washing procedures have been reviewed by Airey (1983). Tissues Information on tissue levels of mercury are generally restricted to the chief target tissues such as brain, kidney, and liver. Measurements have also been reported on placental tissues from the viewpoint of biological monitoring. Determinations on tissue in the early part of the twentieth century are not included because of the possibility of contamination with mercury added to tissue fixatives. Studies of the outbreak of methyl mercury poisoning in Minamata, Japan, in the 1950s provided data on control (nonexposed) groups. Brain levels were reported to be less than 0.1 ppm (for review, see Takeuchi and Eto, 1977). Five autopsy cases in Yugoslavia (Kosta et al., 1975) revealed values in the brain in the range 0.001-0.007 ppm. Mottet and Brody (1974) reported levels in 60 hospital autopsy cases in the United States with no known exposure to mercury in the range 0.006-0.965 for the cerebellum and 0.008-0.470 for the cerebrum. A more recent report (Nylander et al., 1987) suggests that brain levels
in otherwise non exposed individuals increased according to the number of mercury amalgam tooth fillings. Measurements in the occipital lobe cortex in 34 individuals yielded an average value of 0.011 ppm with a range of 0.002 to 0.029 ppm. Linear regression analysis indicated that the mercury concentrations increased with the number of amalgam surfaces in the teeth. Kidney values for nonexposed people are reported to be in the range 0.18-2.6 ppm in Japan (Takeuchi and Eto, 1977), 0.01-0.37 ppm in Yugoslavia (Kosta et aI., 1975), 0.006-0.4 ppm in the United States (Mottet and Brody, 1974), and average values of 0.433 ppm in seven people with mercury amalgams and 0.049 ppm in five people with no mercury amalgams in Sweden (Nylander et aI., 1987). Liver levels in nonexposed people are reported to be 0.161.3 ppm in Japan (Tsubaki and Irukayama, 1977), 0.01-0.05 ppm in Yugoslavia (Kosta et al., 1975), and 0.008-1.43 ppm in the United States (Mottet and Brody, 1974). Fatal cases of methyl mercury poisoning in Minamata had brain levels in the range 2.6-25 ppm, liver levels in the range 22-70 ppm, and kidney levels in the range 21-144 ppm. These individuals had died about 3 months after the end of exposure. Liver levels in the autopsy cases after an outbreak of methyl mercury in Iraq (Magos et aI., 1976) were reported to be in the range 1.4-76 ppm (Takeuchi and Eto, 1977). Little information is available on human tissue levels after exposure to other forms of mercury. Kosta et al. (1975) noted elevated levels (sometimes a thousandfold higher than in controIs) in thyroid, pituitary, kidney, and liver in mercury miners who had died many years after retirement. The average level of mercury in 38 placentas from residents in Iowa (United States) was 2.3 ppm. In another study, in Ohio, 29 placentas yielded an average of 6.7 ppm, of which 5.3 ppm was in the inorganic form. Other Findings A close correlation between selenium and mercury concentrations in a variety of autopsy tissues from retired miners and other residents of the mining village of Idria, Yugoslavia, was reported by Kosta et al. (1975). The atomic ratio of selenium to mercury was almost exactly 1: lover a wide range of concentrations. These observations, along with experimental data on animals (for a more recent report, see Magos et aI., 1984), suggest that inorganic mercury forms a complex with selenium that may persist for long periods in human tissues. 61.6.3.5 Treatment of Poisoning in Animals and Humans Treatment in Animals The literature on the treatment of poisoning by mercury in animals and people was already very extensive in 1967 when Swensson and Ulfvarson reviewed it and added their own thorough studies in animals. On the basis of information on both animals and humans, they emphasized that therapeutic effect depends on the form of mercury causing poisoning and also on whether treatment begins early in the course of acute poisoning or after chronic poisoning already is established. Of compounds that they tested and that
61.6 Mercury have ever been put to clinical use, BAL and D-penicillamine had a therapeutic effect in acute poisoning by inorganic mercury salts. BAL was also useful in acute poisoning by phenyl mercury. No useful treatment was found for acute poisoning by methyl mercury. Whereas penicillamine may have conferred some benefit in acute poisoning by methoxyethyl mercury, BAL did great harm: all animals died in convulsions following the second dose. AH of the chelating agents had some influence on the distribution of mercury in different organs and on its excretion. However, there was no definite connection between the lifesaving effect of an antidote and its effect on distribution or excretion of mercury. An increase in excretion of mercury was not a precondition for lifesaving effect. On the other hand, each chelating agent not only has its own toxicity but also may increase the toxicity of one or more forms of mercury, whether by increasing its concentration in the brain or kidney or in some other way. Zimmer and Carter (1979) confirmed that BAL is useless in methyl mercury poisoning but, contrary to some earlier results, found that D-penicillamine enhanced weight gain and prevented further development of neurotoxic signs in rats poisoned by methyl mercury chloride. In some instances, studies concerned directly with therapy have been confirmed by studies on distribution. For example Berlin and Ullberg (1963) showed that BAL greatly increased the uptake of phenyl mercury or methyl mercury by the brain. Zimmer and Carter (1979) reported a similar result with methyl mercury. More recent work has served to confirm most of the findings, but further progress has been limited. To be sure, an almost endless array of new chelating agents have been synthesized and tested for ability to increase the excretion of mercury, to cause a redistribution of it, and/or to increase the survival of experimental animals. There has been progress in understanding the complications of using systemically acting chelating agents. Following intravenous injection of mercuric chloride, Gabard (1976) tabulated not only the urinary and fecal excretion of 203Hg but also its distribution in red cells, plasma, liver, kidneys, brain, femur, muscle, spleen, and intestine of different groups of rats following treatment with one of 15 chelating agents. It was concluded that the only agent that had a truly favorable effect was sodium 2,3-dimercaptopropane-I-sulfonate (DMPS). Whether this would be true in humans as well as the rat is, of course, unproved. It seems certain, however, that the general problem of complex effects on distribution of metals following systemic chelation applies to all species. The possible danger that a systemic chelating agent will increase the toxicity of mercury may be avoided completely by using a non absorbable chelating agent that serves to reduce the absorption or reabsorption of mercury from the intestine and thus to promote feca1 excretion. One of these mercury-binding polymers (MBP) decreased the half-life of intraperitoneally administered methyl mercury in mice from 10.0 to 4.5 days and increased the LD 50 from 10.0 to 12.7 or 13.4 mg/kg, depending on the exact method of administration of the antidote (Harbison et al., 1977).
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A possible new approach to the treatment of poisoning by mercury, especially methyl mercury, was suggested by Kostyniak et al. (1975). In an in vitro study, they used a soluble and dialyzable chelating agent to free methyl mercury from protein binding and then removed the complex by passing the blood through a semipermeable dialysis tube. Using cysteine at a concentration of 10 mM in whole blood, up to 44% of methyl mercury and up to 94% of the cysteine were removed by one pass through die dialyzer. Treatment in Humans Although BAL has been widely and successfully used in the treatment of inorganic mercury poisoning, the treatment of methyl mercury poisoning with chelating agents has not been successful (Nierenberg et al., 1998). The reason is that, by the time symptoms and signs of poisoning appear, irreversible damage has been inflicted on the target tissue, the brain. There have not been enough cases of poisoning by alkoxy or phenyl mercury compounds to reach any conclusions regarding the effectiveness of chelation therapy. Animal experiments suggest that BAL may be dangerous to use in the case of alkoxyalkyl mercury poisoning (for a review, see Swensson and Ulfvarson, 1967). A neuromuscular disorder similar to myasthenia gravis and responsive to neostigmine was uncovered in the course of electrophysiological testing of Iraqi patients poisoned by methyl mercury. Subsequently, neostigmine therapy (gradually increased to 15-22.5 mg plus 2-3 mg of atropine sulfate intramuscularly daily in split doses) produced a remarkable clinical improvement. Substitution of a placebo resulted in substantial loss of strength that was restored when therapy was resumed (Rustam et aI., 1975). 61.6.4 ALKYL MERCURY COMPOUNDS 61.6.4.1 Identity, Properties, and Uses Compounds and Synonyms Of the entire range of alkyl mercury compounds, ethyl and methyl compounds have been used as pesticides. Methyl mercury was available in the form of several salt-each sold under one or more proprietary names, including the bis-methylmercuric sulfate (Cerewet®), the cyanoguanidine or dicyandiamide (Agrosol®, Morsodren®, Panogen®, Panospray®), the nitrile (Chipcote®), and the propionate (Metasol Mp®). Ethyl mercury also was available in the form of several salts, including the chloride (Ceresan®, Granosan®), the phosphate (Lignasan®, New Improved Ceresan®, New Improved Granosan®), the p-toluenesulfonanilide (Ceresan M®, Granosan M®), the 1,2,3,6tetrahydro-3, 6-endomethano-3,4,5,6,7,7 -hexachlorophthalimide (50-CS46, Emmi®, PX-332), and the thiosalicylate (Elcide®, Merfamin®, Mertorgan®, Merzonin®). Ethylmercuricthiosalicylate also was known by the nonproprietary names mercurothiolate, thimerosal, and thiomersalate. Properties
See Section 61.6.3.2.
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Uses Most of the individual compounds are used as fungicides in treating seeds, especially those of cereals, sorghum, sugar beets, cotton, and flax. Used in this way, they control diseases caused by seed-borne infection and protect germinating seedlings from soil-borne pathogens (WHO, 1974). 61.6.4.2 Toxicity to Laboratory Animals Basic Findings The acute oral LD 50 value for representative compounds in rats is approximately 30 mg/kg (Lehman, 1951, 1952). The total dose of an alkyl mercury compound necessary to produce chronic poisoning in cats is about the same, that is, 6-24 mg/kg expressed as mercury (Kurland et aI., 1960). Rats tolerated 0.5 mg (about 2.8 mg/kg/day) of ethyl mercury chloride daily for 150 days without sign of poisoning, but twice that dosage produced typical signs in 34 to 84 days (Akitake, 1968). Methyl mercury affects mainly the central nervous system. However, the effects differ markedly for adult versus prenatal exposures, so the two situations will be discussed separately. Effects on the Mature Nervous System Methyl mercury produces focal damage to specific areas of the brain-the neuronal cells of the visual cortex, particularly neurons situated in the deep sulci, appear to be most susceptible to damage (for a review, see Berlin, 1986). Effects most similar to those observed in humans are seen in nonhuman primates and rats. Inhibition of protein synthesis is one of the earliest biochemical effects that precede the appearance of overt signs of intoxication in animals. Effects on the Developing Nervous System Spyker et al. (1972) treated mice on day 7 and day 9 of pregnancy with a single (8 mg/kg) dose of methyl mercury. The young did not differ from the controls in size, weight, and appearance. However, when tested at 30 days of age, their behavior was abnormal in an open field test and in swimming. The brain showed abnormalities in the Purkinje and granule cells of the cerebellum (Chang et aI., 1977). More recently, Sager et al. (1984) reported that after a single (8 or 4 mg/kg) dose of methyl mercury to neonatal mice, the cell division of cerebellar granule cells was drastically reduced. The lower dose produced effects in male mice but not in females. It is thought that methyl mercury produces these effects by damaging microtubules, essential for cell division. For further discussion, see Sager and Matheson (1988). 61.6.4.3 Toxicity to Humans Alkyl mercury fungicides commonly have been the cause of occupational poisoning even in developed countries, especially when anything but the best equipment was used for applying them to seed. Furthermore, their application to seed created the possibility that they would be eaten in quantity, especially by peasants who were hungry and did not know about either the delayed onset of symptoms or the impossibility of adequately decontaminating the seed (see Section 7.1.2.4 of Hayes, 1975). Finally, outbreaks of poisoning clinically similar to that caused
by eating treated seed resulted from eating animals that had consumed contaminated feed or from eating fish that had absorbed alkyl mercury from industrial wastes. Symptomatology Acute poisoning by organic mercury has been reported infrequently in humans, although cases of such poisoning by methyl (Swensson, 1952) and other alkyl compounds (Lundgren and Swensson, 1949; Veichenblau, 1932) have occurred. There have been many cases of chronic poisoning involving organic mercury. The classical description of poisoning by an alkyl mercury compound is that of Edwards (1865). The patient may complain of headache; paresthesia of the tongue, lips, fingers, and toes; and other nonspecific dysfunction. In mild cases, the symptoms do not develop beyond this point, and in such instances they usually disappear gradually. Some but not all workers equally exposed to alkyl mercury compounds complain of a metallic taste in the mouth and slight gastrointestinal disturbances, such as excessive flatus and diarrhea (Bloom et aI., 1955; Ritter and Nussbaum, 1945). However, the acute symptoms associated with irritation of the gastrointestinal system and renal failure caused by inorganic mercury compounds are seldom observed in poisoning by alkyl mercury compounds and then almost exclusively in acute poisoning. Even the mild digestive disturbances and sore mouth seen in moderate, chronic, occupational poisoning by inorganic mercury are relatively rare. Instead, the nervous symptoms appear first, sometimes after relatively slight exposure and after weeks or months of latency. Diagnosis is complicated not only by the latency but also by the insidiousness of the onset. The patient may be unable to state with any certainty when he or she first noticed important symptoms. Early signs of more severe poisoning include fine tremors of the extended hands, loss of side vision, and slight loss of coordination, especially with the eyes closed as in the finger-to-nose test. Incordination is especially noticeable in speech, writing, and gait. Incoordination may progress to the point of inability to stand or to carry out other voluntary movements. Occasionally there is muscle atrophy and flexure contractures. In other cases, there are generalized myoclonic movements. There may be difficulty in understanding ordinary speech, although hearing and the understanding of slow deliberate speech often remain unaffected. Irritability and bad temper are frequently present and may progress to mania. Occasionally the mental picture deteriorates to stupor or coma (Ahlborg and Ahlmark, 1949; Ahlmark, 1948; Hemer, 1945; Hunter et aI., 1940). Especially in children, mental retardation may be added to the symptoms of poisoning already mentioned (Engleson and Hemer, 1952; Kurland et al., 1960). Patients frequently become gradually much worse after their illness is recognized and exposure is stopped. The alkyl mercury compounds are strong irritants of the skin and may cause blisters or other dermatitis with or without associated systemic illness (Hunter et aI., 1940; Vintinner, 1940). Study of 43 cases in the 1972 outbreak in Iraq showed that a few entered hospital with complaints that suggested psy-
61.6 Mercury chiatric disturbances, and over half of them were consistently depressed. Blood mercury levels were consistently higher in depressed than in nondepressed patients (Maghazaji, 1974). Signs and symptoms of poisoning in children are similar to those in adults (Nagi and Yassin, 1974). Duration of Illness The duration of illness in fatal cases ranged from about a month to 15 years (Hunter et al., 1940). Intercurrent infection, aspiration pneumonia, and inanition are the immediate causes of death in protracted cases (Kurland et al., 1960). Ten years after poisoning, neurological disorders persisted with little or no improvement among 26 victims of Minamata disease (Tokutomi et al., 1961). Symptoms often persist for years even in mild poisoning (Tsubaki, 1968). Recovery from poisoning by alkyl mercury is so slow and the effects of chelation therapy are so unimpressive that it was thought some years ago that no recovery is possible. Observations repeated at long intervals have indicated that some improvement does occur except perhaps in very severe cases. Twelve patients who were poisoned by ethyl mercury during the 1960 outbreak in Iraq were reexamined in 1973 and showed considerable improvement (AI-Damluji, 1976a). The most severely poisoned patients who survived the 1971-1972 outbreak caused by methyl mercury improved slowly, although ataxia, diminution of visual fields and visual acuity, and paresthesias were still present 2 years later. During the same period, most patients originally graded mild or moderate lost their symptoms completely (AI-Damluji, 1976b). Some children who had suffered mild or moderate poisoning by methyl mercury recovered completely in the course of 2 years. Over half of the children who had suffered severe poisoning remained physically and mentally incapacitated. The degree of clinical progress shown by the children in Iraq was better than that of some other groups, but the difference may have depended on the prompt termination of exposure rather than on the age of the patients (Amin-Zaki et al., 1978). Perinatal Poisoning Infants exposed to methyl mercury in utero may be severely and permanently injured even though their mothers remain asymptomatic. On the other hand, infants who received no mercury in utero but received it in their mother's milk may escape without clinical signs. Such cases were observed in association with the outbreak of methyl mercury poisoning in Iraq during the winter and spring of 1971 to 1972. Four infants had blood levels above 1.0 ppm, and one had levels above 1.5 ppm without sign of injury, even though 0.2 ppm is the minimal toxic level for adults (Amin-Zaki et al., 1974a). In the Minamata area during the period 1955-1959,22 infants of a total of about 400 were born defective. Most of the mothers of these defective children did not have typical Minamata disease, but most of them experienced numbness during pregnancy. All ate fish frequently (Kutsuna, 1966). Experience in Iraq was quite different: in only one of 15 infant-mother pairs exposed to methyl mercury was the infant affected and the mother free of clinical signs (Amin-Zaki et al., 1974b).
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The cause of the difference is uncertain, but it might be associated with the stages of gestation at which exposure occurred. Whether teratogenesis occurs, as distinct from fetal intoxication, has been discussed (Berlin et al., 1969b) but is not yet clear. Follow-up studies of prenatally exposed infants in Iraq have revealed a milder form of affliction characterized by delayed achievement of developmental milestones, abnormal reflexes, and a history of seizures (Marsh et al., 1980, 1981). Possible Distinctions between Poisoning by Methyl Mercury and Ethyl Mercury There is some indication that, compared to methyl compounds, the illness produced by ethylmercuric compounds involves relatively greater injury to the gastrointestinal system (aphthous stomatitis, catarrhal gingivitis, nausea, liquid stool, pain, and laboratory evidence of liver disorder) and the cardiovascular and hematopoietic systems and less disorder of sensation and coordination (deafness, ataxia) (Alekseyeva and Mishin, 1971; Bogomaz, 1969; D'yachuk, 1972; Shustov and Tsyganova, 1970; WHO, 1962). The contrast between the two has been pointed out on the basis of outbreaks in Iraq, the one in 1960 caused by ethyl mercury and the one in 1972 caused by methyl mercury (AI-Damluji, 1976b). However, poisoning by ethyl mercury may be fatal (Gis' and Pozner, 1970; Ljubetskii et al., 1961; Mal'tsev, 1972), and those who survive may have residual nervous symptoms (Gis' and Pozner, 1970; Mal'tsev, 1972). A description of poisoning by ethyl mercury in children (Mal'tsev, 1972) makes it appear impossible to distinguish poisoning by ethyl and methyl mercury. At present it is unclear whether an important, clinical distinction is justified between poisoning by ethyl and methyl mercury either in adults or in children. Outbreaks from Eating Treated Grain Several outbreaks caused by eating seeds dressed with methyl or ethyl mercury are listed in Table 7.14 in the first edition of this Handbook and are discussed briefly in the associated text. Findings reached in study of these and some other outbreaks have been used as appropriate throughout this section and Section 61.6.3. However, no attempt is made here to describe the epidemics themselves. The largest one, in which 6148 patients were admitted to hospital and 452 died there, was described at length in the proceedings of a Conference on Intoxication Due to AlkylmercuryTreated Seed (WHO, 1976) and was summarized by Skerfving and Copplestone (1976). Outbreaks from Eating Poisoned Animals and Fish In December 1969, an 8-year-old girl developed ataxia, decreased vision, and depression of consciousness progressing to coma over a period of 3 weeks. Two weeks after she became sick, her 13-year-old brother developed a similar illness, which also progressed to coma in 2-3 weeks. At the end of December, their 20-year-old sister developed similar symptoms and became semicomatose. All were hospitalized and given supportive therapy. The use of chelating drugs was of very doubtful
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value. In May, two patients remained comatose, the other semicomatose (Likosky et al., 1970; Storrs et aI., 1970a, b). The 40-year-old mother in the family was pregnant at the time. She did not become ill, and she ate no more pork after the sixth month of gestation. However, her urinary mercury levels were elevated during the 7th and 8th months. A 3062-g male infant was delivered at term. He became dusky at 1 minute of life, and intermittent gross tremulous movements of the extremities persisted for several days. The cry was weak and high pitched, but otherwise the child seemed clinically normal. Urinary mercury values were markedly elevated. The electroencephalogram (EEG) and electromyogram (EMG) were normal 3 days after birth and remained so at 6 weeks. However, at 3 months of age the EEG was abnormal, and remained so. Clinical abnormality progressed even more rapidly with increased tone of the extremities at 6 weeks, generalized myoclonic jerks by 6 months, and hypertonicity, irritability, and nystagmoid eye movements without fixation of the eyes at 8 months. These effects were the result of mercury via the placenta, for the child was never breast-fed (Snyder, 1971, 1972). During the investigation it was learned that in October 1969, 14 of 17 hogs owned by the family became ill with blindness and a gait disturbance; 12 of these 14 died, and 2 became blind. In September, one hog had been butchered for family food and the meat was eaten by seven of nine family members from September through December. Further investigation revealed that, in August 1969, the father had obtained floor sweepings from a plant for treating seed grain with methyl mercury dicyandiamide and had included this grain in food for the hogs. High levels of mercury were found in the urine of the patients (0.16-0.21 ppm) and of other members of the family who remained well (0.06-0.20 ppm). The seed grain and pork contained almost equal concentrations of mercury (32.8 and 27.5 ppm, respectively) (Curley et aI., 1971; Storrs et aI., 1970a, b). Mass spectral analysis demonstrated the presence of methyl, ethyl, and probably methoxyethy1 mercury in what was left of the waste grain that had been fed to the hogs (Curley et aI., 1971). Thus, poisoning had been caused by a mixture of alkyl compounds. At least two epidemics were associated with the ingestion of seafood contaminated by industrial waste (Kurland et al., 1960; Tsubaki and Irukayama, 1977). The first of these outbreaks, which occurred at Minamata, Japan, alerted the medical community to the great danger, and Minamata disease became a synonym for chronic poisoning by organic mercury compounds. There is now considerable evidence that the compound in seafood responsible for poisoning is dimethyl mercuric sulfide (CH3-Hg-S-CH3) (Tokutomi et al., 1961). Dosage Response Apparently no useful information on dosage response to alkyl mercury compounds has come from accidents resulting from occupational exposure. Epidemiological studies of populations exposed to unusually high concentrations of methyl mercury in food have defined three categories differing in intensity and duration of exposure. People were poisoned
in Iraq when they consumed contaminated grain for an average of 1-2 months at an average mercury intake of 0.08 mg/kg/day and a maximum of about 0.2 mg/kg/day. People were poisoned in Japan when they ate contaminated fish for months or a few years at average and maximal mercury intakes of about 0.03 and 0.10 mg/kg/day. People in Sweden and various other parts of the world where fish have relatively high levels of methyl mercury were not affected by a maximal mercury intake of 0.005 mg/kg/day lasting many years or for life (WHO, 1976). A dosage-response analysis of both adult and prenatal exposures in Iraq revealed that the developing nervous system is more susceptible than the mature nervous system, probably by a factor of 3 (Clarkson et aI., 1981). Laboratory Findings Three men each swallowed 0.01 mg of 203Hg in the form of methylmercuric nitrate. Whole-body counting indicated that mercury accumulated mainly in the liver (50% of dose) and the head (10%). The main excretion route was the feces but urinary excretion (as a fraction of total excretion) increased with time up to 30 days after intake. The whole-body retention curves followed a single exponential curve with half-lives between 70 and 74 days (Aberg et aI., 1969). Similar data were obtained by Miettinen (1973). See also Section 61.6.3.4. Treatment See Section 61.6.3.6 and see Section 8.2 in the first edition of this Handbook. Pathology Extensive and relatively characteristic pathology has been reported in humans and experimental animals killed by alkyl mercury compounds. The most common findings are (a) bilateral cortical atrophy around the anterior end of the calcarine fissure with disappearances of the striation of Gennari (associated with constriction of the visual fields) and (b) gross atrophy of the folia in the depths of the sulci of the lateral lobes and the declive of the cerebellum involving the granule cell layer (associated with ataxia). The hypothalamus, midbrain, and basal ganglia may be involved. The changes in the brain involve gliosis as well as abnormality and loss of specific neurons. The bodies of the Purkinje cells are spared, although the axons are affected. The classical description of the pathology is that of Hunter and Russell (1954) based on a patient who had had industrial exposure to methyl mercury for only 4 months and who had been followed medically from the time of onset until his death from pneumonia and complications a little over 15 years later. The disease had progressed slowly at first, but by the end of the first year the patient was quite helpless although still alert and intelligent, and he remained so until his death. Reference to poisoning by alkyl mercury as the "Hunter-Russel syndrome" may be traced to this paper, which pointed out, among other things, the marked difference of pathology associated with alkyl mercury in humans and in experimental animals. Changes in peripheral nerves and the posterior columns have been reported in animals (Hunter et aI., 1940; Swensson, 1952), but Hunter was unable to find such changes in humans.
61.6 Mercury 61.6.5 ALKOXYALKYL MERCURY COMPOUNDS 61.6.5.1 Identity, Properties, and Uses Methoxyethyl mercury silicate frequently was formulated with other fungicides or insecticides for protection of cereal grain against both fungi and insects. Examples of proprietary names, which designate each total formulation and must not be confused with formulations of ethyl mercury involving the word "Ceresan®" include Ceresan-Aldrin® (with aldrin and HCB), Cere san Gamma M® (with anthraquinone, hexachlorobenzene, and lindane), Ceresan-Morkit® (with anthraquinone and hexachlorobenzene), and Ceresan Universal Trockenbeize® (methoxyethy 1 mercuric silicate with hexachlorobenzene). Methoxyethylmercuric chloride was known by several proprietary names, including Agallol®, Aretan®, and Ceresan Universal Nassbeize®. For a comparison of the vapor pressures of alkoxyalkyl compounds with those of other organic mercury compounds, see Section 61.6.3.2. 61.6.5.2 Toxicity to Laboratory Animals No information on the toxicity of alkoxyalkyl mercury compounds has been found. 61.6.5.3 Toxicity to Humans Poisoning by alkoxyalkyl mercury compounds usually begins with loss of appetite, flatulence, and diarrhea. The patient may complain of loss of weight, exhaustion, and headache. Albuminuria is a common finding and may be accompanied by generalized edema. Signs of injury of the central nervous system are less prominent than in poisoning by alkyl mercury compounds, but numbness of the fingers and toes and some degree of ataxia and weakness may occur (Bonnin, 1951; Derobert and Marcus, 1956; Wilkening and Litzner, 1952; Zeyer, 1952). Acrodynia was recognized in a 5-year-old boy exposed on a farm to a methy loxyethy1mercury seed disinfectant. The symptoms, which were typical, included an exanthem with scaling of the skin of the hands and feet, extreme apathy, anorexia, photophobia, sleeping during the day and inability to sleep at night, muscular weakness with tremor of the extremities, heavy sweating, hypertension, and tachycardia. The concentrations of mercury in the blood and urine were 0.04 and 0.055 ppm, respectively. The clinical picture improved within 4 weeks; the skin cleared in 2 months, and the patient gained weight; hypertension and tachycardia disappeared slowly (Prediger, 1976). It is unclear whether treatment with cortisone was helpful. A disease of children eventually named acrodynia because of the severe pain it caused in the extremities or "pink disease" because of the dusky pink color of these parts was described at least as early as 1828. There was no constructive clue to its cause until Warkany and Hubbard (1948) reported that mercury was present in the urine in 18 of 20 suspected cases, and the concentration exceeded 0.05 ppm in 15 of them. The two cases without mercury already had been recognized as atypical, but they indicated that an acrodynia-like condition may
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occur without demonstrable mercury. The paper also pointed out that the concentration of mercury in the urine of children treated with calomel often exceeds 0.05 ppm, and, therefore, some special susceptibility must be involved in the relatively few instances in which exposure to mercury is followed by acrodynia. A great many more cases confirmed the relationship (Warkany and Hubbard, 1951, 1953). When measures were taken by physicians, pharmaceutical companies, and regulatory agencies to stop the mercurial medication of infants and children, acrodynia became extremely rare (Warkany, 1966). Following communications with Warkany, Bivings and Lewis (1948) first used BAL to treat acrodynia. The value of such treatment remains unclear. The alkoxyalkyl mercury compounds, like the alkyl ones, can cause toxic dermatitis ranging from mild irritation to the production of slowly healing ulcers. Treatment is symptomatic. See Section 61.6.3.6 and see Section 8.2 of the first edition of this Handbook. 61.6.6 ARYL MERCURY COMPOUNDS 61.6.6.1 Identity, Properties, and Uses Compounds and Synonyms Phenyl mercury was available in the form of several salts each sold under one or more proprietary names, including the acetate (Gallotox®, Liquiphene®, Mersolite®, Nylmerate®, Phix®, Riogen®, Scutl®, Tag Fungicide®, Tag HL-331®, also known by the acronyms PMA, PMAC, PMAS), the nitrate (Merphenyl®, Phenmerzyl nitrate®, Phermerite®), the cyanamide (also known as barbak), the methylenebis(2-naphthyl-3-sulfonic) acid (Conotrane®, phenyl mercuric Fixtan®, Fibrotan®, Hydraphen®, Penotrane®, Septotan®, Versotrane®, also known as P.M.F. and as hydrargaphen), the borate (Famosept®, Imerfen®, Merfen®), the dimethyldithiocarbamate (Merbam®), the monoethanolammonium lactate (Puratized Agricultural Spray®, Puratized N5E®), the N-urea (Puratized Apple Spray®), the propionate (Metasol P-6®), the quinolinate (Metasol DPO®), the triethanolamine lactate (Dowicide la®, Dowicide 1®, Leytosan®, Natriphene®), and the hyrdoxychlorophenol (Semesan®, Uspulun®). Properties For a comparison of the vapor pressure of phenyl mercury with that of other organic mercury compounds see Section 61.6.3.2. Uses The individual compounds are used as fungicides for apples, pears, and other fruits, potatoes, tree wounds, bulbs, textiles, timber, leather, and wood pulp. A few are used as topical antiseptics and bactericides. The acetate is used as an herbicide for crabgrass as well as a fungicide. 61.6.6.2 Toxicity to Laboratory Animals Basic Findings The oral LD 50 of phenyl mercury acetate in rats was found to be 60 mg/kg (Piechocka, 1968). The same parameter for another phenyl mercury compound was 30
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mg/kg (Lehman, 1951, 1952). In mice, oral LD 50 values for phenylmercuric acetate and another phenylmercuric salt were both 70 mg/kg (Goldberg et al., 1950). Dietary mercury levels of 40 ppm (about 2 mg/kg/day) or lower in the form of either mercuric acetate or phenylmercuric acetate were tolerated by rats for 2 years without any change in rate of growth, mortality, or organ weights. A dietary level of 160 ppm (about 8 mg/kg/day) of mercuric acetate produced slight reduction in growth of male rats only, whereas the same rate of ingestion of phenylmercuric acetate interfered to a marked degree with the growth of both sexes and shortened their survival. The action of both compounds in the rat appeared to be on the kidney. A dietary level of 0.5 ppm of the phenyl compound produced detectable histological change in the kidney, but 10-20 times as much mercuric acetate was required to produce the same effect (Fitzhugh et al., 1950). Solecki et al. (1991) observed that rats exposed to phenylmercuric acetate at 5 mg per liter of drinking water for 2 years showed an acceleration in the development of age-associated chronic nephrosis compared to controls.
61.7 THALLIUM Thallium stands between mercury and lead in the periodic table. It has no known essential function in the body. Most cases of poisoning by compounds of thallium, including even the small number that are truly occupational, involve compounds intended for use as pesticides. Most of the cases involve accidental ingestion, suicide, or murder. Moeschlin (1965) stated that in Europe thallium had replaced arsenic as a homicidal poison. Thallium sulfate has been more widely used as a pesticide than any other compound of thallium. It has produced many cases of poisoning and serves as a good example of the toxicity of thallium generally. A very thorough review of the toxicology of thallium is that of Heyroth (1947). 61.7.1 THALLIUM SULFATE 61.7.1.1 Identity, Properties, and Uses
61.6.6.3 Toxicity to Humans
Chemical Name
Accidental and Intentional Poisoning Poisoning by aryl mercury compounds usually involves the blood, with symptoms of weakness secondary to anemia and infection secondary to leukopenia (Cotter, 1947). In some instances, a nonspecific neurasthenia occurs even though anemia is mild (Massmann, 1957) or absent (Swensson and Ulfvarson, 1963). In reporting an unsuccessful suicide with phenyl mercury, Ishida (1970) noted that the patient displayed no significant neurological manifestations.
Structure
Use Experience In many instances phenyl mercury fungicides have been used extensively without untoward effect even though workers may have increased levels of mercury in their hair (Kinoshita and Ogima, 1968; Tokutomi, 1968). Three cases of acrodynia were reported in several thousand infants exposed to a phenyl mercury acetate added to a diaper wash as a fungicide (Gotelli et al., 1985). A follow-up investigation of several hundred of these infants revealed urinary levels of mercury up to 500 J.tg/l. No clinical effects were found. Some infants had a mild enzymuria that subsequently disappeared. Laboratory Findings
Thallous sulfate.
TlzS04.
Synonyms Proprietary names for thallium sulfate include Bonide Antzix ant killer®, GTA ant bane®, GTA bait®, Magikil Jelly ant bait®, Martin's Rat-Stop®, Liquid Mission Brank antroach killer®, and Rex ant bait®. The CAS registry no. is 744618-6. Physical and Chemical Properties Thallium sulfate has the empirical formula 04STlz and a molecular weight of 504.85. It forms a colorless, odorless, dense powder of rhomboid prisms, with a density of 6.77 and a melting point of 632°C. Its vapor pressure is inappreciable. Its solubility in water at 20 0 e is 4.87%; at 100°C, 18.45%. History, Formulations, and Uses Thallium sulfate was introduced as a rodenticide in the United States during the 1930s. Syrups and jeilies for sweet-eating ants and grain baits for ground squirrels and prairie dogs formerly were available with concentrations as high as 3%. The use of thallium compounds as pesticides was banned by the USEPA in 1972 (U.S. EPA, 1985) but its use probably continues in some other countries.
See Section 61.6.3.4. 61.7.1.2 Toxicity to Laboratory Animals
Pathology Decrease in anterior horn cells with demyelination of the lateral columns of the spinal cord (associated with a case said to resemble amyotrophic sclerosis) has been reported; the difference may be related to the fact that a phenyl mercury compound was involved (Brown, 1954). Treatment of Poisoning See Section 61.6.3.6 and see Section 8.2 in the first edition of this Handbook.
Basic Findings Work in the first half of these twentieth century indicated that the oral LD 50 was in the range 7-24 mg Tl/kg (Dieke and Richter, 1946; Gettler and Weiss, 1943; Lehman, 1951, 1952; Munch and Silver, 1931). An extensive toxicological study on a number of thallium compounds by Down et al. (1960) indicated that the LD 50 for the acetate and oxide of thallium was in the range 12-30 mg TI/kg in rats, guinea pigs, rabbits, and dogs. Diets fed to rats for 15 weeks
61.7 Thallium
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showed increased mortality at levels of 4.5 mg Tl/kg as thallic oxide and 2.3 mg Tl/kg as thallium acetate. Rats given lA mg Tl/kg/day in the drinking water as thallium sulfate exhibited 21 % mortality after 240 days (Manzo et al., 1983).
61.7.1.3 Toxicity to Humans
Absorption, Distribution, Metabolism, and Excretion Thallium compounds are well absorbed in the gastrointestinal tract (Lie et aI., 1960). In industry, absorption is mainly respiratory, whereas nearly all accidental and intentional poisoning involves ingestion. Thallium rapidly distributes from the bloodstream to all tissues in the body. The kidneys tend to have the highest concentrations. Thallium also accumulates in active hair follicles, much less so in follicles in the resting state. There is indirect evidence of enterohepatic circulation as the portal vein contains higher levels than in peripheral blood (Frey and Schlechter, 1939). In rats, the fecal route is the predominant pathway of excretion. In studies using a radioactive tracer, Lie et al. (1960) observed a half-life of 3.3 days in rats. Thallium crosses the placental barrier in rats and mice (Gibson et aI., 1967), in rabbits (Frey and Schlechter, 1939), and in humans. Characteristic hair loss and nail bands were seen in a baby born 60 days after the mother ingested 750 mg of thallium sulfate (von Martius, 1953).
Therapeutic Use Apparently thallium sulfate has not been used as a drug. Thallium acetate was used as a dipilatory agent in treating fungal infections of the hair. Single oral doses at the rate of 4-10 mg/kg were employed. An interval of 2-6 months (usually 3 months) was recommended before readministration. Use of thallium was discontinued because it was found impossible to adjust the dosage so that the toxic etlects were restricted to the loss of hair. However, children who survived treatment were reported to grow normally. A review (Munch, 1934) showed there was serious poisoning in 5.5% of 8006 cases treated with a single oral dose of thallium acetate at different recommended dosage rates, mostly 8 mg/kg. There were 8 deaths in the series. No deaths were reported at dosages lower than 8 mg/kg, but such dosages were unreliable in causing hair loss, and one case of poisoning occurred at a dosage of 4 mg/kg. Adults were more susceptible to poisoning than children, so the drug was given orally only to children under 10 or 12 years of age (Felden, 1928). However, thallium acetate ointments were used for treating hypertrichosis in adults. Munch (1934) tabulated 59 intoxications, some involving severe polyneuropathy, muscular atrophy, marked cerebral involvement, and blindness following use of 3-8% ointments for 1 week to several months. There were no deaths in this series.
Pathology By far the most thorough study of tissue changes in rats poisoned by thallium is that of Herman and Bensch (1967), which must be consulted for details. Briefly, renal eosinophilic casts, enteritis, and severe colitis were present consistently in acute poisoning. In subacute poisoning, the brain contained frequent foci of perivascular cuffing and rare foci of recent necrosis. In chronic poisoning, no abnormality was visible by light microscopy, but degenerative changes frequently were present in the mitochondria of the kidney, liver, and intestines, and sometimes in those of the brain, seminal vesicles, and pancreas. Other changes included autophagic vacuoles and lipid droplets. The changes were not specific for thallium. These findings were similar to but more detailed than those of Cortella (1928). Treatment of Poisoning in Animals A number of antidotes have been used in the treatment of thallium poisoning in animals. These include BAL (Braun et aI., 1946; Lund, 1956; Moeschlin, 1965; Thyresson, 1951), cysteamine (Moeschlin, 1965), dithizone (Lund, 1956), cystine (Thyresson, 1951), and potassium chloride (Lund, 1956). The outcomes with these antidotes were equivocal. The agent of choice is Prussian blue [potassium ferrohexacyanoferrate(II), having the empirical formula KFeIII FeII (CN)6· nH20]. Given orally at 200 mg/kg daily for nine days, Prussian blue was found to be effective in rats (Heydlauf, 1969). Other investigators also found that Prussian blue was effective in experimental animals (for a review, see Stevens et al., 1974).
The most complete review of human thallotoxicosis is that of Munch (1934). For more recent reviews, see ATSDR (1992b) and Goyer (1991).
Accidental and Intentional Poisoning Signs and symptoms are referable mainly to the gastrointestinal tract, nervous system, and the hair and nails; the heart and kidneys may also be involved. After large doses, gastroenteritis is evident in about 12-14 hours, whereas neurological symptoms may be delayed 2-5 days. Following smaller doses, onset may be delayed as much as 3 weeks (Reed et aI., 1963). Gastrointestinal manifestations include severe paroxysmal abdominal pain, vomiting, diarrhea, anorexia, stomatitis, salivation, and weight loss. Neurological manifestations during the first days of illness may include paresthesias, headache, cranial nerve damage (ptosis, strabismus, optic atrophy), myoclonic or choreiform movements, convulsions, delirium, and coma. The occurrence of high fever is probably the result of brain injury; it indicates a bad prognosis. Vascular collapse and death may occur in 24-48 hours, but the course is usually more prolonged. Death may be caused by respiratory paralysis, pneumonia, or circulatory disturbances. Peripheral neuropathy, particularly in the legs, is common with severe pain, paresthesias, muscle weakness, tremor, ataxia, and atrophy (Sollmann, 1957). Obstinate constipation may interfere with treatment. Loss of hair begins after 1 to 2 weeks have elapsed. Several days before spontaneous shedding, impending alopecia is indicated by the ease with which the hair may be pulled out. In the more protracted cases, ataxia, choreiform movements, dementia, depression, and psychosis may be prominent. The neurological
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changes must be characterized as diverse (Steinberg, 1961) and sometimes as bizarre. Neurologic damage may be permanent. In fact, chronic neurological defects were found in 54% of 48 patients who survived and were followed up. A blue gingival line and dermatological abnormalities, including white bands in the nails, may appear. Liver damage occurs but is not prominent clinically. Kidney damage may occur (Chamberlain et al., 1958; Munch et al., 1933; Reed et al., 1963). Death may occur promptly from shock, or it may be delayed as much as 19 days after onset (Cavanagh et al., 1974). In a series of 72 children who ingested thallium, 9 died and 26 (36%) were left with persistent neurological damage. Prognostic ally unfavorable signs were gastrointestinal disturbances, cardiovascular abnormalities, coma, convulsions, and mental ion (Reed et al., 1963). An outbreak of poisoning resulting from the use of thalliumtreated grain as food is listed in Table 7.14 of the first edition of this Handbook. In chronic poisoning, symptoms can be nonspecific, and thallium intoxication may not be suspected unless depilation occurs. Although characteristic of thallium toxicity, hair loss also can result from poisoning with other metals and certain drugs. The same is true of the white bands on the nails. The findings already summarized are confirmed in many other reports of accidents, murders, and especially suicides (Bank et al., 1972; Bental et al., 1961; Chinen et al., 1977; Curry et al., 1969; Freund et al., 1971; Gastel et al., 1978; Gefel et al., 1970; Gerdts, 1964; Gharib, 1970; Goulon, 1963; Grulee and Clark, 1951; Hausman and Wilson, 1964; Koblenzer and Weiner, 1969; Marten, 1969; Ossa et al., 1975; Potes-Gutierrez and Del Real, 1966; Reibscheid et al., 1966; Sprensen, 1965; Taber, 1964; Vasil' eva et al., 1978). Dosage Response Even in children, who are less susceptible than adults, serious poisoning has followed a single dose of thallium acetate at a rate of 4 mg/kg. Twice that dosage poif soned 5.5% of treated patients. Moeschlin (1965) reported that death of an adult had followed a dosage of only 8 mg/kg, but that most fatalities involved dosages in the range 10-15 mg/kg or higher. Laboratory Findings Diagnosis of poisoning often can be confirmed by analysis for thallium in urine, blood, or hair. The concentration of thallium in normal tissue is low. The concentration in serum does not exceed 0.03 ppm (Gofman et al., 1964). The values for urine of unexposed individuals are below 1 J.l.g TUg creatinine (Schaller et al., 1980). Urinary levels as high as IQ-15 ppm have been found in a patient who survived (Mathews and Anzarut, 1968). In a group of poisoned children 1-6 years of age, the excretion of thallium soon after hospitalization was usually 1-3 mg/24 hr, but in one patient it reached a level of 15.4 mg/24 hr (Chamberlain et al., 1958). Similar concentrations (2-4 ppm) were found in a poisoned adult (Gettler and Weiss, 1943). In a case of subacute, nonfatal poisoning, the concentration of thallium in the urine varied from 0.2 to 0.3
ppm when first measured a month after ingestion and then declined more and more slowly to values of 0.05 to 0.06 ppm 20 days later (Rodermund et al., 1968). Proteinuria, cylindruria, and sometimes hematuria and oliguria may be present as indications of renal damage. Liver function tests may be abnormal. The blood picture and bone marrow are usually normal. However, there may be increased coproporphyrin and/or, more rarely, uroporphyrin in the urine. The serum iron may be slightly increased, suggesting a disturbance in the uptake of iron by the reticuloendothelial system. According to Moeschlin (1965), the spinal fluid shows a marked increase in protein, sometimes involving the globulin, but no increase in cells; a slight gold-sol and mastix reaction of the parenchymatous type may be present also. In other cases, the spinal fluid may be entirely normal (Chamberlain et al., 1958). Examination of the spinal fluid is not useful in prognosis (Reed et al., 1963). One or more bands of dark pigment may be demonstrated in hair near the root in as little as 4 days after ingestion. The bands do not contain increased thallium (Moeschlin, 1965). In many patients there is flattening or inversion of the T-waves in some leads of the electrocardiogram. In the few patients who have epileptic seizures, the electroencephalogram shows long stretches of low-voltage theta waves and low-voltage fast activity interrupted at intervals by slow high-amplitude waves that may be single or in groups, including regular runs of waves at the rate of 6 or 7 per second (Moeschlin, 1965). Other investigators have emphasized the slow, high-amplitude waves. Patients with abnormal electroencephalographic findings frequently die or fail to recover completely, while those with normal records usually survive without sequelae (Reed et al., 1963). Pathology Only visceral congestion and hyperemia and punctate hemorrhages of the gastrointestinal tract may be found if death occurs soon after exposure. In cases with longer survival, alopecia may be present. The skin is dry, and microscopic examination reveals marked lesions of the epithelium, follicles, and glands. Damage to the intestinal mucosa is common. There may be fatty degeneration of the heart and liver, degeneration of kidney tubules and the adrenal medulla, and edema and congestion of the lungs. At least a part of the varied tissue damage is explained by histological damage in the capillaries. Degeneration and demyelination occur in peripheral nerves. Similar changes have been found in one or more cases in the paravertebral sympathetic chain, the fasciculus gracilis, and in various other tracts or nerves, including the optic nerve. There may be retrograde injury to neurons in ganglia and nerve nuclei (Bank et al., 1972; Gettler and Weiss, 1943; Heyroth, 1947; Karkos, 1971; Moeschlin, 1965). The dying back may affect both motor and sensory nerves (Cavanagh et al., 1974). Finally, severe damage to a wide range of brain nuclei may be found (Kennedy and Cavanagh, 1976). Treatment of Poisoning The treatment of choice (following efforts to remove as much ingested poison as possible) would
61.8 Lead appear to include ferric ferrocyanide (Prussian blue) or activated charcoal, forced diuresis (with a moderate supplementation of total potassium chloride), and supportive measures as required. Among the latter, trihexyphenidyl (Artane) may be of value in selected cases. This compound, which is ordinarily used to treat Parkinsonism, caused a striking reduction in tremors in one series of cases of thallium poisoning (Stein and Perlstein, 1959). Ferric ferrocyanide in divided oral doses at a total rate of 250 mg/kg/day or even greater should be administered to minimize absorption of thallium from the intestine. Daily doses should be given to reduce reabsorption of the poison from the intestine. Although fecal excretion of thallium is ordinarily small in humans, the material is excreted into the gastrointestinal tract as in animals. Frey and Schlechter (1939) reported a case in which moderate quantities of thallium were found in the vomitus as late as the 35th day after onset. Rauws and van Heyst (1979) found that different preparations of Prussian blue differed in their ability to adsorb thallium. In general, colloidally soluble preparations are effective, but ideally every batch intended for antidotal use should be tested for thallium-binding capacity. No side effects were seen from ferric ferrocyanide even at a dosage of 416 mg/kg/day. Duodenal intubation is not required for administering this material. Its effects are dramatic when administered soon after ingestion and rewarding even when given in long-standing intoxication. It may promote substantial fecal excretion, even though the concentration of thallium in the urine has decreased to less than 0.5 mg124 hr (Barbier, 1974). On the contrary, when hemodialysis, forced diuresis, and Prussian blue were used in combined treatment, slightly more thallium was recovered in the urine than in the dialysate and only very little was found in the feces (Drasch and Hauck, 1977). It is not yet clear how the different forms of removal of thallium compete with one another, just which method is best, or whether combined treatment is necessarily better than the best single treatment. If ferric ferrocyanide is not available at once, the patient should be treated with repeated doses of activated charcoal (500 mg/kg twice daily) until Prussian blue can be obtained. Prussian blue, taken orally, is nonabsorbable.1t is believed to trap thallium ions in the intestinal tract via exchange with potassium ions in the Prussian blue mineral matrix (Stevens et aI., 1974). Many other antidotes have been tested but with equivocal results (Bass, 1963; Foreman, 1961; Reed et aI., 1963).
61.8 LEAD The present minor importance of lead compounds as pesticide would justify very little attention to them in this book. However, the great importance of other lead compounds as causes of accidental poisoning and the long persistence of lead arsenate following its earlier heavy use as a pesticide require some discussion of lead here. Those requiring additional information
1383
may find it in several reviews (Aub et al., 1926; Browning, 1969; Cantarow and Trumpeter, 1944; Goldwater, 1972; Goyer, 1974; Kehoe, 1964; U.S. Public Health Service, 1966; WHO, 1977).
61.8.1 LEAD ARSENATE 61.8.1.1 Identity, Properties, and Uses Chemical Name Structure
Lead arsenate.
PbHAs04.
Synonyms Other names for lead arsenate include acid lead arsenate, dibasic lead arsenate, dilead arsenate, dilead orthoarsenate, diplumbic hydrogen arsenate, lead hydrogen arsenate, and standard lead arsenate. Trade names include Arsinette®, Gypsine®, Ortho LlO Dust®, and Ortho L40 Dust®. The GAS registry no. is 10102-48-4. Physical and Chemical Properties Lead arsenate has the empirical formula AsH04Pb and a molecular weight of 347.13. It is a white, odorless powder with a density of 5.79. It decomposes above 280°C. It is soluble in dilute nitric acid and alkali but insoluble in water. It is stable to light, air, and water. It is not stable in acid, alkali, or sulfides. History, Formulations, and Uses Lead arsenate first was used as an insecticide in 1894. It is used for the control of moths, leaf rollers, and other chewing insects and in soil treatments for Japanese and Asiatic beetles in lawns. It is formulated as pastes or powders. Pastes contain at least 28.4% PbO, 14% AS20S; powder contains at least 62% PbO, 32% AS20S; neither contains more than 0.25% AS20S in water-soluble form. Lead arsenate may be used with summer oil emulsions, spray lime, wettable sulfur, and certain hydrocarbons and nicotine sulfate. An indefinite lead arsenate mixture called basic lead arsenate (see Table 61.4 in Section 61.12, also is used in the form of dusts, with sulfur or clay, and in water suspensions. Production of lead arsenate and of calcium arsenate is illustrated in Fig. 1.8) in the first edition of this Handbook. 61.8.1.2 Toxicity to Laboratory Animals Symptomatology The signs of acute poisoning by lead arsenate in rats are similar to those of other arsenic compounds (see Section 61.12.1). The animals show violent gastroenteritis and diarrhea and die from exhaustion and dehydration. Convulsions and general paralysis may occur before the onset of gastroenteritis. Death may occur within a few hours to several days. Following repeated doses, dogs may develop convulsions and paralysis not unlike the signs of encephalopathy seen in subacute poisoning in humans. Rats show nephropathy at least histologic ally and thus reproduce another injury seen in humans. Anemia is commonly present. Many of the subtle effects
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of chronic lead poisoning observed in humans are poorly reproduced in animals, although Aub et al. (1926) produced a convincing "wristdrop" in cats by weighting the paw. Dosage Response Voight et al. (1948) reported a marked difference between the oral LD 50 values of lead arsenate in the rabbit (125 mg/kg) and rat (825 mg/kg). However, Lehman (1951,1952) reported a value for the rat (100 mg/kg) not very different from that reported earlier for the rabbit. Dogs fed lead at rates of 0.33-1.10 mg/kg developed signs of poisoning in 120-215 days, and only some of them survived 229 days until the end of the experiment. Young animals were more susceptible than old ones. Dogs fed lead at rates of 1.8-2.6 mg/kg/day developed the first signs of poisoning in 14-44 days, and most of them died following convulsions and paralysis in 51-84 days after the beginning of feeding. Within the dosage range used, there was an approximately straight-line inverse relationship between rate of intake and the time at which signs of poisoning began. Dogs fed arsenic trioxide at arsenic rates corresponding to those obtained from lead arsenate remained well (Calvery et aI., 1938). A no-effect level for lead was not demonstrated. Male rats, but not females, were said to show decreased growth rate when fed lead arsenate at an average lead rate of about 0.21 mg/ kg/day (Laug and Morris, 1938), but the difference was of questionable statistical significance. There was a slight increase in the dry weight of the kidney relative to body weight, but this may have been the result of a nutritional disturb bance caused by pair feedings. It is possible that 0.21 mg/kg/day is a threshold level for toxicity in the rat. This result is very similar to that reported earlier by Smyth and Smyth (1932), who found a dosage of 0.265 mg/kg/day for 16 weeks harmless, 1.03 mg/kg/day questionably harmful, and a dosage of 3.93 mg/kg/day definitely harmful. Fairhall and Miller (1941) found that a dosage of 24 mg/kg/day caused some failure of growth in the course of a 2-year experiment. Horwitt and Cowgil (hor7) found that an intake of about 5 mg/kg/day produced no effect on growth, but an intake of about 10 mg/kg/day did produce anemia and growth retardation. Slight but reproducible growth retardation was reported by Laug and Morris (1938) in an experiment in which rats on a special low-lead diet received a supplement that added only 0.213 mg/kg/day to their lead intake. Absorption, Distribution, Metabolism, and Excretion Absorption of lead arsenate is generally gastrointestinal. This is likely to be true following respiratory exposure to coarse dusts as well as after ingestion. Dermal absorption is extremely small. Lead and arsenic derived from lead arsenate are distributed separately in the body. Lead is stored in highest concentration in the bone with much lower concentrations in soft tissues. Arsenic, although constituting a smaller proportion of the molecule, is stored in the liver and in some instances in the kidney at higher concentrations than those for lead (Fairhall and Miller, 1941).
Lead is transferred to the fetus of animals and humans (Calvery, 1938; Horiuchi et aI., 1959; Legge and Goadby, 1912; Morris et al., 1938). Dogs and rats fed lead at a constant rate store less lead when fed a normal or high-calcium diet than when fed a lowcalcium diet (Calvery et aI., 1938; Grant et aI., 1938). In fact, the metabolism of lead is very similar to that of calcium. The same factors act on both, and each tends to displace the other. Rats store lead at the lowest measurable dietary intake, and storage is proportional to intake (Laug and Morris, 1938). When rats ingest lead at ordinary dietary levels, it is stored in highest concentration in bone, but when feeding is at a high level for a few weeks, the highest concentration is in the kidney (Laug and Morris, 1938). Mode of Action Because lead arsenate contains two toxic elements, it might be predicted that its toxicity would show an influence of both kinds of poisoning. Fairhall and Miller (1941) reported evidence that this is true; for example, it was clear that increased hemosiderin of the spleen was an effect of arsenic, whereas certain inclusion bodies in the kidney were an effect of lead. Furthermore, as described later, this study suggests that rats are more susceptible to mortality caused by prolonged exposure to the arsenic moiety of lead arsenate, whereas the work of Calvery et al. (1938) suggests that dogs are more susceptible to the lead moiety. One group of rats was fed lead arsenate at an approximate rate of 10 mg/rat/day, which would be about 40 mg/kg/day when they reached adulthood. Other groups were fed lead carbonate and calcium arsenate at such a rate that they received an equivalent dosage of lead and arsenic, respectively. A control group received the same diet but none of the chemicals. During the 2-year course of the experiment, lead carbonate did not increase the mortality significantly over that of the controls (42%). However, both arsenic compounds caused an approximately equal increase in mortality (56% or more after 2 years). At the end of 1 year, half of the living animals were killed and their tissues analyzed. Similar analyses were made on the rats that survived for 2 years. Animals fed lead arsenate stored more arsenic than lead in their soft tissues, even though the rate of arsenic intake was less (21.6% of molecule as compared to 60.7%). The storage of lead in bone was very great, as expected. More interesting was the fact that (with the possible exception of arsenic in the liver) all organs stored less arsenic and less lead when they were fed simultaneously as lead arsenate than when fed separately at the same rates. The difference (at least for lead) was attributed to decreased absorption or increased excretion, but the possibility that solubility might be involved could not be excluded (Fairhall and Miller, 1941). Treatment of Poisoning in Animals CaNa2-EDTA was more effective than unithiol in promoting excretion of lead (Soroka and Sorokina, 1969).
61. 9 Tin
61.8.1.3 Toxicity to Humans Experimental Exposure One man was given lead arsenate at the rate of 40 mg/day or about 0.57 mg/kg/day for 50 days without ill effect (Cardiff, 1940). In a separate study, two men ingested lead arsenate at a rate of 100 mg/man/day for 10 days without injury (Fairhall and Neal, 1938). These levels are far above the level of lead in soluble form found to produce mild poisoning in a somewhat longer period. The difference in effect presumably is due to the insolubility of lead arsenate, although duration of exposure may play a part also. Accidental and Intentional Poisoning There have been astonishingly few reports of poisoning by lead arsenate. Tobaldin et al. (1966) did report cases of nonfatal poisoning in two girls, ages 8 and 10 years. The initial symptoms of gastrointestinal irritation were what one would expect in acute poisoning by a metal salt. The authors concluded that the effects were the result of both elements. Use Experience In an apple-raising area, many people, including children, ate apples sprayed with lead arsenate so that the average intake of the compound was about 5 mg/person/day. They suffered no ill effect, and the only change was a marginal increase in the levels of lead in the blood and of lead and arsenic in the urine as compared with samples from an urban community (Neal et aI., 1941b). In a 14-month study of 542 formulators and applicators of lead arsenate sprays for apple orchards, Neal et al. (1941 b) found only 7 who had a combination of clinical and laboratory findings directly referable to the absorption of the compound. Whether the very mild observed conditions represented poisoning was left moot, but it was noted that not a single case met the criteria established by the Committee on Lead Poisoning of the American Public Health Association. Dermatitis was infrequent and not related to exposure to lead arsenate. The fertility of married orchardists was normal. The degree exposure of these orchardists was indicated by average lead concentrations of 5.74 mg/m 3 during mixing and 0.45 mg/m 3 during spraying. Arsenic values were proportionally smaller. These orchardists, who had had many years of heavy exposure to lead arsenate as applicators and the arsenic content of whose urine in 1938 averaged 0.141 ppm (men) and 0.098 ppm (women), were studied again in 1968 and 1969. The standard total mortality ratio (SMR) for the orchardists compared to the general population of the same state was only 0.65. The corresponding SMRs for heart disease, cancer, and stroke were 0.60, 0.66, and 0.76, respectively, indicating that this cohort experienced less total mortality and less mortality from heart disease, cancer, and stroke (Nelson et aI., 1973).
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a benign pneumoconiosis. Organic tin compounds have produced gastrointestinal disturbances, tremor, convulsions, paralysis, and death in animals. Triethyltin compounds, which readily cross the blood-brain barrier, produce severe neurological disturbances in both animals and humans (see Section 61.9.2). Thus lead and tin may injure the same organs, and some alkyl compounds of both are especially dangerous. However, the situation is far from simple. Alkyl derivatives oflead, like those of antimony, bismuth, and mercury, damage nerve cells, whereas triethyltin produces an interstitial edema of the white matter without damage to the cells. Some tin compounds that are closely related chemically have very different toxic effects. Triethyltin causes edema of the central nervous system; diethyltin causes hypertrophy of the bile ducts in some species. The same compound may be highly toxic to one species and essentially harmless to another. An example is dibutyltin, which is highly toxic to rats but not to guinea pigs. In addition to these differences observed in the same laboratory, there are unusually large quantitative differences in the toxicity of certain compounds (e.g., triphenyltin) as observed in different laboratories. Unexplored possibilities to explain the quantitative differences include the strain of rat and the supply of essential trace elements such as zinc. Traces of inorganic tin are widely distributed in nature. In addition to the amount naturally present in food, tin may be added from foil and plated cans used as containers. Organic tin does not occur naturally. No conclusion about its safety can be drawn from experience with inorganic compounds. Organic tin compounds have some importance as fungicides and they have been proposed as insecticides or antifeeding compounds for insects. So far, these uses have led to no serious difficulty. However, the fact that related compounds have produced serious injury and death in humans serves as a warning that any tin compound proposed for use as a pesticide should be studied with particular care. The following sections are confined to recognized pesticides and to an extremely brief summary of ethyltin compounds. Although the latter compounds have only been proposed as pesticides, it is necessary to keep them in mind in any consideration of pest control based on tin. Comprehensive reviews of the toxicology of tin are those of Barnes and Stoner (1959), LeBreton (1962), and ATSDR (1992c).
61.9.1 FENTIN ACETATE 61.9.1.1 Identity, Properties, and Uses Chemical Name Triphenyltin acetate.
61.9 TIN Tin is in the same periodic group as lead. Under conditions of use, inorganic tin compounds have proved essentially harmless, although inhalation of stannic oxide commonly leads to
Synonyms Fentin acetate (BSI, ISO) is the common name in use. A trade name for the compound is Brestan, Code designations include ENT-25,208, GC-6,936, Hoe-2,824, and VP1,940. The CAS registry no. is 900-95-8.
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Physical and Chemical Properties Fentin acetate has the empirical formula C20H1802Sn and a molecular weight of 409.06. It forms white odorless crystals that melt at 122-124°C. The vapor pressure at 30°C is 1.33 x 10-6 torr. Its solubility in water at 20°C is 28 mg/l. It is poorly soluble in organic solvents. Fentin acetate is stable when dry but rapidly hydrolyzed by water to the hydroxide, a white solid that melts at 118-120°C. The hydroxide practically is insoluble in water but is soluble in most organic solvents. History, Formulations, and Uses Fentin acetate was introduced in 1954. It is used as a fungicide on turnips, potatoes, celery, beans, and other crops. The technical product is 90-95% pure. It is formulated as wettable powders of 20 and 60% concentration and applied at a rate of approximately 160-260 g of active ingredient per hectare. Fentin acetate also is used as an antifeeding compound for insect control. 61.9.1.2 Toxicity to Laboratory Animals Basic Findings Rats poisoned by a single dose of fen tin acetate show sluggishness, unsteadiness, moderate diarrhea, anorexia, bloody stain around the nose and eyes, and wheezing. The acute toxicity of fentin acetate is shown in Table 61.3. It is clear that absorption from the gastrointestinal tract is poor compared with that from the peritoneum. Klimmer (1964) found that 70% of rats dosed by stomach tube at a daily rate considered equivalent to 50 ppm in the diet died of secondary infection in an average of 26.6 days. Stoner (1966) found that rats survived a dietary level of 200 ppm for 10 weeks, but most died when the concentration was raised to 300 ppm. The difference in response almost certainly depended Table 61.3 Acute Toxicity of Fentin Acetate LD50 (mg/kg)
Species
Route
Rat
Oral
136a
Klimmer (1964)
Rat, F
Oral
491 b
Stoner (1966)
Rat
Dermal
450
Rat
Intraperitoneal
Rat, M
Intraperitoneal
8.5
Rat,F
Intraperitoneal
11.9
Stoner (1966)
Mouse, M
Oral
81.3
Stoner (1966)
7.9
Stoner (1966)
13.2
Reference
KIimmer (1964) Klimmer (1964) Stoner (1966)
Mouse, M
Intraperitoneal
Guinea pig
Oral
21
Klimmer (1964)
Guinea pig, M
Oral
10
Stoner et al. (1955)
Guinea pig
Intraperitoneal
5.3
Klimmer (1964)
Guinea pig, M
Intraperitoneal
3.74
Stoner (1966)
Rabbit
Oral
30-50
KIimmer (1964)
Rabbit
Intraperitoneal
10
KIimmer (1964)
a In methylcellulose. bIn arachis oil.
on the dosage schedule. Even a level of 25 ppm may reduce food intake, growth, and the number of leukocytes in the blood (Verschuuren et aI., 1966). Guinea pigs are more susceptible than rats to triphenyltin acetate, on both an acute and a long-term basis. Dietary levels of 50 ppm are fatal in a few weeks. Levels as low as 5 ppm cause growth inhibition and reduction in hemoglobin and white cells in the blood (Stoner, 1966; Verschuuren et aI., 1966). Even at 1 ppm, food intake was reduced (Stoner, 1966). The injury is not merely one of starvation; treated guinea pigs grow more slowly than pair-fed controls (Stoner and Heath, 1967). The effect of fentin acetate is highly cumulative in the guinea pig. The logarithmic mean survival time on 25 ppm was twice that on 50 ppm (Stoner, 1966). However, the log mean survival time (> 190 days) on 12 ppm was greater than would be predicted on the basis of a completely cumulative effect (Stoner and Heath, 1967). Histologically, only a few animals fed triphenyltin acetate displayed interstitial edema of the brain. Dietary levels of 20 ppm and higher significantly increased the water content of the brain and spinal cord in guinea pigs. A dietary level of 500 ppm significantly increased the water content of the spinal cord of rats, but only in the females (Verschuuren et aI., 1966). In spite of these morphological and chemical changes in the brain and spinal cord at high levels of intake, the symptomatology does not suggest that injury to the central nervous system is critical in poisoning by triphenyltin acetate. Triphenyltin acetate (and chloride) produced marked testicular atrophy in rats when fed for 20 days at a concentration giving an intake of 20 mg/kg/day (Pate and Hays, 1968). The effect was more striking than that reported for triphenyltin hydroxide but the difference probably was explained by the higher dosage involved. The second segment of the log(time)-log(dosage) curve for fatal dosages of triphenyltin acetate in guinea pigs was presented by Scholz (1965). As expected (see Section 1.2.3,2), it closely fits a straight line. Fentin acetate was not tumorigenic when administered to mice for about 18 months at the maximal tolerated level (Innes et aI., 1969). Sachsse et al. (1987) found no adverse effects in dogs given 0.6 mg/kg/day in the feed for 1 year. Tennekes et al. (1989), in studies of rats given triphenyltin in the diet for 2 years, found evidence of tubular atrophy of the testes and atrophy of the sciatic nerve at dosage rates down to 0.3 mg/kg/day. Lowest effect levels in mice for chronic oral exposures appear to be higher than those for the rat. For example, a National Cancer Institute study (NCI, 1978) did not find adverse effects at 3.75 mg/kg/day over a 78-week oral exposure. The Agency for Toxic Substances and Disease Registry (ATSDR, 1992c) have reviewed evidence for the tumorigenicity of triphenyltin compounds in animal studies. Most tumors are found in the endocrine glands. Tumors have also been found in testicular Leydig cells and in hepatic cells.
61.9 Tin
Absorption, Distribution, Metabolism, and Excretion Absorption from the skin is poor (Stoner, 1966). During oral dosing of sheep for 20 days with fentin acetate at the rate of 10 mg/day, 113Sn was found in the milk at an average concentration of 0.0017 ppm. Tin was present as fentin acetate and in two unidentified forms. Seventeen days after dosing was stopped, the concentration had fallen to the limits of detectability. During treatment, the concentration 0.0029 ppm in the blood and 0.0075 ppm in the urine. Eight days after dosing was stopped, the liver, kidney, lung, pancreas, gallbladder, and brain contained higher concentrations of 113 Sn than did other organs, and the level was still increased in the liver after 218 days (Herok and G6tte, 1964). In cows and sheep, triphenyltin is excreted chiefly in the feces (Briigemann et aI., 1964; Herok and G6tte, 1964). Triphenyltin is rapidly distributed to all tissues, including the brain of rats. It can still be detected in the brains of rats and guinea pigs more than 30 days after a single dose (Heath, 1966). Mode of Action A single dose in the fatal range produces irritation of mucous membranes and death due to respiratory failure. Repeated doses produce no injury unless they are large enough to interfere with food intake. However, the cause of death following repeated doses is not clear. Starvation undoubtedly contributes. In some but not all laboratories, secondary infection has been an important cause of death of animals receiving this compound (Klimmer, 1964; Verschuuren et al., 1966). A basis for susceptibility to infection was found in guinea pigs, which showed reduced lymphopoiesis, including atrophy of the white pulp of the spleen (Verschuuren et al., 1966). It may be that increased susceptibility to infection explains the morbidity and mortality caused by fentin acetate to a far greater extent than has been proved. Infection is difficult to recognize in the absence of inflammation. 61.9.1.3 Toxicity to Humans Poisoning of two pilots and three mechanics followed the aerial application to field crops of a mixture of fentin acetate and manganese ethylenedithiocarbamate without stringent observation of safety regulations. The pilots were severely ill. Gastrointestinal irritation was the main problem, but the liver was affected in one. Illness in the mechanics was mainly subjective (Horacek and Demcik, 1970). Laboratory Findings Apparently there is no report of tissue levels of tin following poisoning by fentin acetate. The occurrence of tin in normal people is discussed in Section 61.9.2.3. 61.9.2 ETHYLTIN AND RELATED COMPOUNDS 61.9.2.1 Identity, Properties, and Uses Compounds and Their Characteristics Alkyltin compounds occur in the following forms: RSnX3, R2SnX2, R3SnX,
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and &,Sn, where R is an alkyl group linked directly to tin and X is a simple or complex ion. It has been found that the toxicity of these compounds and of certain phenyltin compounds to mammals and fungi depends mostly on the R constituent and to only a minor degree on the nature of X. Tributylin salts have the greatest fungicidal activity, and the effectiveness decreases with changes in the R component in the following order: tributyl> tri-n-propyl = tri-iso propyl> triethyl = diethylphenyl > triphenyl > trihexyl > trimethyl > trioctyl = diethyl. The monoethyl- and tetraethyl-tin compounds were ineffective as fungicides (Van der Kerk and Luijten, 1954). Triethyltin salts have the greatest toxicity for mammals. 61.9.2.2 Toxicity to Laboratory Animals Basic Findings The toxicity of ethyltin trichloride is low. An intravenous dose at the rate of70-150 mg/kg quickly produced hyperpnea, vasodilatation, prostration, and muscular tremors in rabbits, but they recovered in about an hour. Rats showed no distinctive signs after an intraperitoneal dose of 200 mg/kg (Stoner et al., 1955). Diethyltin compounds do not produce clearly neurological effects. They produce a more generalized illness similar to that caused by the triphenyltin. They also produce in the rat and mouse but not the rabbit, guinea pig, cat, or hen a characteristic injury of the bile ducts (Bames and Stoner, 1959). The lesion was first described in great detail in connection with dibutyltin dichloride (Bames and Magee, 1958). The injury starts within 60 min of an intravenous dose or 4 hr of an oral dose as a localized lesion of the lower part of the bile duct. The initial lesion, accompanied by an apparently minor extravasation of bile into the surrounding pancreatic tissue, sets up a chain of events leading to: (a) partial blockage of the bile duct and its great dilatation, (b) inflammation of the walls of the main vessels of the portal tracts, (c) thrombosis in some ofthese vessels, and (d) sharply localized areas of necrosis resembling infarcts within the liver parenchyma. The bile duct may rupture, leading to peritonitis and fat necrosis. It is of interest that this unusual injury apparently occurs only in species in which the pancreatic ducts enter the lower part of the common bile duct rather than entering the duodenum directly. It is also of interest that the general toxic effects of dibutyltin are prevented by BAL, but the local action on the biliary tract is not prevented Triethyltin produces significant swelling of the brain and spinal cord and a striking noninflammatory interstitial edema of their white matter without detectable damage to the nerve cells. Chemical study indicates that the fluid between the fibers is an ultrafiltrate of plasma. Even among compounds that produce brain edema, triethyltin was considered unique in not affecting the gray matter (Magee et al., 1957). Later it was found that hexachloraphene produces a similar lesion (Kimbrough and Gaines, 1971). Poisoning by triethyltin is manifested by progressive weakness, paralysis, and convulsions. In rats, the sulfate is about equally toxic orally, intravenously, or intraperitoneally. The intraperitoneal LD 50 is 5.7 mg/kg (Stoner et al., 1955). Rats fed triethyltin hydroxide at a dietary level of20 ppm
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lost weight, but only a little more rapidly than pair-fed controls. Unlike the controls, they showed ataxia of the hind legs in 7-9 days when they had consumed a total of about 10 mg/kg. Weakness progressed until the hind legs lay motionless while the rats dragged themselves about using their front legs. The hind legs were not completely paralyzed and could be withdrawn if pinched. If feeding were continued, the front legs, too, became weak, and the rats lay helplessly on their sides, although they would still attempt to eat. About two-thirds of them died during the third week, but the remainder appeared to become resistant. Their food intake increased and they showed partial recovery. Recovery was more certain if they were transferred to a dietary level of 10 ppm when they first reached a state of prostration. Such rats appeared almost normal after 6 weeks on the lower concentration, although there was some wasting of the muscles of the hind legs and some difficulty in balancing on the edge of a cage. No further improvement occurred even if the rats were returned to a normal diet, but the histological lesion cleared. If, on the other hand, the rats were continued on diets containing triethyltin hydroxide, they gradually became irritable, tremorous, and unwell. The tremor was made worse by intentional movement. Some of these rats suffered intermittent or almost continuous convulsions (Magee et al., 1957; Stoner et al., 1955). Tetraethyltin has the same effect as triethyltin but acts more slowly because its toxic action really depends on its metabolism to triethyltin (Cremer, 1958). In general, methyltin compounds are less toxic than corresponding ethyltin compounds. For trialkyl compounds, the toxicity decreases gradually as the length of the side chains is increased beyond 2. In the dialkyl series, the propyl and butyl compounds are more toxic than either ethyl or methyl, and toxicity remains high through octyl (Barnes and Magee, 1958). Single oral doses in the hamster, monkey, and gerbil produce neurological damage at 3 mg Sn/kg (Brown et al., 1979). One study in rats claims to find marked proximal tubular necrosis in rats after a single oral dose of 10 mg Sn/kg (Opacka and Sparrow, 1985). It is difficult to find long-term studies on ethyltin and related compounds. Brown et al. (1979) conducted a 4-week study of rats given daily oral doses of triethyltin chloride. Neuronal alterations were seen at a dosage rate of 4 mg Sn/kg/day. Absorption, Distribution, Metabolism, and Excretion There is considerable variation in the absorption of alkyltin compounds. For example, in the rat triethyltin is equally toxic by mouth and by intraperitoneal injection, but in the hen it is not toxic orally. In the rat, diethyltin is toxic when given by stomach tube, but it apparently reacts with some constituent in the food so that even high dietary concentrations are without effect. Ethyltin trichloride injected intraperitoneally is almost all excreted unmetabolized in the urine, with only trace amounts in the bile; when the compound is given by mouth it is excreted in the feces, indicating very limited absorption from the gastrointestinal tract (Bridges et aI., 1967).
Diethyltin dichloride injected intraperitoneally is partially metabolized to ethyltin, much of which is excreted in the urine; a portion of the unmetabolized compound is excreted in the bile and eventually appears in the feces. The conversion of diethyltin to ethyltin apparently occurs in the tissues and in the gut, although the conversion has been demonstrated in vitro only in cecal contents and not in liver homogenate (Bridges et aI., 1967). When the triethyltin chloride is injected intravenously into rats, guinea pigs, or hamsters, it reaches a higher concentration in the liver than in other tissues, including the brain. In the rat, this relative distribution is maintained from 10 min to 5 days after injection. In 4 days, rats excrete about 50% of an intravenous dose, mostly in the feces. The compound is probably excreted in the bile of rats, and it certainly reaches high concentrations in the bile of guinea pigs and hamsters (Rose and Aldridge, 1968). Mode of Action Dialkyltin compounds inhibit a-keto oxidase activity leading to the accumulation of pyruvate. This is considered important in the toxic action. BAL blocks this inhibition and the acute toxicity (Aldridge and Cremer, 1955; Barnes and Stoner, 1959). Triethyltin is the most effective known inhibitor of oxidative phosphorylation, being active at 1 x 10-7 M in vitro. This inhibition is not blocked by BAL (Aldridge and Cremer, 1955). Inhibition of phosphorylation must interfere with energy exchange everywhere in the body. It does not explain why the visible lesion is confined to the white matter of the central nervous system (Barnes and Stoner, 1959). 61.9.2.3 Toxicity to Humans Therapeutic Use Tin and various inorganic compounds of it were promoted during the early part of the twentieth century for treating staphylococcal infection or helminthic infestation. These treatments were usually harmless to patients and pathogens alike, although Bar (1956) reported a case of poisoning by stannous oxide. However, in 1953, a large outbreak of poisoning followed the introduction in France of an organic tin preparation called "Stalinon" for treating staphylococcal infection of the skin, osteomyelitis, anthrax, and acne. The preparation was supposed to be diethyltin diiodide, but it contained about 10% of triethyltin iodide and a little less ethyltin triiodide. One report tabulated 224 cases, including 103 deaths (LeBreton, 1962). There were a few additional cases and deaths reported by others. It was estimated that about 1000 people took the medicine. Symptoms began in 1-30 days after dosage was started. The total dose varied from 81 to 1094 mg or about 1-16 mg/kg in fatal cases. Signs and symptoms included headache, vomiting, photophobia, dizziness, abdominal pain, marked weight loss, hypothermia, bradycardia, retention of urine, alteration of consciousness, psychic changes, convulsions, coma, and paralysis. In spite of this array of abnormalities and the high mortality rate, many patients presented without any diagnostic signs. Some patients had abnormalities of the fundus of the eye, the
61.10 Bismuth
spinal fluid, or the electroencephalogram, but these changes were often absent in patients who went on to die. The level of consciousness was the best prognostic sign; only one patient who became comatose survived. A few patients recovered rapidly, but most who lived required several months to recover, and one required 18 months. Still others were left with permanent sequelae, including blindness and severe, flaccid paraplegia. Death occurred as early as the day of onset; in most instances it came 1-13 weeks after onset but was delayed 36 days in one case. The symptomatology and the autopsy findings indicated edema of the white matter of the brain as the cause of death. Consistent with this view was the finding that surgical decompression of the brain constituted the only useful treatment; six of eight patients treated in this way survived (A1ajouanine et aI., 1959; LeBreton, 1962). Thus, the action was characteristic of triethy1tin, not of diethyltin, which constituted a little more than 80% of the preparation. It appears that humans react the same as other animals to triethyltin but are far more sensitive. Laboratory Findings In the victims of "Stalinon," the concentration of tin was similar in the heart, lung, liver, and kidney, although it tended to be slightly lower in the heart and higher in the liver. The range for pooled viscera was 1.5-16 ppm with a mean of 5.6 ppm. The concentration in die brain was consistently higher, with a range of 5.1-39 ppm and a mean of 17.3 ppm (LeBreton, 1962). In the United States and France, tin may be found in most samples of all organs except the brain, which rarely contains a detectabte quantity in normal people. The median concentration in most organs is no more than 1 ppm (Kehoe et aI., 1940; LeBreton, 1962; Tipton and Cook, 1963). The concentration of tin in the blood of normal Americans averages 0.14 ppm and that in urine 0.018 ppm (Kehoe et al., 1940). The concentration in normal plasma does not exceed 0.47 ppm (Gofman et aI., 1964). The distribution of tin in the organs of the French patients was entirely different from that found in experimental animals injected intravenously with triethyltin chloride, where more tin is retained by the liver than by the brain. Whether the difference depends on species or on the fact that "Stalinon" was a mixture is not clear.
61.10 BISMUTH Bismuth is a heavy metal immediately following lead in the periodic table. It is in the same group as arsenic and antimony. Bismuth compounds used as pesticides are highly insoluble. They have proved safe when used as drugs in doses far larger than any likely to arise from their use as pesticides. In persons without special exposure, the concentration of bismuth in the plasma does not exceed 0.20 ppm (Gofman et aI., 1964).
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61.10.1 BISMUTH SUBCARBONATE 61.10.1.1 Identity, Properties, and Uses Chemical Name Structure
Bismuth subcarbonate.
(BiOhC03.
Synonyms Synonyms include bismuth oxycarbonate and basic bismuth carbonate. The CAS registry no. is 5892-10-4. Physical and Chemical Properties Bismuth subcarbonate has the empirical formula CBizOs and a molecular weight of 510.01. It is an odorless, tasteless powder that is practically insoluble in water or alcohol but soluble in mineral acids and concentrated acetic acid. Use When added to arsenical and fluoride formulations, bismuth subcarbonate tended to increase the feeding of insects on poisoned foliage. 61.10.1.2 Toxicity to Laboratory Animals Basic Findings Apparently the toxicology of the specific bismuth compounds used or proposed as pesticides has not been studied in animals. It seems reasonable to suppose, however, that sufficient dosages of relatively insoluble pesticides would produce changes similar to all but the most acute injuries produced by soluble bismuth compounds. Sollmann and Seifter (1942) found that intravenous doses of sodium bismuth citrate and bismuth glycolate were lethal to rabbits in a dosage range of 2.5-5 mg Bi/kg. Death appeared to be due to renal failure. The oral LD 50 is 22 g/kg for rats and is 0.5 g/kg for rabbits according to Kruger et al. (1985). The acute toxicity of bismuth depends greatly on the solubility of the compound under study. Pathology The tissue changes in rabbits caused by several bismuth compounds used in the treatment of syphilis were found to be characteristic of bismuth and not dependent on specific compounds or any particular route of administration. The major lesion involved the epithelium of the convoluted tubules of the kidney. These cells showed all types of degeneration from cloudy swelling to calcification. The glomeruli did not appear primarily damaged, but the glomerular capillaries frequently contained coagulated masses of erythrocytes. The lesions of the liver were much less conspicuous than those of the kidney, but cloudy swelling and small areas of necrosis flooded by erythrocytes did occur. Other organs showed no dependable or characteristic pathology. 61.10.1.3 Toxicity to Humans Therapeutic Use Bismuth subcarbonate is used to furnish mechanical protection and to exclude irritants from external ulcers, fistulas, or inflamed mucous membranes or from gastrointestinal inflammations or ulcers. It helps to allay diarrhea.
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Formerly it was used as an adjuvant to more active amoebacides against intestinal amoebae and as an opaque medium in roentgenoscopy. The usual oral dose in treating peptic ulcer is 1000-2000 mg suspended in fluid before each meal. The material is harmless unless applied to extensive bums or ulcers or left in fistulas for long periods. Formerly the compound was given as an X-ray contrast medium al a dose of about 50,000 mg. The disadvantage compared to barium sulfate was not toxicity but a delay in emptying of the stomach (Sollmann, 1957). Treatment of Poisoning BAL has been used for treating poisoning by any bismuth compound. Other treatment is symptomatic. 61.10.2 BISMUTH SUBSALICYLATE 61.10.2.1 Identity, Properties, and Uses Chemical Name Structure
Bismuth subsalicylate.
See Fig. 61.1.
Synonyms Other names for bismuth subsalicylate include basic bismuth salicylate and salicylic acid basic bismuth salt. The CAS registry no. is 14882-18-9. Physical and Chemical Properties Bismuth subsalicylate has the empirical formula C7HSBi04 and a molecular weight of 362.11. It forms microscopic prisms mat are decomposed by boiling water and by alkalies into a more basic salt. It is almost insoluble in water or alcohol. Use Bismuth sub salicylate is a fungicide, especially for the control of bluemold disease of tobacco seedlings. 61.10.2.2 Toxicity to Laboratory Animals See Section 61.10.1.2.
o
61.10.2.3 Toxicity to Humans Therapeutic Use The compound is sometimes used orally to allay diarrhea or to soothe gastritis or peptic ulcer. The dose varies from 500 to 2000 mg and may be given three times a day. Before the advent of penicillin, bismuth subsalicylate was much used in the treatment of syphilis, and it may still be used in patients who are intolerant to penicillin. It is given by intramuscular injection as a 10% suspension in oil at the rate of 130 mg/week, usually for 8-10 weeks. This treatment results within about 3 weeks in a urinary excretion of about 1 mg/day. By using a dose of 260 mg, a urinary level of 2 mg/day can be achieved in 1 week. Even when its use was extensive, the gradual intramuscular injection used against syphilis rarely led to serious poisoning. It was customary to stop treatment if gingivitis, albuminuria, cutaneous eruptions, or marked diarrhea appeared. If one of these signs was ignored and dosing continued, serious ulcerative stomatitis with salivation was likely to appear and soon be followed by one or more of the following: malaise, headache, insomnia, depression, asthenia, joint pains, nausea, loss of appetite, diarrhea, loss of weight, albuminuria, and skin reactions (such as puritis, herpes zoster, purpura, and sometimes serious exfoliative dermatitis). Jaundice and conjunctival hemorrhage were rare complications of treatment. A line of bismuth sulfide often appeared at the gingival margin along with similar blotches on the mucosae of the mouth, tongue, throat, colon, rectum, cecum, and appendix. The bismuth line was not a contraindication to treatment. Even when illness occurred, it was usually less severe than that associated with other heavy metals. Specifically, the stomatitis and albuminuria were less severe than those caused by mercury and the cutaneous eruptions less severe than those caused by arsphenamine (Sollmann, 1957). After treatment, the highest concentration of bismuth was found in the kidneys ami the lowest in the brain and blood. Bismuth freely crosses the placental barrier (Sollmann, 1957). Pathology Two patients who died following bismuth therapy but not as a result of it both showed refractile globules in the nuclei and cytoplasm of the epithelium of the convoluted tubules. Similar globules were found in the renal epithelial cells of rats following injection with one of the same compounds (Pappenheimer and Maechling, 1934). Treatment of Poisoning BAL has been used for treatment of poisoning by any bismuth compound. Other treatment is symptomatic.
61.11 ANTIMONY
Antimony potassium tartrate Figure 61.1 Two organometallic pesticides.
Antimony is a metal. It follows tin in the periodic table, and it belongs to the same group as arsenic. The trivalent compounds of antimony, like those of arsenic, are more toxic than the pentavalent ones. The toxic effects of antimony compounds closely resemble those of arsenic compounds, but antimony causes greater vomiting and is excreted more rapidly.
61.11 Antimony Traces of antimony are widely distributed in the environment. Eight or more compounds of antimony had some use as insecticides before DDT became available. However, antimony potassium tartrate was the only one of much commercial importance. The toxicity of this compound is typical of the group. Two useful reviews of the toxicology of antimony, which emphasize the industrial aspects, are those of Fairhall and Hyslop (1947) and ATSDR (l992d).
61.11.1 ANTIMONY POTASSIUM TARTRATE 61.11.1.1 Identity, Properties, and Uses Chemical Name Structure
Antimony potassium tartrate.
See Fig. 61.1.
Synonyms Other names include antimony tartrate, potassium antimonyl tartrate, tartar emetic, tartarized antimony, and tartrated antimony. A code designation is ENT-50,434. The CAS registry no. is 28300-74-5. Physical and Chemical Properties Antimony potassium tartrate has the empirical formula C4H4K07Sb and a molecular weight of 324.92. It forms transparent crystals or powder with a sweetish, metallic taste. The density is 2.6. It is soluble in water and glycerol, insoluble in alcohol. Crystals effloresce upon exposure to air. Formulations and Uses The compound serves as a poison in baits to control insects, especially thrips, and as an emetic (for people or pets) in baits to control rodents. The insect baits usually are applied as sprays containing 0.36-0.48% of the compound in the liquid formulation. The rodent baits are pastes or solids containing 0.3-3% of the compound, specifically 0.30% with thallium sulfate and zinc phosphide, 1.12% with arsenic trioxide, and 3.00% with ANTU. 61.11.1.2 Toxicity to Laboratory Animals Basic Findings The intraperitoneal LD 50 for mice is 46-50 mg/kg for the different isomers of antimony potassium tartrate (Haskins and Luttermoser, 1950). The subcutaneous and intravenous LD 50 values in mice are 55 and 65 mg/kg, respectively (Ercoli, 1968). The oral LD 50 for the same species is 600 mg/kg. The marked difference between these values is an indication of poor gastrointestinal absorption even in an animal that does not vomit. Other species are somewhat more susceptible, with an intraperitoneal LD 50 of 30 mg/kg (11 mg/kg as antimony) in rats and 25 mg/kg in guinea pigs (Bradley and Fredrick, 1941). Lifetime exposure of rats (Schroeder et al., 1970) and mice (Kanisawa and Schroeder, 1969; Schroeder et aI., 1968) to antimony potassium tartrate produced a significant decrease in life span at a daily dosage of about 0.3 mg Sb/kg.
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Animals poisoned by antimony compounds showed dyspnea, loss of weight, weakness, loss of hair, and evidence of cardiac insufficiency. Those that survived began to regain weight within 5-10 days (Bradley and Fredrick, 1941).
Absorption, Distribution, Metabolism, and Storage Ingestion of the compound usually leads to repeated vomiting, Thus, removal of the material from the gastrointestinal tract and inherently poor absorption combine to limit the amount of the compound reaching the tissues. Excretion is mainly urinary. It is much faster than that of arsenic and is almost complete in 72 hr (Osol et aI., 1967). The rate of excretion varies considerably in different species, being slower in mice and monkeys and more rapid in rats. The synergistic action of salts of tris(p-aminophenyl)carbonium on the toxicity of tartar emetic to Schistosoma mansoni was not accompanied by any significant effect on the distribution or rate of excretion of tartar emetic and must, therefore, have involved the parasite only (Waitz et al., 1965). Although the acute toxicity of soluble antimony is greater than that of soluble lead, the excretion of antimony is sufficientiy rapid that no significant storage occurs (Bradley and Fredrick, 1941). Following single or repeated doses of tartar emetic, there was no really marked accumulation of antimony in any organ, but the concentration was always greatest in the liver, and it reached an average of 65 ppm in hamsters kilted by repeated intraperitoneal injection (Gellhom et al., 1946). A similar relationship was found in dogs, where the concentration of antimony decreased in the following order: liver> thyroid> parasites> kidney cortex> other organs (Brady et aI., 1945). Mode of Action Antimony combines with sulfydryl groups including those in several enzymes important for tissue respiration. The antidotal action of BAL depends on its ability to prevent or break the union between antimony and vital enzymes (Sollmann, 1957). The most characteristic toxic effect is vomiting, which is largely reflex in origin. Although vomiting may occur after intravenous injection, a larger dose is required and the action is nore delayed than when the drug is given by mouth (Sollmann, 1957). Furthermore, the action is largely reflex even after injection, for the compound is excreted through the walls of the stomach (Osol et aI., 1967). The cause of death is essentially the same as that in acute arsenic poisoning. Serious injury almost always involves the gastrointestinal system and may involve the cardiovascular system, the kidneys, or other organs leading in any event to a condition of shock. The rapid excretion of antimony makes important sequelae or delayed death unusual. Pathology Rats killed by an intraperitoneal injection of tartar emetic showed degenerative changes of the myocardium, congestion of the glomeruli and degeneration of the tubules of the kidney, and extensive centrolobular hepatic necrosis. Injury to the heart was detected histologic ally in rats receiving daily
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doses of antimony too small to reduce the rate of growth. On the contrary, antimony produced no serious blood changes following either single or repeated doses (Bradley and Fredrick, 1941). 61.11.1.3 Toxicity to Humans Therapeutic Use When antimony potassium tartrate is rubbed on the skin in the form of an ointment, it produces little irritation at first but produces a pustular eruption if applied for long periods. This is due to decomposition of the double salt by the acid secretions of the follicles, leading to formation of the more irritant antimonous oxide and other compounds (Sollmann, 1957). Antimony potassium tartrate was once used as an emetic for treating patients poisoned by a wide variety of compounds. The drug still has some use as a diaphoretic or expectorant in certain cough syrups. It has been used for treating a number of tropical diseases, and it still is the drug of choice for treating infection by Schistosoma japonicum. It is not used to treat S. mansoni or S. haematobium infections because less toxic agents are effective [American Medical Association (AMA) 1977]. The dose of tartar emetic varies greatly according to these different uses. As an emetic, the dose was usually 30-60 mg. The small safety factor is indicated by the fact that in one case a dose of 130 mg proved fatal (Osol et aI., 1967). The dose in cough remedies or expectorants varies from 1 to 8 mg repeated two or three times daily. As an intravenous injection for treating S. japonicum infection in adults, a freshly prepared 0.5% solution should be injected extremely slowly on alternate days according to the following schedule: 8 ml initially increased by 4 ml with each subsequent dose until the 11th day (when 28 ml is given) and then 28 m1 on alternate days until a total of 500 ml (2500 mg) has been given or until side effects become severe. Each dose should be given 2 hr after a light meal and the patient should remain recumbent for 1 hr after treatment (AMA, 1977). These therapeutic doses may be compared with 5 mg/person/day, the rate of respiratory intake of antimony considered safe for workers exposed to various antimony compounds. Use of antimony potassium tartrate has been abandoned as an emetic to give patients, but it is still used as an emetic to combine with certain rodenticides to make them less harmful if they are accidentally consumed by people or pets. Presumably the compound is more stable in rodenticide formulations than ipecac, which now is considered a faster and safer emetic for patients. The doses of antimony potassium tartrate received from cough medicines generally cause no side effects. Excessive doses either by mouth or intravenously produce symptoms resembling those of acute or subacute arsenic poisoning. In fact, the resemblance is so close that a distinction may be impossible unless the cause is known as the result of history or chemical analysis. There is, however, a tendency for antimony to produce earlier more profuse vomiting, less systemic absorption, more rapid excretion, and thus a shorter course without severe neurological sequelae. Symptoms include a metallic taste,
extreme nausea, copious vomiting, frequent hiccough, burning pain in the stomach, colic, frequent stools and tenesmus, fainting, bradycardia often with EGG irregularities, hypotension, difficult and irregular breathing, cutaneous anesthesia, convulsive movements, painful cramps in the legs or joints, jaundice, anuria, prostration, and death (Osol et aI., 1967; AMA, 1977). Laboratory Findings Antimony is found in all tissues and organs of "unexposed" people. Autopsy data on Japanese (Sumino et aI., 1975) indicate a uniform distribution throughout the body with concentrations generally between 0.01 and 0.05 J.lg Sb per gram wet weight. The average hair level was 0.12 J.lg/g but others have reported a range of 0.08-6.6 J.lg/g (Liebscher and Smith, 1968). The average blood level in the Japanese study was 0.016 J.lg/g. Following intravenous injection of tartar emetic, only about 2.5% of the dose was recovered from the urine during the first 24 hr. The rate of recovery fell only very slowly in subsequent days (Boyd and Roy, 1929). Analysis of 315 electrocardiograms taken on 100 patients during various stages of treatment with tartar emetic and Fuadin for schistosomiasis revealed one or another abnormality in 1199% of the patients. Abnormalities in one or more leads included increased amplitude of P waves, a fusion of ST segment and T wave, decreased amplitude of T wave, and prolongation of the Q-T interval. The duration of these changes was variable but was noted up to 2 months after treatment stopped (Schroeder et al., 1946). Pathology At autopsy of persons who have died following ingestion of a large dose, ulcerations usually are found in the esophagus and stomach but not in the intestine. In subacute cases, fatty degeneration of the liver, kidney, and heart may be present; degenerative changes in the nervous system are less common. Persons who died, perhaps as a result of individual susceptibility, following the usual, intravenous, therapeutic dose of tartar emetic showed marked degeneration of the liver, some necrosis of the renal tubular epithelium, and varying degrees of hemorrhage (McKenzie, 1932). Treatment of Poisoning BAL is effective in treating poisoning by antimony potassium tartrate, but this is not true of all organic antimony compounds.
61.12 ARSENIC Arsenic is a metalloid. It belongs to the same group as phosphorus. Each of these elements combines with hydrogen to form a highly toxic gas, arsine and phosphine, respectively. Arsenic is followed in the periodic table by selenium, an essential element which also finds applications as a pesticide. Arsenic compounds occur in many rocks and thus find their way into soil, water, and food, being especially high (3170 ppm) in some seafood (Monier-Williams, 1949). Arsenic is a normal constituent of the human body, a fact recognized
61.12 Arsenic
since the work of Gautier (1899). Two laboratories have independently reported that arsenic, at least in the inorganic form, is an essential nutrient in minipigs and goats (Anke et aI., 1976, 1978) and in rats and chicks (Nielsen et al., 1975; Uthus and Nielsen, 1985). If an element is found to be essential in animals, it is highly probable it is essential in humans. However, we lack knowledge of a biochemical mechanism and physiologic role of arsenic. For reviews of arsenic, see WHO (1981), U.S. EPA (1983), Ishinishi et al. (1986), and ATSDR (1998). 61.12.1 ARSENICAL PESTICIDES 61.12.1.1 Identity, Properties, and Uses Compounds and Their Characteristics Elementary arsenic forms two oxides: the trioxide, AS203, and the pentoxide, AS20S. Arsenic trioxide (trivalent) reacts with water to form arsenous acid, H3As03, which is known only in solution and forms three series of salts: orthoarsenites (e.g., Na3As03), metaarsenites (e.g., NaAs02), and pyroarsenites (e.g., Nll4As20S). Arsenic pentoxide (pentavalent) reacts with water to form three acids that may be isolated: orthoarsenic acid, H3As04; metaarsenic acid, HAs03; and pyroarsenic acid, H4As207. These acids form the corresponding salts: orthoarsenates, metaarsenates, and pyroarsenates. A few organic arsenic compounds also are used as pesticides. Although a great many arsenicals have had some use as pesticides, the ones shown in Table 61.4 are of the greatest importance. It is beyond the scope of this book to discuss the highly toxic compounds of arsenic (arsine and certain war gases) or those of
Table 61.4 Some Arsenical Pesticides Name
Synonym
Formula
Arsenic trioxide
White arsenic
AS203
Sodium arsenite
Na3As03. NaAs02
Paris green
Cooper aceto-meta-
Lead arsenate
Acid lead arsenate
N'4As20s CU(CH3COOh 3Cu(As02h
arsenite PbHAs04
Standard lead arsenate Dilead arsenate Basic lead arsenate
Lead hydroxyarsenate
Calcium arsenate Dimethylarsinic acid
A complex mixture Arsan® Cacodylic acid Phytar® Silvisar®
Disodium methyl arsenate
(CH3)zAsO(OH)
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low toxicity employed therapeutically in medicine and agriculture (drugs and feed additives). Formulations, Uses, and Production Arsenites are more soluble and more rapidly toxic than corresponding arsenates; therefore, arsenites are used as rodenticides and herbicides and in insecticidal baits. Paris green, although an arsenite, may be applied to foliage, but inorganic arsenates are less phytotoxic and, therefore, preferred for application to crops as insecticides. Some organic arsenates (dimethylarsinic acid and disodium methyl arsenate) are herbicides. The use of arsenical insecticides in agriculture decreased greatly following the introduction of DDT and later poisons, but the use of arsenical herbicides has increased. Starting about 1975, the uses of arsenic compounds as wood preservatives began to grow. By 1980, 70% of the arsenic consumed in the United States was used by the wood preservative industry (Loebenstein, 1994). 61.12.1.2 Toxicity to Laboratory Animals Symptomatology The acute effects of arsenic in animals are similar to those observed in humans (see Section 61.12.1.3). The degree of irritation of the gastrointestinal tract involved in poisoning by arsenic trioxide depends on its purity. A commercial preparation that was 97.8% pure was much more irritant yet slightly less toxic than a sample of greater man 99.999% purity (Harrisson et al., 1958). Dogs fed sodium arsenite at a rate of about 2.7 mg/kg/day showed anorexia, listlessness, and weight loss leading to cachexia and eventually death (Kiyono et al., 1974). Mild to moderate anemia was present prior to death. Neither skin lesions nor polyneuropathy apparently has been observed in experimental animals. Thus, except for weight loss and anemia in some species, arsenic poisoning as it occurs in humans is poorly reproduced in animals. Dosage Response The literature indicates a tremendous variation in the oral toxicity of arsenic compounds. Table 61.5, prepared from a careful study of arsenic trioxide by Harrisson et al. (1958), shows that the variation is real but can be accounted for largely by whether the compound is administered dry or in solution and to a lesser extent by the purity of the sample and by the species, strain, and weight of the experimental animals. Harrisson and his colleagues found no significant difference in the response of males and females. They acknowledged that arsenic of coarse grind might be less easily dissolved and hence less toxic than fine powder. However, in the particular samples they studied, the coarser preparation was actually more toxic because of its greater purity. Thus a combination of variables must be taken into account. The LD 50 values of suspensions of calcium arsenate (298 mg/kg) and of Paris green (100 mg/kg) in rats are similar to that of powdered arsenic trioxide. The LD 50 value of lead arsenate is distinctly higher (l050 mg/kg) (Gaines, 1969). Long-term feeding of dogs (Byron et at., 1967) or monkeys (Heywood and Sortwell, 1979) causes death at a dosage rate
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Table 61.5 Effect of Purity and Dosage Form of Arsenic Trioxide on Its Oral Toxicity in Male Animals of Different Species, Strains, and Weightsa LD 50 value (mg Aslkg) Experimental
Weight
Dosage
97.8%
99.999%
animal
(g)
form
AS203
AS203
± 2.1
39.4 ± 4.7
Swiss mouse
20-25
Solution
Swiss mouse
35-40
Solution
47.6 ± 3.3
Solution Solution
25.8 ± 1.8 32.4 ± 2.3
C57HI6 mouse Dba2 mouse C3H mouse Sprague-Dawley
42.9
Solution 125-200
Solution
125-200
Dry
25.8 23.6 ± 1.4
15.1
± 1.8 ± 1.8
rat Sprague-Dawley
214.0 ± 8.4
145.2 ± 8.7
rat aFrom data of Harrisson et al. (1958).
of about 3 mg/kg/day as arsenate or arsenite. A marked enlargement of the common bile duct is a characteristic feature. A no-adverse-effect level in rats is about 1.6 mg As/kg/day as arsenite and about twice as high for sodium arsenate. Dogs appear to be somewhat more susceptible than rats. Neither skin lesions nor polyneuropathy has been observed in experimental animals.
Absorption, Distribution, Metabolism, and Excretion Arsenic is absorbed chiefly by the gastrointestinal and respiratory tracts. However, some is absorbed by the intact skin, and systemic illness may follow application of arsenical ointment to eczematous skin. Gastrointestinal absorption of the soluble trivalent compounds of arsenic appears to be high (Buchet et aI., 1981a, b; Charbonneau et al., 1978; Coulson et al., 1935; Crecelius, 1977; Freeman et al., 1993; Mappes, 1977; Marafante and Vahter, 1987; Marafante et al., 1987; Vahter, 1981; Vahter and Norin, 1980; Yamauchi and Yamamura, 1985) in both animals and humans. The organic form of arsenic present in shellfish and other marine foods is also well absorbed from the gastrointestinal tract. Soluble arsenates are also well absorbed. Arsenic is rapidly cleared from the bloodstream except in the case of rats, where its methylated derivatives bind to hemoglobin. In fact, both trivalent and pentavalent forms of arsenic are rapidly methylated, so their distributions are somewhat similar. Initial accumulation is in liver, kidneys, and lung, from which arsenic is cleared rapidly; long-term accumulation is in skin, squamous epithelium of the upper gastrointestinal tract, thyroid, the lens of the eye, and the skeleton. It is also accumulated in hair. Human autopsies have also confirmed accumulation in keratinized tissues and the skeleton. Both trivalent and pentavalent forms of arsenic readily cross the placenta (for review, see Clarkson et al., 1983). The pentavalent form accumulated in the skeleton of the fetus in late gestation. Little arsenic crossed the blood-brain barrier.
Most arsenic in tissues is found to be protein bound. Trivalent arsenic is well known to bind to tissue -SH groups (Webb, 1966). Both trivalent and pentavalent arsenic can be interconverted by oxidations-reduction reactions in mammalian tissues. The biochemical mechanisms are not known. Monomethyl and dimethyl derivatives are produced in experimental animals and humans presumably by methylation of the trivalent form. Substantial species differences exist in methylation rates and in the relative proportions of the monomethyl and dimethyl forms. Arsenobetaine, the organic form of arsenic in marine organisms, is not further metabolized in the body. Methylation of arsenic may be regarded as a detoxication process, as the methylated derivative is less toxic than the inorganic form. Factors affecting the methylation rate are therefore important. There is evidence that the efficiency of methylation decreases at higher doses. Populations having certain dietary deficiencies (e.g., in methionine) may be more susceptible to arsenic poisoning. After absorption of either the trivalent or pentavalent form, urinary excretion is the dominant pathway. In chronic exposure, this route accounts for 60-70% of total excretion. Methylation rate is important, as the methylated species is more rapidly excreted than the inorganic form. For example, urinary excreition is lowest in the marmoset monkey, which is unable to methylate arsenic (Vahter et al., 1982). Excretion also takes place by feces, skin, hair, and milk. Fecal excretion may be preceded by extensive enterohepatic recirculation according to data for experimental animals (Klaassen, 1974).
Tolerance It was reported during the nineteenth century that the mountaineers of Styria and certain other regions ate arsenic once or twice a week as a tonic and thus accustomed themselves to doses of 400 mg or more per day. Their blood and urinary arsenic levels and even the absorbability of the material they ate are apparently unknown. However, the reality of tolerance to inorganic arsenic has been demonstrated by tests in dogs (Cloetta, 1906), rabbits (Cloetta, 1906), and rats (Joschimoglu, 1916; Norris and Elliott, 1945). Tolerance does not depend on decreased absorption from the gastrointestinal tract, for it may be induced by intraperitoneal injection (Norris and Elliott, 1945). Biochemical Effects Arsenic inhibits pyruvate oxidase and the phosphatases. The blood level of pyruvate increases in poisoned animals or people. There is a reduction of tissue respiration leading to a wide range of functional and some morphological changes. Many other sulfhydryl-containing enzymes are involved also, and it is impossible to assign relative importance to them. However, studies on antidotes have made it clear that chemical reaction of trivalent arsenic with sulfhydryl groups, including those in susceptible enzymes, is the biochemical lesi on (Peters, 1952). Further studies by Peters' group (for review, see Clarkson, 1983) revealed that the inhibition of pyruvate oxidase was due to the binding of trivalent arsenic, as the oxyanion, to neighboring -SH groups of a-lipoic acid. Pentavalent arsenic, also as the oxyanion, is able to substitute for phosphate
61.12 Arsenic
anions in the cell's transport and enzymic processes. The result is replacement of phosphates by arsenate in high-energy phosphorylated substrates, leading to uncoupling of oxidative phosphorylation reactions (for a review, see Jennette, 1981). Altered biochemical functions in the liver mitochondria of rats fed arsenic are associated with considerable physical distortion of the mitochondrial membrane components. The same degree of biochemical change is associated with less electron microscopic change in the mouse (Fowler and Woods, 1979). Although arsenic inhibits many enzymes, it increases the activity of liver microsomal enzymes. Arsenic trioxide at a dietary level of 1000 ppm for 15 days induced several enzymes, and even at 100 ppm hexobarbital sleeping time was decreased. Rats develop some tolerance to arsenic, and its toxicity is decreased by phenobarbital (Kourounakis et al., 1973; Wag staff, 1972).
Effects on Organs and Tissues The chief pharmacodynamic action is dilatation and increased permeability of the capillaries. This action is strongest in the intestines, regardless of route of absorption. Local action on capillaries causes congestion, stasis, thrombosis, ischemia, and necrosis. Such necrosis extends into the bone in some instances. Injury to the kidneys is due primarily to capillary injury, but there is always some injury to the epithelium. Initial injury to the nervous system is also based on circulatory disturbance, but later there is direct injury to the nerve cells. Injury to the liver by arsenic is usually minor. A few single cases or clusters of cases involving cirrhosis have been described. In many of these cases, alcohol was a complicating and perhaps critical factor in the etiology. The cause of death depends on the size of the dose and, therefore, on the speed of action. If death occurs within a day or two, it is caused by shock characterized by a severe fall in blood pressure. Vasodilatation is most marked in the splanchnic area, but death occurs in experimental animals even if the intestines are tied off early. The action on capillaries is direct, for it follows perfusion of excised organs with solutions of arsenic. The dilatation is not due to a loss of contractility; the vessels continue to react to splanchnic stimulation until very late, and they react to epinephrine even later (Loeb, 1912). If death is delayed 3-14 days, it is caused by dehydration, electrolyte imbalance, and a more gradual fall in blood pressure. Most animal studies testing for carcinogenicity of arsenic have been negative (for review, see U.S. EPA, 1983). Studies on rats should be viewed with caution in view of the unusual metabolism of arsenic in this species. Effects on Reproduction and Development Few studies have been made on the effect of arsenic on reproduction and development. Treatment of pregnant animals with high doses of inorganic arsenic results in fetal resorptions and, in surviving offspring, defects in the genitourinary tract. Studies to date indicate that effects on the fetus are not produced at dosages below those causing maternal toxicity (for reviews, see ATSDR, 1998; Clarkson et al., 1983).
1395
Pathology The injury produced by arsenic compounds in most species is not characteristic although the enlargement of the rat's common bile duct by sodium arsenite and sodium arsenate is indeed striking. Using a strain of rat that occasionally showed spontaneous enlargement of the duct, Byron et al. (1967) found that the duct enlarged in 45 of 49 rats fed sodium arsenite at a dietary As level of 250 pptn and in 42 of 48 rats fed sodium arsenate at a dietary As level of 400 ppm. Lower dietary levels produced a lower incidence of enlargement. Some of the ducts measured more than 7 mm in diameter, and their walls were as much as 10 times thicker than normal. The condition was reminiscent of that produced in rats by dibutyltin dichloride and a few nonmetal compounds. No parallel condition in humans has been recognized. Treatment of Poisoning in Animals Dimercaptosuccinic acid increased the rate of excretion of radioactive arsenic by poisoned rats and increased their survival compared with controls (Okonishnikova, 1965). 61.12.1.3 Toxicity to Humans Experimental Exposure
See Laboratory Findings.
Therapeutic Use Arsenic in the form of arsenic trioxide, potassium arsenite solution (Fowler's solution), or arsenious acid solution formerly was used extensively as a tonic in treating nutritional disturbances, neuralgia, rheumatism, arthritis, asthma, chorea, malaria, syphilis, tuberculosis, diabetes, skin disease, and every kind of blood disturbance. Some skin conditions were treated locally, and arsenic was used to destroy some superficial epitheliomas. In fact, both Hippocrates and Galen recommended a naturally occurring arsenic disulfide for treating ulcers, and cautery of tumors by arsenic salts was practiced by Avicenna in the tenth century and by Guy de Chauliac in the fourteenth. Fowler's solution was used in some treatment of leukemia until rather recently (Sollmann, 1957). Experience in the systemic use of inorganic arsenicals showed that their prolonged administration at the usual rate of 0.04-0.09 mg/kg/day frequently produced mild poisoning. The various systemic and cutaneous effects described under Use Experience have been produced by arsenical therapy also. Accidental and Intentional Poisoning For many years, arsenic has been the most important single cause of accidental deaths associated with pesticides in the United States. It caused 36, 26, 29, 31, and 17% of such cases in 1956, 1961, 1969, 1973, and 1974, respectively (Hayes and Vaughn, 1977). Accidental poisoning by arsenic pesticides is almost always and often involves children. Experience in many other countries, for example, Poland (Brodniewicz and Szuber, 1960), Australia (Southby, 1965), and New Zealand (Bailey, 1964; Kennedy, 1961), has indicated the disproportionate importance of arsenical pesticides as a cause of pesticide poisoning. Contamination of well-water by arsenical pesticides has been a cause of outbreaks of poisoning in Russia (Khasanov,
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1971; Planques et al., 1960) and possibly in Hungary (Nagy et aI., 1975a, b). The first symptom is often a feeling of constriction of the throat, followed by difficulty in swallowing and epigastric discomfort. Abdominal pain and vomiting often start within an hour of ingestion, although the onset may be delayed until the next day, particularly if foul play is involved and the dosage is controlled. Death may result from a severe fall in blood pressure and collapse as in "dry" cholera. Generally, death is delayed for 1.5-3 days after onset and sometimes as much as 14 days. In this event, death follows vomiting and profuse, watery, painful diarrhea; the clinical picture is similar to cholera because of the "rice water" character of the stools and the great dehydration of the patient. Although arsenic is not truly corrosive, the extreme distention of the capillaries may lead to their rupture and thus to ecchymoses or even bleeding into the stomach or intestine and eventually, but infrequently, bloody vomiting and diarrhea. The patient has great thirst and difficulty in swallowing and articulation. Although the abdomen is painful, it is not tender. Descriptions of cases of this kind can be found in great profusion in the older literature, for example, Orfila (1814-1815) and Christison (1845). If the patient survives the acute phase, skin eruptions, a moderate depression of blood cells, and polyneuropathy may appear. The pseudoinflammatory reaction may extend to the conjunctiva, gums, mouth, pharynx, and bronchi, and it may persist in the gastrointestinal tract. Thus, conjunctivitis, rhinitis, cough, or bronchopneumonia may be present. Dermatitis is especially prominent in the palms and soles and other areas subject to pressure. White transverse bands in the nails frequently appear in about 6 weeks and may accompany polyneuropathy, which often appears 1-3 weeks after ingestion. The bands migrate to the free edge of the nails at approximately the normal rate of about 5 months from matrix to edge. Although these lines, first described by Mees (1919), may have other causes, their occurrence in conjunction with polyneuropathy is almost pathognomonic of arsenic poisoning. Skin eruptions may progress to exfoliative dermatitis, especially following exposure to organic arsenicals. Peripheral circulatory difficulty characterized by blanching or flushing of the skin may be present, especially in the fingers. Some authors (Leng-Levy et aI., 1969) have emphasized the importance of cardiac involvement in the total syndrome, and the frequency of ECG changes would support this view. Among various forms of cardiac involvement, Wang and Mazzia (1969) emphasized the unusual susceptibility of persons severely poisoned by arsenic to ventricular fibrillation and cardiac arrest during anesthesia. Arsenical polyneuropathy involves paresthesia, pain, burning, and tenderness of the affected limbs (Heyman et al., 1956). Trouble in walking or in grasping objects may at first be secondary to pain. Later the pain may disappear with astonishing suddenness, but the patient may be left with loss of (a) proprioception, (b) other sensory functions, and (c) motor function. The relative importance of different changes varies from case
to case. The legs and arms are affected about equally, although difficulty usually is first noticed in the feet. Muscle atrophy may be pronounced (McCutchen and Utterback, 1966). Tendon reflexes are, of course, weak or absent. Mental confusion may be present and may be evident as an apparent inability of the patient to grasp the seriousness of the condition or the possible implications of foul play. Tsuchiya (1977) reviewed an outbreak of arsenic poisoning that differed from others in two ways: most of the victims were less man 12 months of age, and the compound involved was found to be pentavalent. The arsenic was a contaminant of sodium phosphate intentionally added to milk as a stabilizer. Different lots of the powdered milk product contained 0-34 ppm of arsenic. Symptoms began early in the summer of 1955; following investigation, sale of this milk powder was banned on August 24, 1955, Many of the signs and symptoms (diarrhea, vomiting, anorexia, and others) were typical of acute poisoning by trivalent arsenic but the prominence of many other difficulties (dermatitis, loss of hair, melanosis, leukoderma, and irritation of the eyes and upper respiratory system) was typical of subchronic or chronic poisoning, as one might expect with exposure lasting several months. However, neuritis was not observed, and electrical measurements indicated no dysfunction of the peripheral nervous system. In view of the fact that there were a total of 12,131 cases and 130 deaths, one would have expected hundreds if not thousands of cases of polyneuropathy if trivalent arsenic had been involved. Other differences from ordinary arsenic poisoning were the presence of fever and of liver swelling in the majority of cases. The contrast is especially remarkable in view of the demonstrated interconversion of the two valence forms. The poisoned infants were treated with BAL; rapid weight gain was considered the most striking benefit. Survivors were examined in June 1956, and it was concluded that most were normal-another contrast in view of the severity of the initial illness. It seems clear that the majority of cases of arsenic poisoning have been associated with trivalent compounds, but in some instances the identity-and thus the valence-of the offending compound has not been reported. The large outbreak just mentioned suggests that there are two distinctly different forms of arsenic poisoning, depending on valence. Arsine (AsH3) is not used as a pesticide and ordinarily plays no part in the danger of arsenical pesticides. However, arsine was the cause of the illness of eight children and one adult who 48 hr earlier had helped to clean a dip vat that had been charged with chlordimeform and monobasic calcium phosphate. All of the patients showed albuminuria, and some showed hematuria, abdominal pain, dysuria, and headache. A diagnosis of arsine poisoning was made in the absence of any known source of arsenic. However, the next day, the farmer recalled that an arsenical dip had been used in the vat 2 years earlier, and arsenic later was measured in the acid mud (pH 6.5) of the vat and in the urine of the patients. Fortunately (and consistently with the delayed onset), all the cases were mild and the recoveries complete (Rathus et aI., 1979).
61.12 Arsenic Use Experience Poisoning has not been a significant problem in pesticide applicators; cases reported in vineyard workers in Germany may have involved the drinking of contaminated wine. Poisoning is well known from other occupational sources, including the manufacture of pesticides. However, according to Buchanan (1962), who reviewed the industrial toxicology of arsenic in detail, even the use of arsenic compounds in industry has not proved an important cause of occupational morbidity and mortality. Poisoning that is caused by repeated occupational exposure usually involves the frequently insidious onset of loss of appetite, weight loss, weakness, nausea, alternating diarrhea and constipation, colic, peripheral neuropathy, dermatitis, some loss of hair, giddiness, and headache. In general, gastrointestinal involvement is less and dermal involvement is greater than in poisoning caused by one or a few doses. Prolonged exposure may lead to gradual mental and physical deterioration and a state of cachexia suggestive of a malignant or endocrine disorder. Cyanosis of the face may be present as a result of statis in the injured capillaries rather than systemic anoxia. The dermatitis may be erythrematous, pustular, or even ulcerative. Burning and itching may be present and there may be serous discharge. With most arsenic compounds, the skin lesions tend to be most marked in the area of greatest contact. They are considered mainly the result of direct toxic action. The eruption may involve the face, eyelids, conjunctivae, or even cornea. There may be irritation of the nose, pharynx, and trachea leading to hoarseness and chronic cough. Perforation of the nasal septum has occurred. A highly characteristic dermatitis confined to the scrotum, inguinal area, and nasolabial folds may follow moderate occupational exposure to Paris green. The lesions begin with erythema, frequently become eczematous and weeping, and may start to heal with the formation of a black scab. A sensitization reaction may be involved because the distribution does not correspond to the distribution of insecticide on the skin, and the dermatitis generally occurs in the absence of typical systemic poisoning. In less acute cases, hyperkeratosis, hyperhidrosis, or melanosis may occur. These changes are considered evidence of chronic systemic action. The hyperpigmentation is most marked on surfaces exposed to light; it does not extend to mucous membranes. There may be a speckled depigmentation of pigmented areas giving the so-called raindrop appearance. In persons exposed to sufficient arsenical dust, the onset of illness is characterized by dyspnea and oppression and pain in the chest. These symptoms may be followed by nausea, diarrhea, and other usual signs of poisoning but generally of a mild degree. The polyneuropathy that may follow repeated exposure to arsenic resembles that seen in persons who survive one or a few doses. Disturbance of sight, taste, and smell may occur. There may be disturbance of bladder function. It is said that arsenical polyneuropathy is particularly severe and unremitting in chronic alcoholics.
1397
There is no doubt that compounds similar or identical to those used as pesticides have caused skin cancer in humans. The earliest evidence comes from patients treated with potassium arsenite (Fowler's solution) (Fierz, 1966; Hutchinson, 1887; Minkowitz, 1964; Neubauer, 1947; Sanderson, 1963; Sommers and McManus, 1953) and from people who consumed water with a naturally high arsenic content (Hsueh et al., 1995; Neubauer, 1947; Tseng et aI., 1968). In these instances, intake of arsenic was oral. The situation was somewhat less clear in connection with certain winegrowers in Germany and France (Denk et al., 1969; Galy et aI., 1963a, b; Latarjet et aI., 1964; L'Epeeetal., 1973; Liebegott, 1952; Roth, 1957a). Whereas the initial impression was that exposure to lead arsenate was dermal and respiratory-as, indeed, it was-the critical intake of arsenic may have been oral and associated with contamination of wine by arsenic, making the exposure of the winegrowers entirely similar to that of people who drank arsenic in Fowler's solution or in water. Two histopathological types of skin cancer have been associated with arsenic-squamous carcinomas in the keratin areas and basal cell carcinomas. Skin cancers caused by arsenic differ from those resulting from ultraviolet light by occurring in areas of the body not exposed to sunlight, e.g., soles of the feet. The appearance of skin cancer is preceded by a characteristic sequence of changes in skin epithelium. Hyperpigmentation is followed by hyperkeratosis, which, histologically, has been described as keratin proliferation of a verrucose nature with derangement of the squamous portions of the epithelium. The latent period for initiation of exposure to appearance of skin cancer ranges from 6 to 50 years when arsenic was used medicinally, e.g., Fowler's solution. When exposure was from contaminated drinking water, the shortest latency was about 24 hours. There is now mounting evidence that ingestion of arsenic may increase the risk of internal cancers as well. In large-scale epidemiological studies in Taiwan, where exposure is mainly from drinking water, clear associations are reported with tumors of the bladder, kidney, liver, and lung (Chen and Wang, 1990; Chen et aI., 1985, 1986, 1988a, b, c; 1992; Chiang et al., 1988; Chou et aI., 1995; Wu et aI., 1989). Increased risk of bladder cancer has been reported for populations in Argentina (Hopenhayn-Rich et aI., 1996) and in Chile (Moore et aI., 1997) from arsenic in drinking water. Cuzick et al. (1992) reported on follow-up findings of a cohort in Lancashire who had been tested with Fowler's solution (potassium arsenite) during 1945-1969. A significant excess bladder cancer mortality was found. A considerable body of evidence now associates lung cancer with occupational exposure to arsenic. Perry et al. (1948) in the United Kingdom and Ott et al. (1974), Baetjer et al. (1975), and Mabuchi et al. (1979) in the United States found an excess risk of lung cancer associated with exposure to arsenic in industries manufacturing arsenic pesticides. A number of studies of copper smelters in the United States, Sweden, and Japan (for review, see U.S. EPA, 1984) have also implicated arsenic as a cause of lung cancer.
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Arsenic has been mentioned in connection with the carcinogenicity of cigarettes (Satterlee, 1956), and it is true that residues of arsenic in tobacco increased from an average of 16.2 ppm in 1932-1934 to an average of 42 ppm in 1950-1951 (Satterlee, 1956). However, the concentration of arsenic in tobacco decreased later, so that in 1958 the average concentration in 17 brands was only 6.2 ppm (Guthrie et aI., 1959). In a few instances, people with arsenical keratoses or other late effects of arsenic ingestion have presented with a wide variety of internal cancers, and the possibility that the internal cancers were caused by arsenic has been raised (Atkinson, 1969; Brady et aI., 1977; Roth, 1957b). However, more recent reviews (e.g., U.S. EPA, 1984) regard the evidence as equivocal.
Dosage Response A dose of 5-50 mg of arsenic trioxide is toxic. A dose of 128 mg (about 1.8 mg/kg) is said to have proved fatal to an unhabituated adult, but recovery has occurred after much larger doses. In fact, recovery occurred in one case after what was thought to be 20,000 mg (Co sic and Kusic, 1966). Very young children may be more susceptible. Study of 291 exposure incidents involving a single kind of ant bait in the form of a bottle containing one-half ounce (15 ml) of sweetened 0.61 % sodium arsenite solution permitted a poison control center to conclude that a dosage of less than 1 mg/kg can cause serious illness in a child and 2 mg/kg can cause death (Peoples et al., 1977). The effectiveness of arsenical rat poisons varies greatly with the grind of the powder; very fine powders approach the toxicity of solutions containing an equivalent amount of arsenic. Thus, the ease of absorption influences the toxicity to a marked degree. The repeated dose necessary to produce poisoning is less well known. The "therapeutic" dose of arsenic trioxide (12 mg three times daily) that used to be employed as a tonic frequently led to mild poisoning. Two milligrams three times a day for an adult is a rate of only about 0.06 mg/kg/day expressed as As. The Agency for Toxic Substances and Disease Registry (ATSDR, 1998) has estimated an upper no-adverseeffect level for skin lesions (blackfoot disease, hyperkeratosis, hyperpigmentation) to be 0.8 J-Lg As/kg/day. Cancer associated with arsenic has not been reported in persons exposed to air concentrations of 0.1 mg/m 3 or less even for long periods. It has been estimated that exposure to air concentrations of 50 J-Lg As/m3 occupationally for 24 years would result in a 200% excess risk of lung cancer (WHO, 1981). Data from a population in Taiwan (Tseng, 1977) led a WHO Task Group on Environmental Health Criteria for Arsenic (WHO, 1981) to conclude that the lifetime risk for skin cancer due to arsenic in drinking water is about 5% for a total dose of 10 g in an assumed life span of 70 years.
Absorption, Distribution, Metabolism, and Excretion Arsenic has been measured in human tissue and body fluids in both "nonexposed" and exposed individuals. In people exposed to normal environmental levels, hair and nails have the highest
concentration with skin and lung next in order (Liebscher and Smith, 1968). Unfortunately, only total arsenic is reported, and we do not know the levels of specific forms of arsenic. Tissue levels change rapidly after a single dose of arsenic. In six human subjects who ingested a tracer dose of arsenate labeled with the 74 As isotope, more than 5% of the dose was excreted in urine within 5 days (Tarn et al., 1979). In three people who ingested 500 J-Lg of arsenic as arsenate in drinking water, 45% of the dose was excreted in urine within 45 days. Ingestion of the same amount of arsenic as the mono- or dimethyl derivative resulted in about 75% of the dose being excreted in urine within 4 days (Buchet et aI., 1981a). Arsenic is rapidly cleared from blood in humans (for review, see U.S. EPA, 1984). If exposure is continuous, blood levels should quickly attain a steady state and should be proportional to average daily intake. However, it was not possible to detect any increase in blood levels in individuals exposed to arsenic in drinking water until the level in drinking water rose to 100-300 J-Lg As/I (Valentine et aI., 1979). Most drinking water samples are below 10 J-Lg As/I (WHO, 1981). Normal concentrations in blood of nonexposed adults are in the range 1---4 J-Lg As/I. Individuals exposed to elevated arsenic in drinking water may have blood levels up to 50-60 J-Lg Asll (for review, see Vahter, 1988). When considering blood levels in nonexposed people, the possibility of intake of organic arsenic in seafood should be taken into account. Ingestion of shrimp with a high natural arsenic content causes blood levels to rise to 50 J-Lg Asll within 2 hr (Vahter, 1988). Levels of total arsenic in urine in nonexposed people are in the range 5-50 f.Lg Asll. However, ingestion of seafood can increase urinary levels up to 1000 J-Lg Asll. Urinary arsenic in occupational exposure is usually in the range of hundreds of micrograms per liter-values that can easily be confounded by ingestion of seafood. Vahter (1988) has reviewed published studies indicating that long-term daily ingestion of drinking water at 100 J-Lg Asll results in an average urinary concentration of 60 J-Lg As/I. Vahter (1988) goes on to suggest that a better index of exposure to inorganic arsenic may be obtained by measuring the urinary concentrations of inorganic arsenic and the two metabolites mono- and di-methyl arsenic. Hair concentrations of total arsenic in nonexposed individuals are usually less than 2 J-Lg As/g. In subjects occupationally exposed or ingesting arsenic in contaminated drinking water, hair concentrations can rise to 50 J-Lg As/g. The methylated derivatives of arsenic and the organic arsenic in seafood are not accumulated in the hair (Vahter, 1988), so total arsenic in hair should reflect the body levels of inorganic arsenic. Inorganic arsenic is accumulated into hair at the time of formation of the hair strand. Once incorporated into the strand, its concentration remains unchanged. Thus the "segmental" analysis of hair section by section measured from the scalp should quantitatively reveal the sequence of part absorption (Smith, 1964). The use of hair as a monitor of absorbed inorganic arsenic is limited by the possibilities of external contamination from water, soaps, and shampoos. No satisfactory washing procedure
61.13 Phosphorus has been developed for removing external arsenic from the hair sample (Atalla et aI., 1965). Because trivalent arsenic is known to bind selectively to the -SH groups of keratin in hair, it is likely that accumulation in hair may depend not only on inorganic versus organic forms of arsenic but also on the oxidation state of inorganic arsenic. Concentrations of arsenic in nails in nonexposed persons are in the range of 0.01-3 ).lg As/g with an average of about 0.3 ).lg As/g. Values over 100 ).lg As/g have been reported in cases of chronic arsenic poisoning (see Vahter, 1988). Arsenic is probably incorporated into the growing tissue of nails, so its concentration may vary along the growth direction of the nail. The concentrations of arsenic in the organs of people in the general population and in persons killed by arsenic are shown in Table 61.6. There should be no difficulty in confirming a diagnosis of poisoning by analysis of organs if the deceased person had no occupational exposure to arsenic and if arsenical embalming fluid does not complicate the picture. Apparently no information is available on the concentration of arsenic in the organs of occupationally exposed persons who died of unrelated causes. Thus, it is not certain whether the analysis of organs always offers clear evidence about whether the death of a person with occupational or other special exposure to arsenic was or was not caused by the material. In any event, levels less than 0.2 ppm (dry weight) should be ignored; levels higher than 0.2 ppm should be evaluated carefully, although they occasionally occur in normal persons. Pathology In acute poisoning, erosion and inflammation of the stomach and upper intestinal tract may be marked. The liver may show degenerative lesions. Unless death is very rapid, the severe dehydration produced by acute poisoning gives the body an emaciated appearance, even though a normal amount of fat may remain. The alimentary canal shows a large amount of Table 61.6 Concentration of Arsenic in Human Organs Persons killed
fluid, shreds of mucus, and false membrane in the absence of marked corrosion-a picture similar to that in cholera. Central necrosis may be found in the follicles of the spleen and the tonsils. The body may decay more slowly than would be expected in the same amount of time at the same temperature. When death follows long repeated exposure, there is usually fatty degeneration of the myocardium, kidney, and liver, and the liver is often enlarged. Cachexia may be marked, and severe edema may be present. The nerves may show demyelinization and disintegration ofaxons. Treatment of Poisoning If ingestion is suspected, the stomach should be emptied by vomiting or lavage with warm water and activated charcoal followed by a saline cathartic. BAL (dimercaprol) is a specific antidote. Dehydration should be combated with saline infusions, guided, where possible, by laboratory studies. If available, an artificial kidney may be used. According to Lasch (1961), much arsenic can be removed by hemodialysis. The diet should be liquid and supplemented with vitamins. D-Penicillamine has been found at least as effective as BAL fof treating human cases, and it has been recommended in all situations in which oral administration is appropriate (Peterson and Rumack, 1977). Soviet authors generally prefer unithiol to BAL for treating poisoning by arsenic compounds other than arsine (Mizyukova and Petrun'kin, 1974), and they are almost certainly correct on the basis of both effectiveness and safety. The relative value of unithiol and D-penicillamine is uncertain.
61.13 PHOSPHORUS Compounds of phosphorus are a major constituent of protoplasm and essential to life. However, the element itself in its reactive white or yellow form is highly toxic. Red phosphorus and the less common black phosphorus are much less reactive and, therefore, relatively harmless. 61.13.1 PHOSPHORUS
General population
by arsenic
Sample
(ppm)
(ppm)
61.13.1.1 Identity, Properties, and Uses
Brain
0.00I-O.036a ,b
0.5-20C
Chemical Name
Kidney
0.002-O.363 a ,b
Liver
0,005-O.246a ,b
10-500c llO-132a,e
1.2-73g 0.001-0. 132a.b
Lung aConcentration in dry tissue, bLiebscher and Smith (1968). cSollmann (1957). d Gonzales et al. (1954). eBoylen and Hardy (1967). f Buchanan (1962). gHayes and Vaughn (1977).
White phosphorus.
5-150C 7-127 d
Spleen
1399
5-250c
Synonyms Because of discoloring impurities, white phosphorus is sometimes known as yellow phosphorus. Trade names for the rodenticide include Bonide Blue Death rat killer®, Common Sense cockroach and rat preparations®, and RatNip®. The CAS registry no. is 7723-14-0.
1.61
Physical and Chemical Properties Phosphorus has an atomic weight of30.97376. It has three main allotropic forms: white, black, and red, all of which melt to form the same liquid. White phosphorus is the most highly reactive solid form, and it is the only one used as a pesticide. Red phosphorus is used for the manufacture of some fertilizers and pesticides. Under ordinary
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conditions, phosphorus is present as the molecule P4. White phosphorus is a colorless or pale yellow crystalline solid with a waxy appearance and garlic-like odor. When stored in water, it increases the water's acidity and corrodes the container at the liquids-air interface. The density of white phosphorus at 20°C is 1.83. It has a melting point of 44.1 °C and a boiling point of 280°C. It is practically insoluble in water but soluble in absolute alcohol (1 g/400 ml); in benzene (l g/31.5 ml); in chloroform (1 g/40 ml); in oil of turpentine (l g/60 ml); and in almond oil (1 g/100 ml). White phosphorus ignites spontaneously in moist air at about 30°C and in dry air at a higher temperature. Formulations and Uses Pastes in collapsible tubes and jars containing 1.5-2% or even as much as 5% phosphorus were formerly used for rat, mouse, and cockroach control. 61.13.1.2 Toxicity to Experimental Animals Oral and subcutaneous lethal doses of 2-12 mg/kg for rabbits and dogs are quoted from the literature around 1900 (Spector, 1955). It appears that no measurement of the acute or chronic toxicity of phosphorus has been made in a way that would permit proper quantitation. Rabbits and guinea pigs were not killed by a dosage of 0.66 mg/kg/day but they developed a cirrhosislike condition (Mallory, 1933). To assess risks to wildlife arising from use and environmental dispersal of white phosphorus pellets, mallard ducks were subjected to acute toxicity tests (Sparling et al., 1997). The 24-hour median oral lethal dose (LD 50) for white phosphorus dissolved in oil was similar in both sexes: 6.55 mg/kg in adult males and 7 mg/kg in adult females. The LD 50 for the ecologically more relevant pelletized form of white phosphorus was 4 mg/kg in adult males. Absorption, Distribution, Metabolism, and Excretion Phosphorus is absorbed from the respiratory and gastrointestinal tracts. It can cause severe burns to the skin, but it is said that not enough is absorbed from the burned areas to cause systemic poisoning. The dead tissue may be protective by permitting time for complete oxidation to phosphoric acid. Whether dilute formulations such as may occur in the gastrointestinal tract would be absorbed from the skin in harmful amounts has not been tested. Unreacted elemental phosphorus may be demonstrated in the tissues of people who die several days after ingesting phosphorus but not in those who die after longer periods. No quantitative study of the excretion of phosphorus seems to have been made. Mode of Action The mode of action is unknown. It has not been possible to associate the main clinical or pathological features of intoxication with inhibition of any particular enzyme or class of enzymes, although some are inhibited. It is common to speak of phosphorus as a protoplasmic poison, but it is difficult to distinguish its possible direct effects on the liver, kidney, brain, and heart from the effects of anoxia on those organs. The peripheral vascular dilatation, which is the first and
most pervasive systemic effect of phosphorus, contributes to all the disorders that may be seen in various organs. However, the mechanism of this dilatation is not clear. Phosphorus not only leads to structural damage of vital organs but also produces serious disruption of their metabolic function, as evidenced by hypo glycemia, azotemia, inhibition of glycogen formation in the liver, and many other disorders. Apparently there has been no recent review of the metabolic effects of phosphorus. Early reviews (Rubitsky and Myerson, 1949; Sollmann, 1957) list a great variety of biochemical effects, but nothing that could be termed a biochemical lesion. It is interesting that the signs and symptoms of poisoning by phosphorus are similar to those of poisoning by phosphine. This is especially true if one considers poisoning by zinc phosphide or aluminum phosphide, in which phosphine is released in the stomach and symptoms involving direct irritation of the lungs are minimal. Poisoning by phosphorus apparently has not been studied quantitatively in rats. Therefore, it is not possible to compare in this species the toxicity of preformed phosphine and of phosphine equivalent from phosphorus. It can be stated that in the rat the LD 50 for phosphine gas is 8.9 mg/kg (Section 14.6.2 of the first edition of this Handbook) and that for phosphine derived from zinc phosphide is 10.6 mg/kg (Section 61.4.2.2), whereas that for white phosphorus (expressed as phorphine equivalent) in humans may be estimated at roughly 16 mg/kg. The values are certainly of the same order of magnitude. The difference is in the direction one would expect if the conversion of phosphorus to phosphine is incomplete. Such differences as exist in the kind of injury caused by ingested phosphorus and ingested phosphine (in the form of a phosphide) might be due to differences in distribution of phosphine in the tissue related to the place of its formation. These qualitative and quantitative relationships could be explained if phosphorus were converted to phosphine in the intestine before absorption or in the liver after absorption. Such conversion probably has not been looked for and certainly has not been demonstrated, although it commonly is stated that small quantities of phosphine are formed in the putrefaction of organic matter containing phosphorus. Treatment of Poisoning in Animals Mineral oil (50 to 100 ml) prolonged the life of some 10- to 12-kg dogs and saved others from a 500-mg dose of phosphorus, which was uniformly fatal to controls that received saline cathartics or no treatment (Atkinson,1921). 61.13.1.3 Toxicity to Humans Therapeutic Use It was observed about 100 years ago that continued small doses of phosphorus to growing animals result in a layer of dense bone under the proliferating epiphyseal cartilage. In adults, the haversian and narrow canals are gradually filled with dense bone. These observations led to the illadvised use of phosphorus in treating rickets, osteoporosis, and fractures (Sollmann, 1957). It was claimed that children up to 8 years old suffered no dangerous effects from 0.5 rag/day, al-
61.13 Phosphorus though some vomited at first. The smallest dosage that exceptionally produced dangerous effects was 1 mg/day. The effects included gastrointestinal disturbance, necrosis of the jaw, and rarely typical phosphorus poisoning. Accidental and Intentional Poisoning Apparently the most thorough clinical study of poisoning by phosphorus is that of Dfaz-Rivera et al. (1950, 1961) based on a series of 56 cases involving suicide. Practically all poisoning by phosphorus has been caused by accidental or suicidal ingestion or occasionally by murder. The mortality was high. It has been conventional to describe phosphorus poisoning in three stages: initial, latent, and systemic. This convention was recognized by the turn of the last century (Hann and Veale, 1910) and was followed in a review by Rubitsky and Myerson (1949). It is universally agreed that persons who ingest large doses die quickly in profound shock. Thus, the three stages are possibly applicable only in cases in which the dose is small enough to permit survival for a week or more. Even in regard to such cases, there is apparent disagreement concerning the distinctness of the three stages. The initial symptoms include nausea, severe epigastric pain, headache, dizziness, and weakness. Diarrhea is uncommon. Vomoting is frequent but does not occur in all cases. Hematomesis may be present. The patient may be very nervous. These symptoms are commonly attributed directly or indirectly to irritation of the gastric mucosa (Dfaz-Rivera et aI., 1950). However, absorption of phosphorus must begin at once, for the very first symptoms are similar regardless of the severity of poisoning, and in severe cases the absorption of a fatal dose requires only a few bours. The initial symptoms are usually very distressing so that victims of accidental poisoning and many suicides seek medical attention promptly. However, LaDue et al. (1944) reported the death of a 15-month-old girl on the fifth by of illness, although the symptoms (abdominal pain and interminent vomiting) during the first 4 days were so mild they did not interfere with the child's play; diagnosis was on the basis of autopsy findings and not confirmed by chemical analysis or a history of ingestion. The onset of poisonmg may be delayed as much as 5 hr but is often within half an hour or less. The latent stage is characterized by a temporary and misleading improvement in the patient's condition and sense of well being. However, according to Dfaz-Rivera et al. (1950), the socalled asymptomatic period is seldom as pleasant as may appear in the literature. In spite of some apparent improvement, their patients were seldom free of nausea, anorexia, a dissagreable taste in the mouth, eructations, or epigastric or generalized abdominal pain; some had constipation, and a few vomited. This period of relative relief starts within the first 48 hr, and the severe symptoms observed in those not destined to survive begin within the first 4-5 days. In some cases the latent stage may last until 10 days after ingestion (Rubitsky and Myerson, 1949). In some cases, the improvement leading to what might otherwise constitute a latent phase is retained, and the patient goes on to complete recovery (LaDue et al., 1944).
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The third stage is ushered in by the reappearance of severe vomiting, abdominal pain, and sometimes hermitemesis. Severe damage to several organs becomes apparent. The liver is usually enlarged and very tender; jaundice, hypoprothrombinemia, and a bleeding tendency appear. Hypoglycemia is severe in some cases. Bleeding may be from the gums, stomach, intestine, or kidney, or in rare cases there may be extensive superficial ecchymoses (Hann and Veale, 1910). Injury to the kidney is evident by oligoria and azotemia. Injury to the brain rezults in severe restlessness, toxic delirium, and toxic insanity. These have been attributed to cerebral anoxia, but it seems impossible to exclude the possibility of direct toxic effects. The early appearance and rapid progression of hepatic, renal, or cerebral signs portend a best prognosis, and death frequently occurs by the eighth day or earlier. Severe agitation, coma, shock, earty azotemia, and severe hypoglycemia are the most certain signs of impending death. However, the actual occurrence of death is often sudden. Occasionally death is due to massive hemsternesis but more often has been attributed to cardiac arrest. In cases destined to survive, the third or systemic stage of poisoning begins relatively late, and the signs and symptoms are relatively mild compared to those in fatal cases. Recovery is usually established by the 14th day, although asymptomatic hepatomegaly may persist as long as 30 days after intoxication, and abnormalities of the electrocardiogram have been found as much as 27 days after ingestion. Although cirrhosis of the liver may be a sequela of poisoning (LaDue et aI., 1944; Moeschlin, 1965), recovery in other cases is complete. A 30-year-old man who had swallowed a substantial dose of phosphorus presumably with suicidal intent developed in addition to the usual symptoms and signs of poisoning also symptoms, signs, and ECG changes indicating myocardial infarction. The patient recovered. On the ninth day, the acute S-T segment changes began to resolve, and 2 days later the Q wave changes diagnostic of infarct were no longer visible. Following more specialized ECG studies, it was concluded that ischemic myocardial necrosis had not occurred, and the ECG changes had been the result of metabolic derangement secondary to poisoning (Pietras et al., 1968). A positive Chvostek sign and increased neuromuscular irritability were observed on the 15th day of acute, typical phosphorus poisoning from which the patient had already recovered to a considerable degree. Blood levels of calcium and phosphate were low, and urinary excretion of calcium exceeded intake. Both the clinical and laboratory findings spontaneously reverted to normal in 3 days. The lesion was considered to have been in the proximal renal tubules. Three other cases of hypocalcemia in phosphorus poisoning were reviewed and it was concluded that this complication may be more frequent than formerly supposed (Cushman and Alexander, 1966). Phosphorus may be effective as an abortifacient, although it may kill the mother also (Piribauer and Wallenko, 1961). Use Experience Formerly, chronic poisoning, characterized by necrosis of the mandible and maxillary bone, was caused
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by prolonged inhalation of phosphorus in industry. This kind of phosphorus poisoning might be accompanied by signs of mild liver and kidney damage, but in some cases necrosis of die jaw was the only evidence of injury (Heimann, 1946). The history of the condition commonly called "phossy jaw" has been reviewed in detail by Hughes et al. (1962), who also described the much milder condition that still occasionally occurs in industry. In spite of continuing clinical observation and modern biochemical studies, the cause of the condition remains obscure. However, it appears to be an osteomyelitis that progresses because of a local, probiotic effect of elemental phorus. Although the condition is an osteomyelitis, it differs clinically and radiologically from other osteomyelitis. The organisms involved vary from case to case and may not even be recognized pathogens. Under present conditions, phosphorus is not an important problem in industry. In a similar way, phosphorus used as a rodenticide is rarely a source of occupational poisoning, but it is highly dangerous to children who find and ingest it. Dosage Response A daily dose of 1 mg given with therapeutic intent sometimes produced gastrointestinal disturbance, necrosis of the jaw, and rarely typical phosphorus poisoning. A dose of 15 mg may be severely toxic, and as little as 50 mg (about 0.7 mg/kg) has proved fatal to an adult (Sollmann, 1957). As little as 2 mg is reported to have killed an infant (Rabinowitch, 1943). On the conirary, patients have recovered after doses thought to range from 350 to 715 mg (Caley and Kellock, 1955; F1etcher and Ga1ambos, 1963). Phosphorus is better absorbed and more toxic if ingested in a finely divided state rather than in lumps. Diaz-Rivera et al. (1950) gave evidence that even a uniform, finely ground paste of phosphorus was more quickly and completely absorbed when suspended in water, or especially when dissolved in an alcoholic beverage. Study of a series of 56 cases, in which a single formulation of phosphorus was ingested with suicidal intent, revealed a clear relationship between the amount ingested and the severity and outcome of intoxication. In this particular series, the smallest fatal dose was estimated as 190 mg (about 3 mg/kg). In 33 cases in which the dose ingested was 780 mg or less (up to about 13 mg/kg), the mortality was 18% and the time from ingestion to death was 4 days or more. In 21 individuals who ingested 1570 mg (about 26 mg/kg), the mortality was 90% and death occurred in a few hours to 3 days. Higher doses were uniformly, rapidly fatal. Laboratory Findings Urinalysis may show albuminuria, cylindruria, hematuria, and grossly abnormal concentrations of several amino acids. Liver function tests, including prothrombin time, become abnormal but usually not before hepatomegaly and jaundice are evident. Hypoglycemia may be severe; blood urea nitrogen, creatinine, and ammonia may be elevated. Any changes in the peripheral blood are small in degree and in no way characteristic. The phosphorus content of the blood is usually normal. Electrogardiogram changes have been described in detail by Diaz-Rivera et al. (1961).
In the absence of a history of exposure, the garlic-like odor of the patient's breath often gives the first indication that phosphorus is involved. In addition, phosphorus may be detected in vomitus and feces-and, according to Rubitsky and Myerson (1949), in the breath-by the light it emits in a dark room. Elementary phosphorus may be detected in excreta, gastric contents, and tissues in smaller concentration but with no greater certainty by chemical analysis. It is claimed that phosphorus and phosphine in tissue can be separately demonstrated (Curry et al., 1958). Pathology If death is sufficiently prompt, there is no pathology except irritation of the esophagus and stomach. Perforation may occur. Following survival for several days, fatty degeneration is striking in the liver, heart, and kidney but may be found in all organs, including the brain. Hepatic necrosis may be extensive, with changes occurring first in the periphery of lobules. The earliest signs of definite morphological change of the liver were found in a patient who died 6 hr after ingesting phosphorus. Periportal necrosis with degenerative changes extending toward the center of the lobules have been observed in biopsies from patients who survived. Later biopsies showed that fibrosis occurred in some but not all cases. When present, the fibrosis was periportal, sometimes forming septa between portal areas. The degree of fibrosis could not be explained by dosage or by the degree of injury evident during the acute phase and must have depended in part on one or more unidentified factors (Fletcher and Galambos, 1963; Greenberger et aI., 1964; LaDue et al., 1944). Differential Diagnosis If history of phosphorus exposure is unavailable, the initial symptoms may be confused with the gastroenteritis caused by agents such as arsenic. There is a characteristic odor of garlic to the breath and vomitus in phosphorus poisoning. Luminescence of the gastric contents or feces in a darkened room is pathognomonic. Diagnosis in the absence of a history of exposure is much easier if the patient survives long enough to develop signs of hepatic, renal, and central nervous system dysfunction. Treatment of Poisoning Because there is no specific therapy, the removal of phosphorus by vomiting or gastric lavage with large volumes of fluid is of utmost importance. According to Dfaz-Rivera et al. (1950), early vomiting or gastric lavage may be of benefit, and at times lifesaving among patients ingesting doses of 780 mg or less, but for larger doses the outcome is almost always death, regardless of treatment. Potassium permanganate, 0.1 % solution, or 2% hydrogen peroxide is used in preference to water because they may oxidize phosphorus to harmless phosphates. Mineral oil (200-250 ml), which helps to prevent phosphorus absorption, should be administered by gastric tube. Additional doses of 30-40 ml should be given by mouth every 3 hr for the first 24 hr. Absorption can be increased by digestible fats and oils, and for that reason they are contraindicated. The treatment of shock and acute hepatic or renal
61.14 Sulfur failure is instituted when necessary. According to some authors, shock caused by phosphorus responds to vasopressor agents, but in connection with poisoning associated with suicide DfazRivera et al. (1950) found these agents generally ineffective. A high carbohydrate, high protein, low fat diet supplemented with heavy doses of vitamin B and crude liver extract is generally advised. It has been reported but not confirmed that a combination of methionine and cystine can prevent necrosis of the liver (Rubitsky and Myerson, 1949). The use of cortisone acetate resulted in dramatic improvement in one case of severe poisoning after the ingestion of 825 mg of phosphorus, which is ordinarily a fatal dose. The drug was given in an initial intramuscular dose of 200 mg followed by 50 mg ever 6 hr. Oral administration was used after 48 hr. Doses were decreased on the 8th day and again later, and eventually stopped on the 65th day. The drug was given in the hope that it would improve both glycogen deposition and detoxification in the liver (Bayne et aI., 1952). It should be mentioned that although BAL has been tried in the treatment of phosphorus poisoning, there is neither clinical nor theoretical justification for it. If signs of calcium deficiency appear, calcium gluconate should be given. If phosphorus contacts the skin, it should be removed with water and later with a 1% solution of copper sulfate. Additional details about the treatment of phosphorus bums were given by Rabinowitch (1943).
61.14 SULFUR Sulfur follows phosphorus in the periodic table. It is in the same group as oxygen and selenium. Sulfur is an important constituent of protoplasm. 61.14.1 ELEMENTAL SULFUR 61.14.1.1 Identity, Properties, and Uses Chemical Name
Sulfur.
Synonyms Synonyms for sulfur include brimstone, colsul, flour sulfur, and flowers of sulfur. Trade names include CorosuI D and S®, Kolofog®, Kolospray®, Magnetic 70, 90, and 95®, Spersul®, Sulforon®, and Thiovit®. The CAS registry no. is 7704-34-9. Physical and Chemical Properties The atomic weight of sulfur is 32.064. The element exists in several allotropic forms of which the orthorhombic, cyclooctasulfur (also called a-sulfur) is stable at ordinary temperatures and pressures. The stable form consists of amber-colored crystals with a density of 2.06. When this is heated to 94.5°C, it forms monoclinic, cyclooctasulfur (B-sulfur), which melts at 115°C and boils at about 444.6°C. The vapor pressure of sulfur is 3.96 x 10-6 tOff at 30.4°C. Sulfur is soluble in carbon disulfide, slightly soluble in petroleum ether and alcohols, and insoluble in water.
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Formulations and Uses Sulfur is used as an acaricide and fungicide in the form of dusts, wettable powders, and pastes. 61.14.1.2 Toxicity to Laboratory Animals In the intestine, sulfur is converted to hydrogen sulfide; the reaction is more rapid and complete if the sulfur is in colloidal rather than crystalline or powdered form. With the colloid, an oral dosage of 175 mg/kg was rapidly fatal to some rabbits; prior to death they showed convulsions, unconsciousness, a hydrogen sulfide odor of the breath, a fall in blood pressure, bradycardia, and stimulation of respiration followed by respiratory arrest (Greengard and Woolley, 1940a). Part of the absorbed sulfur is excreted and part enters into the general metabolic pool. When radioactive sulfur was fed to sheep, activity appeared in their wool within 2 weeks, and continued feeding of sulfur changed the physical properties of the fibers by changing the number of disulfide linages (Clark and Buhrke, 1954). 61.14.1.3 Toxicity to Humans Experimental Exposure Volunteers who ingested daily doses of 500 or 750 mg of colloidal sulfur absorbed it completely and excreted most of it, mainly as the sulfate, within 24 hr. Fecal excretion of sulfur was not measurably increased. After ingestion of a single dose, excretion rose to a peak within the first 2-4 hr and then declined gradually to normal in 1420 hr. Absorption of powdered (100 mesh) sulfur was far less complete and was delayed until 8-16 hr after ingestion (Greengard and Woolley, 1940b). Therapeutic Use Binz (1897) cited an earlier reference on the medicinal use of sulfur by the followers of Hippocrates. In relatively recent times elemental sulfur has been used as a laxative and as an ointment for treatment of scabies and fungal infections. These uses were effective. In addition, sulfur ointments were used for various other skin diseases, and colloidal sulfur was injected either intravenously or intramuscularly for treatment of tuberculosis, syphilis, and especially arthritis (Sollmann, 1957). The injections especially for arthritis were probably an extension of the traditional treatment of rheumatisni and various other conditions by baths in water from sulfur springs. The oral, purgative dose of non colloidal sulfur was 20004000 mg. Sulfur in combination with molasses constituted a favorite "spring tonic" of earlier times. Its purgative action was due to hydrogen sulfide, which was formed by gradual reduction of a part of the sulfur. If the dose were retained, a dangerous amount of hydrogen sulfide might be formed, but this was rare and is unlikely in the absence of mechanical obstruction. It has long been recognized that hydrogen sulfide was formed and was exhaled in the breath and apparently secreted by the skin. Silver ornaments worn by patients treated with sulfur were turned black (Binz, 1897). Accidental and Intentional Poisoning Poisoning by lime sulfur (calcium polysulfides) probably should be attributed to
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sulfur because it decomposes to form sulfur in the presence of acid, including that in the stomach. Wakasugi and Fukui (1974) reported the suicide of a 22-year-old student who was found dead beside a 400-ml bottle 1abeled calcium polysulfide. The yellowish material with a sulfide odor was found on the face and hands as well in the stomach. Findings at autopsy were numerous but nonspecific.
61.15 SELENIUM Selenium is a metalloid. It follows arsenic in the periodic table and falls in the same group as sulfur, which it resembles in many of its chemical reactions. The toxic properties of plants grown on certain soils and eventually shown to depend on selenium have been known for a long time. In fact, the loss of hooves by horses and cows that eat these poisonous plants is so characteristic, one can be confident this was the disease observed by Marco Polo (1254-1327) in western China and described by him. The disease has been carefully studied and its quantitative relationship to selenium established. It is only recently that selenium has been shown to be an essential trace element. Thus, selenium is one of those elements known first as a poison but later found necessary for life. The difference is one of dosage. Deficient diets contain less than 0.1 ppm of selenium; normal diets contain 0.1 ppm or somewhat more; toxic diets may contain as little as 5 ppm, but toxicity is more likely at concentrations of about 25 ppm. Selenium is widely, but unevenly, distributed in nature. Traces of it ranging from 0.0000 to 0.01 ppm are found in community drinking water (Taylor, 1962). Food also contains selenium. For example, commercial wheat contains from 0.1 to 1.9 ppm. Selenium often is found in relatively high concentrations in acid soils of semiarid areas. Soils containing more than 0.5 ppm may lead to concentrations greater than 5 ppm in plants sometimes eaten by livestock and thus constitute a potential hazard. Much higher concentrations may occur in plants either because the soil in which they grow contains up to 30 ppm or because certain ones, namely some species of Astragalus and all species of Stanleya, Oonopsis, and Xylorrhizia (Kingsbury, 1964), are especially adapted to high concentrations of selenium and, in fact, are dependent on the element. The highest concentration of selenium repotted in field-grown wheat is 63 ppm (Moxon, 1958). Most vegetables contain far less but onions may contain up to 17.8 ppm (Underwood, 1956). (Indicator plants of the genera just mentioned may contain concentrations approaching 15,000 ppm (Trelease and Beath, 1949).) Livestock do not eat these plants unless driven by starvation, but their presence indicates the possibility of a problem in the soil, and if they are eaten, only a small amount of such plants may cause poisoning. At least a portion of the selenium in plant tissues is in the form of analogs of ordinary sulfur compounds. Thus Semethyl-selenocysteine has been identified in plant tissue but
is difficult to separate from its sulfur analog (Trelease et aI., 1960). If a high proportion of their diet contains high residues of selenium, livestock, especially cattle, at first become depressed and unaware of their surroundings. Gastrointestinal stasis leads to pain indicated by grunting, grinding of the teeth, and excessive salivation. A period of excitement may follow in which the animals wander aimlessly. There may be partial blindness and usually is some degree of paralysis. The aimless, stumbling wandering is recalled by the term "blind staggers," which is the common name for this form of poisoning. Autopsy usually reveals multiple small hemorrhages in the heart, kidney, and spleen. Gastroenteritis may be present and the kidney may be soft and friable (Radeleff, 1964). If the intake of selenium is smaller but more prolonged, livestock develop a more chronic condition called alkali disease. It may even be intermittent if the animals receive excessive residues only part of the time. The disease is characterized by emaciation, lack of vitality, loss of long hair from the tails of horses and cattle and from the manes of horses, and deformity of the hooves leading to pain and lameness. In severe cases, the hooves may be lost. The animals are often anemic. Congestive heart failure is common, and on autopsy atrophy and fibrosis of the heart are prominent. Cirrhosis of the liver and scarring of the kidney and erosion of the joints of the long bones may also be found at death (Radeleff, 1964). Rats fed grain grown in affected areas show many of the signs seen in poisoned livestock, including reduced food in.. take, weight loss, paralysis of the hind legs, hemorrhage into the gastrointestinal tract and the subcutaneous tissues and muscle fascia near the joints, and atrophy of the thymus and reproductive organs. Pathology is not marked in rats that die early because it requires some time to develop. Animals receiving smaller doses develop anemia and characteristic liver changes involving necrosis and atrophy accompanied by regeneration and leading to a grossly nodular appearance and to edema and other secondary changes (Franke, 1934). Adult sheep are not very susceptible to selenium poisoning. However, lambs born to ewes fed high levels of this element may have abnormal eyes and deformed feet. Dietary levels above 5 ppm are teratogenic to poultry (Franke and Tully, 1935). It was shown by Schwarz et al. (1957) and by Schwarz and Foltz (1958) that rats maintained on a diet containing an undetectable amount of selenium (less than 0.1 ppm) develop liver necrosis, and chicks onthe same diet develop an exudative diathesis. The selenium normally present in the diet of animals is in organic form. However, the rats and chicks remained healthy and developed normally when their deficient diet was supplemented by selenium at a dietary level of 0.04 ppm (rats) or 0.1 ppm (chicks). Selenium was equally effective whether in the form of sodium selenite or some organic selenium compounds. However, selenium is an integral part of Factor 3, a normally occurring dietary agent mat prevents liver necrosis in the rat (Schwarz and Foltz, 1957), and, on a molar basis, the selenium in Factor 3 is some three to four times more effective
61.15 Selenium
than the other forms of selenium studied (Schwarz and Foltz, 1958). Even elementary selenium was effective if fed at a substantially higher level. This result was confirmed and extended by many authors. Although the actions of selenium and vitamin E are similar in many respects, it has been shown that selenium is an essential nutrient in both rats (McCoy and Weswig, 1969) and chicks (Thompson and Scott, 1970) even when their diet is fully supplemented by the vitamin. More recent progress in showing that selenium is a necessary constituent of certain enzymes and other critical molecules was reviewed by Stadtman (1974) and by Underwood (1977). Although most details cannot be given here, the essentiality of sclenium is emphasized by the finding that, at least in Escherichia coli, selenium in the form of 4-selenouridine is an integral part of RNA (Hoffman and McConnell, 1974). Selenium is known to be an essential part of the enzymes glutathione peroxidase and type I iodothyronine 5'-deiodinase (Lockitch, 1989). It was later found that muscular dystrophy (white muscle disease), a spontaneous myopathy of lambs raised in certain areas, could be prevented to a large extent by supplementing the deficient diet of their mothers with selenium at the rate of 0.1 ppm using sodium selenite (Muth et aI., 1958). The disease can be prevented (or, if not too advanced, it can be cured) by subcutaneous injection of 1 mg of sodium selenate per lamb (Lagace, 1961). Radeleff (1964) mentions that traces of selenium will prevent certain myodystrophic conditions of horses as well as those of cattle and sheep. In addition to muscular dystrophy in several species and exudative diathesis in chickens, other diseases of domestic animals that occur spontaneously in selenium-deficient areas and may be corrected by selenium supplements include pancreatic fibrosis in chickens, hepatosis dietetica in pigs, unthriftiness in sheep and cattle, and reproductive disorders in various species (Underwood, 1977). The beneficial effects of selenium may be due in part to interaction with other metals. For example, micromolar concentrations of selenium partially prevented the in vitro inhibition of colony formation by mouse bone marrow cells and by certain tumor cells caused by low concentrations of organic and inorganic mercury (Strom et al., 1979). Even dietary levels that permit reproduction and lead to no recognizable disease may be inadequate for maximal growth. On the other hand, the optimal level of dietary selenium is astonishly close to the toxic level, as was discussed in connection with the logprobit model and quantitative study of the effects of small dosages (Section 1.2.7.4). Although illness of people has been attributed to excessive industrial environmental exposure to selenium (Dudley, 1938; Lemley, 1940; Lemley and Merryman, 1941), there is no agreement about the nature of the disease, and its existence has been questioned (Smith, 1941). The fact that people remain healthy in areas where livestock are subject to alkali disease has been explained by the facts that human diet is more varied and that much human food is brought in from other areas so that the dosage received by people is less than that received by livestock. This argument was less applicable in pioneer times, when both food and feed were raised locally. It remains true, of
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course, that people eat only plants that are relatively inefficient in accumulating selenium. Animals may be driven by starvation to eat plants of high selenium content. Thus people tend to receive a lower dosage than livestock living in the same areas and the people probably have a higher protein intake. However, the urinary excretion of selenium by healthy people living on seleniferous soil varies from 0.02 to 1.33 ppm (Smith and Westfall, 1936), and the daily output of some workers may reach 5 mg (Vesce, 1947). There is, therefore, some evidence that humans are less susceptible than livestock to poisoning by selenium. Nearly all injury caused by a wide variety of selenium compounds in industry arises not from their systemic toxicity but from their irritation of the skin, eyes, and respiratory tract (Browning, 1969). Apparently no form of uncomplicated selenium deficiency has been reported in people. However, increased growth in children suffering from kwashiorkor (Schwarz, 1961) and other malnutrition (Majaj and Hopkins, 1966) has been reported following administration of selenium. Additional information on the toxic and nutritional effects of selenium may be found in reviews by Frost and Lish (1975), Underwood (1977), and Lockitch (1989). 61.15.1 SODIUM SELENATE 61.15.1.1 Identity, Properties, and Uses Chemical Name
Sodium selenate.
Synonyms Trade names include Sel-Kaps®, Sel-Tox®, SS02®, and SS-20®. Code designations for the compound include P-40. The CAS registry no. is 10112-94-4. Physical and Chemical Properties Sodium selenate has the empirical formula Na204Se. The anhydrous salt and the hydrate have molecular weights of 188.94 and 369.11, respectively. The compound forms white, nonflammable crystals which have a density of 3.098 and are very water-soluble. Use Sodium selenate is an insecticide used in horticulture for the control of mites, aphids, and mealybugs. It is also used as a fungicide (ATSDR, 1996). 61.15.1.2 Toxicity to Laboratory Animals Basic Findings Animals given enough of a selenium compound to produce poisoning in 15 min show spasmotic contractions of the flanks, dyspnea, tetanic spasms, and death from respiratory failure (Franke and Moxon, 1936). Pathological changes include congestion of the liver with areas of focal necrosis; congestion of the kidney; endocarditis; myocarditis; and atony of smooth muscle of the gastrointestinal tract, gallbladder, and bladder.
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The intraperitoneal dosage of selenium (as sodium selenate) that kills 75% of rats in 48 hr is 5.25-5.75 mg/kg. The lowest dosage causing death is 3.75 mg/kg. Sodium selenate is only slightly less toxic than sodium selenite (Franke and Moxon, 1936). Lehman (1951, 1952) reported the oral LD 50 values of both compounds as 2.5 mg/kg. The intravenous dosage of selenium (as sodium selenate) fatal to 50% of rats lies between 3 and 4 mg/kg. Rabbits are more susceptible; an intravenous dosage of 2.5 mg/kg killed all those tested, as was true of an oral dosage of 4.0 mg/kg (Smith et al., 1937). Animals showing signs of poisoning have urinary concentrations of selenium ranging from 0.6 to 3.0 ppm (Smith, 1941). The National Toxicology Program (NTP, 1994) conducted a 13-week study of rats and mice given sodium selenate in their feed. 100% mortality was seen in rats at a dosage rate of 2.5 mg Se/kg/day, cholestasis at 1.6 mg Se/kg/day, minimal degeneration of renal papilla at 0.5 mg Se/kg/day, and an 11 % decrease in epididymal sperm count at about 0.2 mg Se/kg/day, the latter probably being the lowest adverse effect dosage. In mice, increased kidney weight associated with decreased water intake was seen at about 2 mg Se/kg/day and changes in estrus timing were seen at about 1 mg/kg/day, which appears to be the lowest adverse effect dosage level. Rats maintained on a diet deficient in protein (10%) were severely poisoned by a wheat selenium at a dietary level of 10 ppm, but the same concentration of selenium had virtually no effect on rats maintained on an isocaloric diet containing adequate protein, whether obtained from casein, wheat protein, lactalbumin, ovalbumin, yeast protein, liver protein, or even gelatin. Several individual amino acids were not protective (Gortner, 1940; Lewis et aI., 1940; Smith, 1939, 1941; Smith and Stohlman, 1940). Oral intake of selenium compounds leads to restricted food intake and marked loss of weight. Weight loss also occurred in an experiment involving subcutaneous administration, even though food intake was increased above the control level (Cameron, 1947). In spite of this difference, the rats showed many changes found in the other studies, including both atrophy and hypertrophy of liver lobes. Curiously enough, there was no cirrhosis (the injured liver was smooth), and the lobular distribution of hypertrophy and atrophy were reported to be opposite to that described and clearly illustrated by Franke (1934). Absorption, Distribution, Metabolism, and Excretion Most selenium compounds including sodium selenate are well absorbed (80-100% of intake) in laboratory animals, including rats, mice, dogs, and monkeys (Furchner et al., 1975; Thomson and Stewart, 1973). Absorption takes place mainly in the duodenum (Whanger et aI., 1976). In rats and dogs, selenium arising from intake of sodium selenite in the diet or in drinking water distributes widely throughout the body although concentrated in liver and kidneys (Furchner et aI., 1975; Sohn et aI., 1991; Thomson and Stewart, 1973). Selenomethionine is accumulated in tissues to a greater extent than the inorganic compounds (Behne et aI., 1991; Butler et al., 1990; Ip and Hayes, 1989; Salbe and Levander, 1990).
Selenium, being an essential trace element, undergoes extensive metabolism. It is incorporated into numerous selenoproteins (Sunde, 1990). Selenomethionine, which cannot be synthesized in the body, competes with methionine at methionine codons as does se1enocysteine with cysteine at cysteine codons. Selenium can be methylated. Dimethylselenide is exhaled and trimethylselenium ion is a major metabo1ite in urine. Methylation is believed to be a detoxication pathway (ATSDR, 1996). Selenium is excreted chiefly in the urine but about 3-10% is metabolized and excreted by the lungs, and there is some fecal excretion even when sodium selenate is administered subcutaneously (McConnell, 1942). When sodium se1enite is administered in the same way, 17-52% is exhaled within 8 hr (Schultz and Lewis, 1940). Selenium crosses the placental barrier in both animals and humans (Hadjimarkos et aI., 1959; Westfall et aI., 1938) and is found in human milk (Archimbaud et aI., 1992; Chhabra and Rao, 1994; Choy et aI., 1993; Hawkes et aI., 1994; landia1 et aI., 1976; Parizek et aI., 1971; Willhite et aI., 1990). The concentration varied from 0.13 to 0.24 ppm in the placenta and from 0.07 to 0.18 ppm in cord blood. Carcinogenicity When the possible carcinogenicity of selenium and its compounds was reviewed by an IARC Working Group, it was concluded that the available data were insufficient to allow an evaluation of the carcinogenicity of large dosages of selenium compounds in animals. The group concluded that the data provided no suggestion that selenium is carcinogenic in humans. They found evidence for a negative correlation between regional cancer death rates and selenium not convincing, even though they cited considerable evidence along this line as well as evidence that small dosages of selenium tend to protect animals from naturally occurring cancer or cancer induced by classical carcinogens (WHO, 1975). The Agency for Toxic Substances and Disease Registry (ATSDR, 1996) also concluded that "the majority of subsequent studies of humans and animals have revealed either no association between selenium intake and the incidence of cancer or a chemopreventive association." 61.15.1.3 Toxicity to Humans Accidental and Intentional Poisoning Apparently there is no record of human poisoning by sodium selenate. Ingestion of about 1000 mg of sodium selenite caused a condition similar to arsenic poisoning with death in 5 hr (Moeschlin, 1965). Laboratory Findings The levels of selenium in human tissues and body fluids is affected by geographic differences in dietary intakes (for a review, see ATSDR, 1996). For example, average plasma or serum level in the United States is 0.13 j.lg Se/ml; in Scandinavia, 0.052 j.lg Se/ml; and in New Zealand, 0.03 j.lg Se/ml. Most human tissue levels in the United States are below 1 j.lg/g with the thyroid showing some of the highest levels. Treatment of Poisoning The treatment of selenium poisoning is essentially unknown. A high-protein diet may be helpful.
61.16 Fluorine Arsenical feed supplements are beneficial in animals. BAL is not indicated. Some evidence from animal studies suggests that BAL decreases injury to the liver, increases injury to the kidneys, and has no effect on survival.
61.16 FLUORINE Fluorine is the halogen of lightest atomic weight. Its salts are widely distributed in soil and water and occur in all tissues. The toxic effects of excessive doses of various fluorides have been known for many years. Since about 1940 it has been known that children develop a high level of dental caries if they live where the drinking water contains much less than 1 ppm of the fluoride ion. Children who live where there is naturally a good supply of fluoride in the local water or children whose supply of fluoride is supplemented develop teeth that are more resistant to rotting. Thus fluoride is essential to human health if not to human life. Conflicting results have been obtained in animal experiments addressing the question of whether fluorine is an essential trace element. The difficulty is in preparing animal feed free of fluorine without changing other key nutrients in the diet. The National Research Council concluded that the data do not yet justify classifying fluorine as essential (NRC, 1989). On the other hand, the World Health Organization lists fluorine as essential for animal life without giving the supporting data (WHO, 1973). Sodium fluoride is the prototype of inorganic fluorides used as pesticides. It has received some study in this connection and a great deal more in connection with industrial air pollution on the one hand and the prevention of dental caries on the other. The remaining inorganic fluoride pesticides, cryolite, sodium fluorosilicate, and even zinc hexafluorosilicate, owe their toxicity to fluoride ion, which they release slowly on contact with water. Through their often careless use, the inorganic fluoride pesticides have been an important cause of accidental acute poisoning. There is no indication that they have been a source of any chronic condition. The fluorine compounds that have given rise to chronic effects (mainly fluorosis of teeth and bone) in both humans and animals are in all instances not pesticides. For those several reasons, it seems best to confine the sections on fluoride pesticides to information on the compounds themselves, emphasizing acute effects, and to outline what is known of the long-term environmental and occupational effects of fluorine compounds in this section. These long-term effects offer the best available indication of what continuing rate of intake of fluoride pesticides would be required to produce chronic injury. The only moderately common effect of continued excessive intake of fluoride in humans is mottling of the enamel of the teeth. This effect involves chalky patches alternating with areas of staining. No mottling has been reported where the concentration of fluoride in drinking water is less than 0.3 ppm and practically none at levels below 0.6 ppm. Mottling is exceptional and slight when the concentration in drinking water is
1407
1 ppm. The incidence and degree of mottling increase in proportion to the fluoride concentration in water within the range 2-lOppm. Chronic fluorosis was first described as occurring among animals that obtained excessive fluorides from their pastures. The residues deposited on vegetation consist of dust from indigenous soil (certain phosphate deposits or volcanic ash) or from the stack of gases of factories making phosphate fertilizer, some other phosphorus chemicals, steel, and aluminum. The compounds include native minerals, silicon tetrafluoride, and hydrofluorosilicic acid. Affected animals have stiff-legged gait, swollen hock joints, palpable lumps on their bones, and severely and irregularly worn teeth (Heyroth, 1963). The pain the animals show is a direct result of the deformity of their bones and the erosion of their articular surfaces. Lameness may lead to inability to feed; this contributes to weight loss, rough coat, reduced milk production, and infertility often seen in poisoned animals (Radeleff, 1964). Under field conditions, cattle are most susceptible, and other species are less so in the following order: cattle > sheep > swine> horses> turkeys> chickens (Radeleff, 1964). An extensive review of die literature (Schmidt and Rand, 1952) revealed that the highest dosages that did not produce fluorosis were 1-3 mg/kg/day in cattle, 5-12 mg/kg/day in swine, 1020 mg/kg/day in rats and guinea pigs, and 35-70 mg/kg/day in chickens. Mottling of the enamel in humans does not occur if excessive exposure to fluorides begins after the permanent teeth are formed. Although the bones may be affected at any time of life, even mild effects of this kind generally occur only after prolonged, poorly regulated, generally industrial exposure and, therefore, only in adults. It is probably because of this age distribution that essentially all reports of fluorosis in people involve osteosclerotic changes in bone marked by greater opacity to X-ray, thickening of the lamina, exostoses, and calcification or the ligamentous attachments. Table 61.7 shows various indices of excessive repeated fluoride intake that have been associated with detectable effects in humans. In most instances, even workers who exhibit radiological evidence of skeletal fluorosis suffer no disabling symptoms. However, about half of a group with an average urinary concentration of 16.1 ppm and with urinary levels up to 43.4 ppm complained of lack of appetite, nausea, and shortness of breath and showed some degree of anemia. A smaller proportion of the men had other complaints (Brun et aI., 1941, or Roholm, 1937). The beneficial effect of fluoride in reducing the prevalence of dental caries was discovered much more recently than the toxic effects of higher doses. The beginning of this discovery, which has permitted our children's teeth to be so much better than our own, has been traced back to (a) the observation of Eager (1901) connecting "Chiaie teeth" (now called mottled teeth) with drinking water and air charged with volcanic vapors now known to be fluorides and (b) the subsequent demonstration by Dean (1942a, b) of a quantitative relationship between caries in 7257 children and the concentrations of fluoride in their drinking waters. The decrease in caries is a linear func-
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CHAPTER 61
Inorganic and Organometal Pesticides
Table 61.7 Effect of Prolonged Intake of Fluoride in Humans Concentration Water (ppm)
Approximate Air (mg/m3)
<0.6
intake (mg/kg/day) <0.01
Effect
<0.6ppm
Rotting of teeth
~1
0.03
Reference
Urinary output
ppm
Optimal tooth development; no injury
6
Mottled dental enamel
7.8 ppm
0.17 0.2-0.35
Fluorosis in many
M011er and Gudjonsson
Fluorosisb
Roholm (1937)
Fluorosis progressive with
Brun et al. (1941)
(1932)
12-26a
0.2-1.0
16.1 2.4--43.4 ppm
duration of employment 0.14--3.43
<9.03 mg124 hr
Fluorosis by X-ray in some
Agate et al. (1949)
.=::10 6ppm
Fluorosis by X-ray
Largent et al. (1951)
Slight fluorosis in a few
Heyroth (1952)
aFluoride content calculated from gravimetric measurement of cryolite dust, slightly over half of which was of respirable size. bFirst detectable by X-ray after an avenge of 8.0 years of exposure, the shortest time observed being 2.8 years.
tion of the logarithm of fluoride concentration (Hodge, 1950). The use of fluoridation has been reviewed in detail (Campbell, 1963; Shaw, 1954). Briefly, when the concentration of fluoride is optimum, no ill effects will result and caries rates will be only 35-40% of those in communities using water supplies with little or no fluoride. The optimal concentration is about 1 ppm. However, the dosage obtained from drinking water depends on the average amount of water consumed, and this depends on temperature. Therefore, the official recommendation for a particular community depends on its annual average maximal daily air temperatures as shown in Table 61.8. The recommendations take into account the usual range of fluoride intake from food. Of 165 community water supplies analyzed for fluorine in 1961, 81 fell below the lowest optimal average of 0.7 ppm, whereas only 21 were above the highest optimal average of 1.2 ppm, and only 6 supplies had an average concentration above 1.29 ppm. The highest observed value was 10 ppm (Taylor, 1962). Table 61.8 Recommended Limits for Fluoride in Drinking Watero Annual average of maximum daily air temperatures b
Lower
Optimum
Upper
level
level
level
n)
eC)
(ppm)
(ppm)
(ppm)
50.0---53.7
10.0---12.0
0.9
1.2
1.7
53.8-58.3
12.1-14.6
0.8
l.l
1.5
58.4--63.8
14.7-17.6
0.8
1.0
1.3
63.9-70.6
17.7-21.4
0.7
0.9
1.2
70.7-79.2
21.5-26.2
0.7
0.8
1.0
79.3-90.5
26.3-32.5
0.6
0.7
0.8
aModified from U.S. Public Health Service (1962). bBased on temperature data obtained for a minimum of 5 years.
More information on the value of fluoride and on the injuries produced by both deficient and excessive intake of it may be found in a WHO monograph entitled Fluorides and Human Health (WHO, 1970). Another valuable source of information that emphasizes effects on domestic animals is a chapter by Underwood (1977). More general works on the pharmacology and toxicology of fluorides are those by Roholm (1937), Largent (1961), Hodge and Smith (1965), Smith (1966,1970), and ATSDR (1991). 61.16.1 SODIUM FLUORIDE 61.16.1.1 Identity, Properties, and Uses Chemical Name Structure
Sodium fluoride.
NaF.
Synonyms Trade names for sodium fluoride include Floridine®, Florocid®, Flura-Drops®, Karidium®, Pergantine®, T-Fluoride®, and Villiaumite®. Code designations include FDA-101. The CAS registry no. is 7681-49-4. Physical and Chemical Properties Sodium fluoride has the empirical formula FNa and a molecular weight of 41.99. It forms an odorless, noninflammable, white crystalline powder with a salty taste. It has a density of 2.78, a melting point of 993°C, and a boiling point of 1704°C.1t is soluble in water and slightly soluble in alcohol. Formulations and Uses Sodium fluoride is toxic to all forms of life. It has been used as an insecticide, rodenticide, and herbicide and as a fungicide for preservation of timber. Its toxicity
61.16 Fluorine to plants generally has restricted its use as an insecticide to bait formulations. The commercial product varies in purity from 93 to 99%. It should be colored to help distinguish it from table salt or flour. 61.16.1.2 Toxicity to Laboratory Animals Oral LD 50 values ranging from 31 to 200 mg/kg have been reported in rats (DeLopez et aI., 1976; Lehman, 1951, 1952; Lim et al., 1978; Skare et al., 1986; Smyth et al., 1969). Strain and gender differences contribute to this wide range in values. An LD 50 of 44 mg of fluoride per kilogram was reported for mice (Lim et aI., 1978). Animal studies on reproductive toxicity of sodium fluoride have been equivocal. Some have reported no effects on sperm and no accumulation of fluoride in the testes (Dunipace et aI., 1989; Li et al., 1987; Skare et al., 1986) whereas others have reported decreased male fertility (Araibi et aI., 1989). Decreased female fertility has been observed at levels producing other signs of toxicity (Messer et aI., 1973). However, others have claimed that fluoride can actually enhance fertility by increasing the absorption of iron (Messer et aI., 1973; Tao and Suttie, 1976). The National Toxicology Testing program conducted two studies of chronic oral toxicity on rats (F334/N) and mice (B6C3Fl) of sodium fluoride added to drinking water (Bucher et aI., 1991; NTP, 1990). The first study was compromised due to suspected dietary deficiencies. Based on the findings of the second study, the NTP concluded that there was "equivocal evidence of carcinogenic activity of sodium fluoride in male rats." No evidence of carcinogenicity was found in female rats or in mice of either sex. Absorption, Distribution, Metabolism, and Excretion Dietary fluoride is well absorbed from the gastrointestinal tract as the undissociated hydrogen fluoride molecule by passive absorption (Whitford and Pashley, 1984). The absorption of soluble fluoride is therefore rapid and complete in all species, including humans (Carlson et al., 1960a; Ekstrand et al., 1977, 1983), with maximum plasma fluoride concentrations attained as early as 30 min following exposure (Ekstrand et aI., 1977). Long-term retention of fluoride is mainly in calcified tissues (Wagner et al., 1958) but soft-tissue levels do rise transiently following ingestion (Carlson et aI., 1960b). Excretion is mainly via the urine (Spencer et aI., 1970). Biochemical Effects In fluorosis uncomplicated by systemic illness, fluoride ion appears to be deposited in place of hydroxyl ion in the hydroxyapatite of bone (McCann, 1953; Neuman et aI., 1950). There is little or no change in the calcium (Lantz and Smith, 1934; McClure and Mitchell, 1931) or phosphate (Phillips, 1932; Smith and Lantz, 1935) content of the bone. A part of the fluorine deposited in bone is readily excreted after fluoride feeding is stopped, but another part is more firmly held (Miller and Phillips, 1953; Savchuck and Armstrong, 1951).
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61.16.1.3 Toxicity to Humans Therapeutic Use In areas where there is a deficiency of fluoride in the drinking water, the incidence of dental caries can be reduced by administering sodium fluoride at a rate of 2.2 mg/day (Arnold et aI., 1960). However, to be of value the drug must be taken consistently during the years when the permanent teeth are being formed. Because few children or their parents can remember to carry out this preventive procedure, it is much more efficient to treat municipal water supplies so that the concentration of fluoride ion is in the range 0.7-1.2 ppm. This may be done by adding sodium fluoride at a rate of 1.5-3.0 ppm. Larger doses given therapeutically are discussed under Dosage Response. Accidental and Intentional Poisoning Acute, nonfatal poisoning is characterized by gastroenteritis (sudden nausea and vomiting, abdominal cramps, burning pain, and diarrhea) lasting for 3-6 hr and followed by collapse involving stupor and weakness lasting for about 36 hr (Sharkey and Simpson, 1933; Vallee, 1920). Tetany may occur as a result of calcium depletion. In fatal poisoning, muscular weakness appears early and is accompanied by a marked fall in blood pressure; tremor may be followed by clonic convulsions; dyspnea may be accompanied by a gray ish-blue cyanosis. Death from respiratory and cardiac arrest may occur a few minutes to 10 hr or more after ingestion but usually in 3 or 4 hr. Table 7.14 in the first edition of this Handbook lists a large outbreak of poisoning caused by the eating of sodium fluoride. In a much smaller outbreak, two people were killed by the compound when it was sold and used as flour (Fazekas, 1968). Dosage Response Hodge and Smith (1965) estimated the certainly lethal dose for a 70-kg adult to be in the range 5-10 g of sodium fluoride (32-64 mg of fluoride per kilogram) based on reports of fatal poisonings. Children may be more sensitive. Whitford (1990) reported that a dose of only 8 mg of fluoride per kilogram was responsible for the death of a 27-month-old child. Two adults were able to withstand a dose of 250 mg with minimal illness. One volunteer experienced slight nausea and epigastric distress lasting about 5 hr and salivation, which was intense for 15-30 min and stopped in half an hour; an itching sensation on his hands and feet lasted for about a week (Rabuteau, 1867). Another volunteer, who took 250 mg of sodium fluoride on an empty stomach, experienced nausea, which appeared in 2 min and in 20 min increased to a maximum accompanied by greatly increased salivation and some retching but no vomiting. Two hours after the dose, lunch was eaten but was immediately vomited. Slight nausea continued throughout the following day but disappeared on the second day (Baldwin, 1899). Therapeutic usages of sodium fluoride in the early part of the twentieth century suggested that adults might tolerate doses of 80 mg four times daily and children 20-50 mg four times daily for periods of several months (Black et at., 1949).
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CHAPTER 61
Inorganic and Organometa1 Pesticides
Table 61.9 Concentration of Fluoride in Drinking Water and in the Urine of Persons Who Drink This Water" Fluoride
Mean urinary
in water
output
(ppm)
(ppm)
2
2.09
5.5
5.46
6
7.80
8
8.71
aModified from Heyroth (1963), by permission of John Wiley and Sons, Inc.
Laboratory Findings In the general population, levels of fluorine in tissues and body fluids depend, inter alia, on consumption of fluoridated water. In fact the concentration of fluoride in urine corresponds numerically with the concentration in drinking water as shown in Table 61.9. When the concentration of fluoride in water was reduced from 8 ppm to 1 ppm, the corresponding concentrations in urine decreased from 6-8 ppm to about 2 ppm in the course of 27 months (Likins et aI., 1956). The mean fluoride concentration in plasma is in the range 0.14-0.19 ppm even when the fluoride concentration in drinking water varies from 0.15 to 2.5 ppm (WHO, 1970). Sequestration in bone and urinary excretion may act to control plasma levels (Hodge, 1961). The normal concentration of fluoride in soft tissues is in the range 0.5-1.0 ppm. Bones and teeth contain much higher concentrations, usually in the range 100-300 ppm (Hodge and Smith, 1965). In cases of poisoning, plasma calcium may be decreased and there is generally a disruption of fluid balance as a result of vomiting. Albuminuria is common. Pathology In acute fatal poisoning there is corrosion of the stomach and sometimes other parts of the gastrointestinal tract. Frequently there is edema of the brain and lungs and petechial hemorrhages of the lungs and heart. Treatment of Poisoning Vomiting should be promoted or gastric lavage carried out if the poison itself has not already produced copious vomiting. Liquid given between bouts of vomiting or used for gastric lavage should contain calcium (lime water, 1% calcium chloride solution, calcium gluconate, or even milk) to form the highly insoluble calcium fluoride. After the stomach has been cleared, a cathartic should be given. Milk of magnesia is best because it precipitates fluoride as well as clearing the intestine. Vomitus and excreta should be washed away quickly to prevent bums (Peters, 1948). Calcium gluconate (20 ml of a 10 or 20% solution) should be given intravenously at once. It should be repeated as required by the blood calcium level. It not only may alleviate carpopedal spasm but may raise the blood pressure to normal from a shock level (Rao et al., 1969).
Additional treatment should be symptomatic depending on the progress of the case but should emphasize restoration of fluid balance. In one very severe case, an intravenous atrial pacemaker was necessary to control ventricular fibrillation; feeding through a gastrostomy was necessary for 30 days (Abukurah et aI., 1973). 61.16.2 SULFURYL FLUORIDE 61.16.2.1 Identity, Properties, and Uses Chemical Name
Sulfuryl fluoride.
Synonyms Sulfuryl fluoride is manufactured under the trade name Vikane®. The CAS registry no. is 2699-79-8. Physical and Chemical Properties Sulfuryl fluoride has the empirical formula F202S and a molecular weight of 102.07. It is an odorless, colorless gas with a melting point of -135.82°C and a boiling point of - 55 .38°C. The vapor pressure is 13 x 103 torr at 25°C. The solubility in water at 25°C is 0.75 g/kg. Sulfuryl fluoride is of low solubility in most organic solvents but is miscible with methyl bromide. It is stable and noncorrosive. It is not hydrolyzed by water, but by NaOH solution. History, Formulations, and Uses Sulfuryl fluoride was first introduced in 1957. It is used for the fumigation of structures against drywood termites. Technical sulfuryl fluoride is 95% pure. 61.16.2.2 Toxicity to Laboratory Animals Both sexes of rats, guinea pigs, and rabbits and female rhesus monkeys tolerated air concentrations of 100 ppm (417 mg/m 3 ) without apparent adverse effect when exposed 7 hr a day, 5 days a week, for 6 months. Observations included survival, general appearance, behavior, and the appearance of internal organs of animals killed at the end of the experiment (Stewart, 1957). Later it was reported that a concentration of 20 ppm produced detectable effects in rats, mice, and guinea pigs exposed 7 hr a day for 6 months, but the injury present after 12 months was reversible when exposure was discontinued. Some evidence of fluorosis was observed in the incisors of mice but not in those of rats or guinea pigs (ACGIH, 1971). Although the long-term effects of sulfuryl fluoride are those of excess fluoride, it seems possible that some or all of the acute effects are those of the intact molecule. Biochemical Effects In the absence of studies on mammals, it is necessary to refer to an excellent study on termites. First it was shown that termites fumigated with a nonlethal concentration of [35 S]sulfuryl fluoride excreted inorganic sulfate, indicating that fluoride had been released. Then, using the labeled metabolic pool technique (see Section 3.3.3 in Hayes,
61.17 Chlorine 1975), separate studies of termites prefed sodium [14C]acetate or on 2p]phosphate showed that fumigated termites exhibited a spectrum of metabolic changes characteristic of fluoride toxicity (Meikle et aI., 1963).
e
61.16.2.3 Toxicity to Humans Accidental and Intentional Poisoning A case attributed to sulfuryi fluoride involved a 30-year-old man who was exposed for about 4 hr to unknown concentrations of a 99: 1 mixture of it with chloropicrin. While still at work, he experienced nausea, vomiting, crampy abdominal pain, and itching. When admitted to hospital soon afterward, vital signs were normal; the only abnormalities observed were reddening of the conjunctival, pharyngeal, and nasal mucosae; diffuse rhonchi; and paresthesia of the lateral surface of the right leg. The serum was positive for fluoride. The signs and symptoms resolved quickly; the patient was discharged on the fourth hospital day. He returned three times as an outpatient, complaining of persistent scratching of the throat, flatulence, and difficulty in reading. Ophthalmological examination revealed no abnormality, and the patient was discharged with a strong suspicion that emotional factors played a significant role in his disorder (Tax ay, 1966). In his discussion of the case, Taxay reviewed unpublished reports of animal experiments apparently indicating that dosages sufficient to produce illness from a single exposure produce respiratory irritation, central nervous system depression, and possible liver and kidney injury. Certainly the patient's major objective findings involved irritation of the eyes and respiratory tract. Only a series of cases would reveal whether the patient's signs and symptoms were typical or even whether they were due primarily to sulfuryl fluoride. Dosage Response The threshold limit value (20 mg/m3) would permit occupational exposure at the rate of 2.86 mg/kg/ day, equivalent to a fluoride exposure of 0.42 mg/kg/day. Treatment of Poisoning Treatment is symptomatic (see Section 61.16.1.3 and see Section 8.2 of the first edition of this Handbook). 61.16.3 ZINC HEXAFLUOROSILlCATE 61.16.3.1 Identity, Properties, and Uses Chemical Name Structure
Zinc hexafluorosilicate.
ZnSiF6' 6H20.
Synonyms Other names for zinc hexafluorosilicate include zinc fluosilicate, zinc fluorosilicate, and zinc silicofluoride. The CAS registry no. is 16871-71-9. Physical and Chemical Properties The anhydride of zinc hexafluorosilicate has the empirical formula F6SiZn and a molecular weight of 207.46. The hexahydrate forms white crystals which are soluble in water.
Use
1411
Mothproofing agent.
61.16.3.2 Toxicity to Laboratory Animals The oral lethal dose for the guinea pig was reported as 100 mg/kg (Simonin and Pierron, 1937). 61.16.3.3 Toxity to Humans Accidental and Intentional Poisoning A suicide served to show that poisoning by zinc hexafluorosilicate is typical of poisoning by the fluoride ion. A 35-year-old man drank half a glassful of a 5-10% solution of a commercial formulation. Following ingestion, emesis and tetanic convulsions occurred in 3.4 hr and death in 4.5 hr. Pathology was typical of fluoride poisoning (von Kraemer and Giebelmann, 1975). Laboratory Findings In two cases, including the suicide just mentioned, analysis showed that the concentration of zinc was increased over normal values more dependably in the blood than in the liver or kidneys. In fact, in the suicide, the concentration of zinc in these organs was within the range of normal (Giebelmann and Peplow, 1974). Even in cases in which analysis of stomach contents for zinc hexafluorosilicate is definitive, it would seem best to analyze the tissues for fluoride. Treatment of Poisoning See Section 61.16.1.3 and see Section 1.8 of the first edition of this Handbook.
61.17 CHLORINE Chlorine is a halogen that can occur as the highly toxic gas C12. However, the chloride ion is essential to life. Its concentration as NaCl in normal human blood is 4500-5000 ppm. Unlike compounds of lead or mercury, which have toxic properties common to the element, compounds of chlorine, whether inorganic or organic, have little in common, although certain groups of them may exhibit similarities. Thus there seems to be no characteristic chlorine toxicity as there is a mercury toxicity. This does not deny that the toxicity of many organic compounds can be modified and often substantially increased by chlorine substitution in the molecule. Such substitutions have been developed by living organisms and much more recently by chemists. Examples of naturally occurring organic chlorine compounds include ochratoxin A (Van der Merwe et aI., 1965), the compound now called penitrem A (Wilson et aI., 1968), and cyclochlorotine (Ishikawa et aI., 1970). 61.17.1 SODIUM CHLORATE 61.17.1.1 Identity, Properties, and Uses Chemical Name Structure
Sodium chlorate.
NaCl03.
1412
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Inorganic and Organometa1 Pesticides
Synonyms Trade names for sodium chlorate include Altacide®, Chlorax®, De-Fol-Ate®, Drop Leaf®, Klorex®, Ortho-C-I-Defoliant®, Rasikal®, Shed-A-Leaf®, Val-Drop®, and Weed Killer®. The CAS registry no. is 775-09-9. Physical and Chemical Properties Sodium chlorate has the empirical formula CINa03 and a molecular weight of 106.45. It forms an odorless white powder with a density of 2.5 and a melting point of 248°C. It liberates oxygen at about 300°C and decomposes upon heating. The solubility of sodium chlorate in water at O°C is 790 g/l. It is soluble in ethanol and glycerol. A strong oxidizing agent, sodium chlorate reacts with organic materials in the presence of sunlight. History, Formulations, and Uses Sodium chlorate has been in use as a weed killer since 1910. The commercial product is about 99% pure and is applied at rates of about 100 to 200 kg/ha. In some formulations, sodium chloride or other salts are included as fire retardants. Sodium borate may be formulated with sodium chlorate for both its fire retardant and herbicidal action. 61.17.1.2 Toxicity to Laboratory Animals The oral LD 50 of sodium chlorate in the rat is 1200 mg/kg (Edson, 1960). The intraperitoneal LD 50 in the mouse is 596 mg/kg (Nofre et aI., 1963). Sodium chlorate is a strong oxidizing agent. In the body it produces methemoglobin, a process involving the conversion of iron from the normal ferrous state to the ferric state. In addition, chlorate destroys red blood corpuscles, liberating hemoglobin and other proteins. The intact red cell has considerable power to reduce methemoglobin, but the mechanism apparently cannot operate after hemolysis. Thus, a high percentage of hemoglobin in the plasma may be in the form of methemoglobin whereas the percentage of metaglobin in the intact cells is low (Knight et aI., 1967). Repeated doses of chlorate, large enough to produce illness and weight loss but too small to be harmful if given only once, injure the kidney tubules severely without producing detectable methemoglobinemia (Richardson, 1937). This situation apparently has no parallel in human clinical experience but could be involved with livestock. 61.17.1.3 Toxicity to Humans Accidental and Intentional Poisoning The maJonty of deaths caused by sodium chlorate have been the result of suicide (Mengele et al., 1969; Motin et al., 1970; Oliver et aI., 1972; Timperman and Maes, 1966). The chance of ingesting a fatal dose accidentally is small unless the compound is mistaken for a drug and taken purposely, as occurred when the potassium salt mistakenly was substituted for potassium chloride (Cochrane and Smith, 1940). However, completely typical, near-fatal poisoning occurred when a 13-year-old boy "tasted" crystals of this weed killer which he found in his father's shed. In spite of
intensive treatment, recovery did not begin until about the 15th day and required a little over 40 days (Starvou et aI., 1978). Poisoning is characterized by gastritis (nausea, vomiting, and pain), anoxia (cyanosis, collapse, and terminal convulsions) secondary to methemoglobinemia, possible liver injury, and nephritis (lumbar pain and oliguria). Nephritis presumably is the direct result of chlorate ion as well as secondary to the destruction of corpuscles. The blood pressure tends to fall and the heartbeat becomes irregular. The liver and spleen may be enlarged and tender. The urine, if any, is brown or black in color and contains casts, red cells, free hemoglobin, and methemoglobin. The blood is brownish in color, and the plasma contains free hemoglobin and free methemoglobin. The red cell count is very low and the white cell count high (Knight et aI., 1967; Sollmann, 1957). Onset may be delayed as much as 12 hr (Mengele et al., 1969). Death from sodium chlorate poisoning has occurred from 4 hr to 34 days after ingestion with an average of just over 4 days (Knight et aI., 1967; Mengele et aI., 1969; Motin et aI., 1970). An entirely different kind of danger also arises from the strong oxidizing action of sodium chlorate. Its storage constitutes a special fire hazard. Sodium chlorate can explode if subjected to intense heat with or without sudden pressure. If mixed with sulfur, sugar, or some other oxidizable materials, it forms an explosive mixture that may be more powerful than gunpowder (in which the oxidant is KN03). McGregor and lackson (1969) described the pattern of hand injuries resulting from the accidental explosion of home made sodium chlorate mixture bombs in 11 teenage patients and discussed the management of the injury. Use Experience When used as a pesticide, sodium chlorate may cause irritation of the skin, eyes, or upper respiratory tract. Dosage Response Dermal absorption associated with agricultural use of sodium chlorate is not sufficient to cause systemic poisoning. Even by mouth, a large dose is required to produce illness. A 6.35% solution of potassium chlorate was long used as a gargle, or a 300-mg tablet was allowed to dissolve slowly in the mouth to treat pharyngitis before modem antibiotics became available. The toxicities of the sodium and potassium salts are similar. It was considered that a dose of 10,000 mg was toxic and 15,000-20,000 mg was fatal (Cochrane and Smith, 1940; Sollmann, 1957). The smallest recorded fatal dose was 7500 mg (Bernstein, 1930). However, vigorous treatment saved one person who had ingested about 40,000 mg (Knight et aI., 1967). Laboratory Findings In a fatal case in which chlorate could not be demonstrated in the blood or organs, it was found at a concentration of 6000 ppm in the urine (Oliver et aI., 1972). Methemoglobin levels of workers exposed to sodium chlorate were greater than those of a control group but always less than 10% (Maki, 1972).
61.18 Boron
In acute poisoning, blood potassium and urea may be increased. Even in a patient who eventually survives, the plasma may have a dark, opaque, muddy brown appearance, and spectroscopy may show hemoglobin, oxyhemoglobin, methemoglobin, and methemalbumin. The red cells may appear black and shiny like coal dust. During recovery, kidney function may return to normal very slowly and not always completely (Knight et aI., 1967). Treatment of Poisoning In case of irritation of the skin or mucous membranes, the area should be thoroughly flushed with water. Every effort should be made to remove the material if ingestion has occurred. There is no specific antidote, but oxygen, peritoneal dialysis, and exchange transfusions may be lifesaving even after a dose as high as 40 g. Dialysis is important because 95% of small doses of chlorate are excreted by the kidney, but larger doses so injure that organ that the body is almost powerless to remove the poison unaided (Knight et al., 1967).
61.18 BORON Boron has atomic number 5 and an atomic weight-of 10.81. It precedes carbon in the periodic table and is in the same group as aluminum. Although boron has long been recognized as essential to the growth of higher plants, there is no evidence so far that it is essential for animals. The lowest dietary level yet attained was 0.15 ppm (Underwood, 1977). On the other hand, the addition of 5 ppm boron (as sodium metaborate) to the drinking water of mice during their entire lifetime was without effect on their body weight, tumor incidence, or longevity (Schroeder and Mitchener, 1975). As might be expected of an element essential to plants, boron is found in all animal tissues. As reviewed by Underwood (1977) measurements of daily intake of boron have varied from 0.4 to 20 mg/person/day, depending largely if not entirely on the quantity of fruits and vegetables in the diet. The concentrations in normal human soft tissues, including brain, range from 0.06 to 0.6 ppm; concentrations in bone and teeth are somewhat higher, especially in hard-water areas. Boron has been used as an insecticide in the form of boric acid and borax, both mainly for the control of cockroaches. The acid is slightly more toxic, but the kind of illness produced by the two compounds is the same.
61.18.1 BORIC ACID 61.18.1.1 Identity, Properties, and Uses Chemical Name
Boric acid.
Synonyms Boric acid also is known as boracic acid and as orthoboric acid. The CAS registry no. is 10043-35-3.
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Physical and Chemical Properties Boric acid has the empirical formula BH303 and a molecular weight of 61.84. It forms odorless, white crystals with a pearly lustre and faintly bitter taste. It is slightly corrosive at room temperature. The density of boric acid is 1.435 at 15°C. It has a melting point of about 171°C but also may decompose upon heating. Its solubility in water at 20°C is 4.88 g/lOO ml; in cold alcohol, 1 g/l8 ml; in boiling alcohol, 1 g/6 ml; and in glycerine, 1 g/4 ml. It is stable up to 100°C. Formulations and Uses Boric acid has slight fungicidal properties. Its main use as a pesticide is in the form of tablets for the control of roaches.
61.18.1.2 Toxicity to Laboratory Animals Basic Findings The oral LD 50 values of boric acid are 2660 and 3450 mg/kg in rats and mice, respectively. The corresponding intravenous values are 1330 and 1780 mg/kg (Pfeiffer et aI., 1945), indicating relatively rapid and complete absorption from the gastrointestinal tract. Animals poisoned by boric acid showed depression, ataxia (occasionally convulsions), a fall in body temperature, a violet-red color of the skin and mucous membranes, and, in dogs, persistent vomiting and meningismus (Pfeiffer et al., 1945). Growth of rats was inhibited when their drinking water contained boric acid at a concentration of 2500 ppm, resulting in a dosage of about 325 mg/kg/day; growth was unaffected at a concentration of 1000 ppm (about 130 mg/kg/day). The mechanism of the toxic action of boron is not known. However, it is clear that no injury is done unless one or a few doses overpower the body's rather considerable ability to excrete the ion so that its concentration in the tissues and especially in the brain increases to > 10 ppm from the normal level of <1 ppm. Absorption, Distribution, Metabolism, and Excretion The similarity of the oral and intravenous LD 50 values indicates that absorption from the gastrointestinal tract is rapid and virtually complete. Following intraperitoneal injection, a peak concentration was reached in about 1.0-1.5 hr in the brain and in about 0.5 hr in other tissues. The concentrations of borate in the tissues were directly proportional to dosage over the range 18-700 mg/kg (Locksley and Sweet, 1954). In acute poisoning, the concentration of boric acid reaches high levels in all tissues, the high concentration in the brain being especially noteworthy. In a typical experiment, the concentrations were about 1110,910, and 260 ppm in brain, liver, and body fat, respectively (Pfeiffer et aI., 1945). Boric acid is excreted unchanged in the urine. Following intravenous injection in dogs, this excretion became maximal in 1-2 hr and then declined gradually; at the same time, there was an initial, brief depression of phosphorus excretion followed by a gradual rise, which at 6 hr exceeded the control value for phosphorus by five times (Pfeiffer et al., 1945). The significance of this increase in excretion of phosphorus is unknown.
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The half-life of boric acid in the blood of mice is about 65 min (Locksley and Sweet, 1954). 61.18.1.3 Toxicity to Humans Experimental Exposure Pfeiffer et al. (1945) demonstrated by animal studies and by a review of human case histories that boric acid is absorbed easily from injured skin. By studies in adult volunteers who were heavily exposed to 5% solution or to 10% boric acid ointment, they showed by analysis of the urine that no detectable boron was absorbed from the intact skin. Goldbloom and Goldbloom (1953) confirmed this negative result in 10 volunteers. Accidental and Intentional Exposure Most poisoning by boric acid has occurred in connection with its former use as a local antiseptic applied to irritated skin, bums, or wounds or from its mistaken inclusion in the feeding formula for babies. However, children also have ingested the compressed tablets made to combat roaches. Illness usually begins about 8 hr after ingestion. Signs include vomiting, diarrhea, rapidly progressing prostration, tremors, meningismus, and convulsions. An erythematous eruption of the skin that may progress to exfoliative dermatitis is characteristic. The eruption tends to be prominent on the palms, soles, and buttocks. Death may occur in less than a day or after as much as a week (Goldbloom and Goldbloom, 1953; Pfeiffer et aI., 1945; Wong et aI., 1964; Young et al., 1949). In very severe cases, onset may be within an hour and death within 4 hr (McNally and Rukstinat, 1947). Goldbloom and Goldbloom (1953) raised the possibility that Ritter's disease was, in fact, acute boric acid poisoning. Dosage Response The fatal dose is thought to be 2000-3000 mg for infants, 5000-6000 mg for children, and 15,000-20,000 mg for adults (Young et al., 1949). When through error a 42-year-old patient received an intravenous infusion of about 15,000 mg of boric acid as a 2.5% solution with 10% dextrose, she showed slight flushing, slight nausea, one episode of vomiting, and no further trouble. A total of 14,650 mg of boric acid was recovered from the urine (McIntyre and Burke, 1937). It is not clear whether this case is to be viewed as an example of individual variation-and good luck-or whether it reflects the relative protective effect of very gradual administration. In newborns, who may be more susceptible than older infants, active treatment was successful in saving all those thought to have ingested 4000 mg and one thought to have ingested 4500 mg, but was not successful against higher doses (Wong et al., 1964). Laboratory Findings Fisher and Freimuth (1958) reported the cases of two children who failed to show any illness after drinking a solution of boric acid; their blood boron levels when first examined were 13.0 and 13.8 ppm, respectively. The authors questioned an earlier report that 8.7 ppm is a fatal level
and stated that, in their experience, levels of 87-175 ppm or higher (500-1000 ppm or more expressed as boric acid) are found in cases of fatal poisoning. These results are consistent with those of Boggs and Anrode (1955), who reported the recovery of a mildly ill infant whose initial blood boron level of 48 ppm was soon reduced to 31 ppm by an exchange transfusion. The report by others of low blood boron values associated with serious or fatal poisoning may have been the result of faulty analytical techniques. One baby whose initial blood boron level was 94 ppm died in spite of vigorous treatment, but two with initial levels of 49 and 34 ppm survived (Segar, 1960). In fatal cases the concentration of boric acid in the brain has ranged from 250 to 2555 ppm (Fellows et aI., 1948; McNally and Rust, 1928; Young et al., 1949), that is, boron levels of 44447 ppm. In many cases, the highest concentration was found in the brain, but in some instances the highest concentration was in the lung or liver. Metabolic acidosis, jaundice, and increased blood urea may be present (Wong et aI., 1964). Pathology In fatal cases, pathological changes may be minimal but may include degeneration of the kidney tubules; slight degeneration of liver cells; engorgement, focal hemorrhage, and leukocytic infiltration of the skin; and edema and congestion of the brain and spinal cord (Goldbloom and Goldbloom, 1953; Wong et aI., 1964). Treatment of Poisoning Treatment of poisoning by boron is symptomatic. There is no pharmacological or specific antidote. Therefore removal of the poison is of particular importance. Boggs and Anrode (1955) demonstrated the effectiveness of exchange transfusion in removing boric acid from the body. Later, Segar (1960) showed that 4 hr of peritoneal dialysis removed about the same amount of bone acid as one exchange transfusion and concluded that continuous dialysis for 24-28 hr is, therefore, much more efficient than exchange transfusion for this purpose. Peritoneal dialysis was also used and recommended by Wong et al. (1964). Animal experiments showing that the compound is excreted readily in the urine (Pfeiffer et al., 1945) would justify the use of forced diuresis.
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CHAPTER 61
Inorganic and Organometal Pesticides
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Inorganic and Organometal Pesticides
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CHAPTER
62 Boric Acid and Inorganic Borate Pesticides Philip L. Strong D.S. Borax Inc.
62.1 INTRODUCTION Boric acid and inorganic borates are water-soluble, low-acutetoxicity, low-volatility white powders, which have been used in registered pesticide applications since 1948. They are nonmutagenic and noncarcinogenic. Animal studies at higher doses show reproductive toxicity, primarily male effects (decreased spermiation and testicular atrophy). Exposures during pregnancy to mice, rats, and rabbits resulted in developmental effects (i.e., low birth weight and skeletal defects). The reported NOAEL (No-Observed-Adverse-Effect Level) is 10 mg B/kg/day based on the most sensitive effect, low fetal body weight in rats. Recent reports involving adult accidental poisonings report minimal or no toxicity and recommend that aggressive treatment is not necessary for most patients. Earlier reported infant poisonings are based on outdated applications and misuse. The chronic "No Effect" level for humans was reported to be about 2.5 mg B/kg/day and the chronic "Adverse Effect" level was 5 mg B/kg/day. Cattle have been reported to eat sufficient quantities of borate fertilizer from open bags left in the field to result in death. Oral absorption is nearly 100%. Dermal absorption from intact skin is very low. Borates are distributed throughout the total body water. Bone is the only tissue that accumulates boron significantly above blood levels. Boric acid and inorganic borates are not metabolized below the B(OHh structure. Excretion is primarily in urine with the reported excretion half-life between 13 and 24 hours. Recent studies are demonstrating that low levels of boron are important nutritionally for both animals and humans.
62.2 BACKGROUND This chapter focuses on boric acid and inorganic borates, primarily borax and disodium octaborate tetrahydrate and their derivatives, since the major pesticide uses involve these three Handbook of Pesticide Toxicology Volume 2. Agents
materials. Dilute aqueous solutions of boric acid and borax yield identical species, most frequently boric acid, at physiological pH; therefore, toxicological data for systemic and chronic toxicity are used interchangeably for boric acid and the borate chemicals considered here. In order to allow comparisons from one chemical to another, the data are often reported in terms of the boron content, expressed as boron equivalents. The element boron does not exist in nature; it is always combined with oxygen in the form of borates. To convert boric acid data to boron equivalents, the multiplication factor is 0.1750, for borax it is 0.1134, and for disodium octaborate tetrahydrate the factor is 0.2096. Acute data, such as for dermal and respiratory toxicity, are considered separately based upon studies of the individual chemicals (i.e., boric acid and borax).
62.2.1 CHEMICAL NAMES (SYNONYMS) AND FORMULAS 1. Boric acid (boracic acid, orthoboric acid) B(OHh 2. Disodium tetraborate decahydrate (sodium tetraborate decahydrate, borax, borax decahydrate, borax lO-mole) Na2B407 . IOH20 3. Disodium octaborate tetrahydrate (none) Na2Bg013 . 4H20 (The chemical composition formula given above is an approximate composition of the solid material. The solid is amorphous and the exact structure is not known. It is the composition that gives the maximum water solubility of all the sodium borate species.)
62.2.2 PHYSICAL AND CHEMICAL PROPERTIES
1. Boric acid (CAS No. 10043-35-3)-white crystalline powder; molecular weight, 61.88; melting point, 170.9°C ± 0.2°C; vapor pressure, less than 10-4 tOff at 20o e; water solubility, 4.72% at 20°C; octanol/water partition coefficient: 0.175; pH: 5.1 (1 % solution) at 20°C; specific gravity, 1.51.
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2. Disodium tetraborate decahydrate (borax) (CAS No. 1303-96-4)-white crystalline powder; molecular weight, 381.87; melting point, 62°C (begins to dissolve in water of hydration); vapor pressure, less than 10-6 torr at 20°C; water solubility, 4.71 % at 20°C; pH, 9.24 (1 % solution) at 20°C; specific gravity, 1.71. 3. Disodium octaborate tetrahydrate (CAS No. 12280-03-4)white amorphous powder; molecular weight, 412.52; melting point, 815°C; vapor pressure, negligible at 20°C; water solubility, 9.5% at 20°C; pH, 8.5 (1 % at 23°C); bulk density, 320 to 480 kg/m 3 . 62.2.3 HISTORY, FORMULATIONS, AND USES Boric acid was first registered in the United States for pesticide use in 1948. As of September 1993, there were 189 registered products containing boric acid and its salts. Many formulations are marketed, including liquids, soluble and emulsifiable concentrates, granulars, powders, dusts, pellets, tablets, solids, paste, baits, and crystalline rods (V.S. Environmental Protection Agency, 1993). Vses include insecticides, fungicides, and algaecides, with very little herbicide usage remaining. The major borate pesticide usage includes control of cockroaches, wood-destroying insects (including termites) and fungi, and ants. Borates are used for stump treatment (sapstain) and algae control in swimming pools. They are also used for flea control in carpets, although this use has been controversial due to some high application rates reported and to concern for exposures to children and pets.
62.3 TOXICITY TO ANIMALS A number of excellent reviews have been published on the toxicology vf boric acid and inorganic borates (ATSDR, 1992; Culver et al., 1994b; ECETOC, 1995; Fail et al., 1998; Hubbard, 1998; Hubbard and Sullivan, 1996; IPCS, 1998; Moore, 1997; Murray, 1995). 62.3.1 EXPERIMENTAL LABORATORY STUDIES 62.3.1.1 Acute Studies In general, acute toxicity of boric acid and inorganic borates is considered to be low. Acute oral LD50-values in rats range from 2500 to almost 5000 mg/kg body weight (Smyth et al., 1969; Weir and Fisher, 1972). Dermal LD50 values in rabbits are greater than 2000 mg/kg body weight. Symptoms of acute toxicity include central nervous system (CNS) depression, ataxia, convulsions, and death. They are not skin irritants or sensitizers. Boric acid and disodium octaborate tetrahydrate produced mild eye irritation in the Draize test in rabbits. Borax yielded severe eye irritation in the same test. The V.S. EPA reports a Category 3 for eye irritation for boric acid and disodium octaborate
tetrahydrate and a Category 1 for borax (V.S. Environmental Protection Agency, 1993). 62.3.1.2 Chronic Studies Mutagenicity No evidence for mutagenic activity has been reported from a number of in vivo and in vitro studies (Bakke, 1991, 1992; Benson et al., 1984; Hayworth et al., 1983; Landolf, 1985; McGregor et al., 1988; National Toxicology Program, 1987; O'Loughlin, 1991; Stewart, 1991). Carcinogenicity The V.S. EPA Office of Pesticide Programs Carcinogenicity Peer Review Committee has classified boric acid as a "Group E" carcinogen, "evidence of noncarcinogenicity for humans" (V.S. Environmental Protection Agency, 1993). This classification is based on the results of two 2-year chronic feeding studies: one in rats (Weir and Fisher, 1972) and one in B6C3F1 mice (National Toxicology Program, 1987). Reproductive Toxicity Most of the toxicological studies on boric acid and inorganic borates have involved either reproductive or developmental end points. The reviews, mentioned at the beginning of Section 62.3 do a good job of discussing these effects. The first reports of reproductive effects from borate exposure came from France (Caujolle et al., 1962; Truhaut et al., 1964) in which sterility was reported in rats after ingestion of repeated high doses in the range of 1 to 24 g/kg body weight. These reports were followed up by a series of studies known as the Weir and Fisher studies (Weir and Fisher, 1972) utilizing rats and dogs in parallel studies with both boric acid and borax. This series of studies, published as final reports to V.S. Borax by Hazelton Laboratories in eight separate volumes, was summarized in one publication (Weir and Fisher, 1972). These reports were far ahead of their time for the details reported and included 2-year dietary rat and dog studies, 38-week dog studies, and three-generation reproduction studies in rats. Each study was carried out separately for boric acid and borax and confirmed that effects observed are equivalent for both materials based on boron content. Testicular degeneration, including atrophy and sterility in both rats and dogs, was reported at 1170 ppm boron levels in the feed (equivalent to 6880 ppm boric acid and 10,300 ppm borax in the feed). The NOAEL dose was 350 ppm B (2000 ppm boric acid and 3090 ppm borax). The dog study NOAEL results are equivalent to a dose of 8.8 mg B/kg body weight, which is the lowest reported NOAEL. This Weir-Fisher dog study NOAEL has been used as the key study for risk assessment in the past, and still is the study upon which the current EPA reference dose of 0.09 mg B/kg/day is based (V.S. EPA, 1998). However, the U.S. EPA is currently reevaluating the RID for boron. Reasons for not selecting the older Weir and Fisher dog study as the pivotal study for risk assessment of boric acid and inorganic borates were recently elucidated and include use of only one dog per group for some observations and problems with the control group (Murray, 1996). Recent risk assessments use a more current
62.3 Toxicity to Animals rat developmental study as the key study for risk assessment purposes (ECETOC, 1995; IPCS, 1998; Moore, 1997; Murray, 1995). This new rat developmental study is discussed in Section Developmental Toxicity. Following the Weir and Fisher studies, there have been numerous animal studies reporting reproductive effects from boric acid and inorganic borates. Species included are mice, rats, and dogs (Chapin and Ku, 1994; Dixon et aI., 1976, 1979; Fail et aI., 1989,1991; Krasovskii et aI., 1976; Ku and Chapin, 1994; Ku et aI., 1991, 1993a, b; Lee et aI., 1976; National Toxicology Program, 1987; Silaev et aI., 1977; Treinen and Chapin, 1991). Rats were fed 9000 ppm boric acid in their chow for up to four weeks; the first testicular effect observed by light microscopy was inhibited spermiation (Treinen and Chapin, 1991). At similar dose levels over a one-week-period, rats were killed at 1, 2, 3,4, and 7 days after exposure was begun, and several softtissue boron levels were measured and compared to bone and serum levels. Steady-state levels (12-40 J.-Lg Big) were reached in 3 to 4 days; no accumulation of boron above plasma levels was observed in any tissue except for bone. Bone boron levels increased throughout the 7 days, reaching 40-50 J.-Lg Big (Ku et aI., 1991). In a third study, rats were fed boric acid in their feed at 3000, 4500, 6000, and 9000 ppm for up to 9 weeks. Rats were sacrificed weekly and examined for multiple end points including serum and tissue boron levels, testis histology, weight, and sperm count. Inhibited spermiation was separated from atrophy based on dose, with inhibited spermiation observed at 3000- to 4500-ppm levels and atrophy at 6000 to 9000 ppm. Inhibited spermiation was reversible and atrophy was not (Ku et aI., 1993a). An in vitro study using several testicular cell culture systems was carried out to investigate possible mechanisms for boric acid reproductive toxicity. No effects on steroidogenic function were observed in the isolated Leydig cells. The most sensitive in vitro effect involved a reduction in DNA synthesis of mitotic/meiotic germ cells, and next was an effect on energy metabolism in Sertoli or germ cells (Ku et aI., 1993b). In addition to the three-generation rat study (Weir and Fisher, 1972), a mouse reproductive study utilizing the continuous breeding protocol has been reported (Fail et aI., 1991). Doses of 0, 1000, 4500, and 9000 ppm boric acid in the feed were utilized. No litters were obtained at 9000 ppm. All measured parameters, including fertility and litter size, were reduced at 4500 ppm. There were no differences from the controls noted for the 1000ppm group during the 14-week continuous breeding period. The final litters from the 1000-ppm and control groups were allowed to mature and breed an F2 litter. A decrease in the adjusted body weight of the F2 was reported. In addition, a crossover mating trial was carried out which established the male as the affected sex (Fail et aI., 1991). Developmental Toxicity Developmental effects related to exposure to boric acid and inorganic borates have been reported in mice, rats, and rabbits (Heindel et aI., 1992; Price et aI., 1996a, b). Boric acid was dosed in the feed of Swiss mice at average doses of 248, 452, and 1003 mg boric acid/kglday throughout gestation. Some maternal toxicity was reported at
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all dose groups (mild renal lesions in the low-dose mice). A reduction in fetal body weight was observed in the mid-dose group, and increased incidence of resorptions and malformed fetuses was observed at the highest dose. Short rib XIII (a malformation) and an increased incidence of rudimentary or missing ribs at lumbar I (a variation) were also reported. The developmental NOAEL was 248 mg boric acid/kg/day (Heindel et aI., 1992). In Sprague Dawley rats, the average dose of boric acid in the feed was 78, 163, and 330 mg boric acid/kg/day (also throughout gestation). However, a higher dose, 539 mg boric acid/kg/day, was tested on gestational days 6 to 15 only. Maternal rats showed some toxicity at all doses except the lowest dose. A reduction in fetal body weight was observed at all doses. Prenatal mortality was increased at the 539 mg/kg dose. Fetal malformations were increased over controls at all doses except the lowest, with enlarged lateral ventricles of the brain and shortening of rib XIII the most frequent. The NOAEL for maternal toxicity was 78 mg boric acid/kg/day. The NOAEL for developmental toxicity in rats was not reached in this study (Heindel et aI., 1992). Two rat studies were conducted to elucidate the Heindel observation of enlarged lateral ventricles of the brain at the two highest doses, 330 and 539 mg boric acid/kg/day (Price et aI., 1994a, b). The first study was carried out at 0.8, 1.6, and 2.4% boric acid in the diet during gestation days 14-17, and showed that this was not a sensitive period for boric acid-induced ventricular enlargement (Price et aI., 1994a). The second study incorporated boric acid in the feed at 0.4,0.5,0.6, and 0.8% during gestation days 6-15, attempting to reproduce the effects reported by Heindel et al. and to demonstrate dose dependence. After adjusting for body weight effects, there were no significant dose-related enlarged lateral ventricles, either incidence or severity. Hydrocephaly was reported to be significant in the 0.8% group (2% in controls and 15% at the highest dose) (Price et aI., 1994b). In New Zealand White rabbits, the boric acid dose was delivered by gavage on gestation days 6-19 (Price et aI., 1994a). In this study, boric acid was provided at 62.5, 125, and 250 mg boric acid/kg/day. Maternal toxicity was observed only at the highest dose. At this high dose, significant prenatal mortality was reported (90% versus 6% in controls). Malformed fetuses were also increased at the highest dose, with defects being primarily cardiovascular (intraventricular septal defect). No significant maternal or developmental toxicity was reported at the lowest two doses. The NOAEL, for both maternal and fetal developmental effects, was 125 mg boric acid/kg/day. There has been significant interest in the increase in reduction of extra lumbar ribs and in whether a reduction of a "variation" is adverse or not (Moore, 1997). A recent study in rats (Narotsky et aI., 1998) has investigated this effect, utilizing two daily gavage doses of 500 mg boric acid/kg. This very high twice-daily gavage dose level was utilized in order to increase the incidence of rib skeletal changes for meaningful observation and to provide a methodology for studying the mechanism of these changes. In this study, Narotsky noted that the types of rib changes observed depend on the days of exposure during gestation and that the changes are specific for specific exposure days. Shortening or absence of rib XIII was
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CHAPTER 62 Boric Acid and Inorganic Borates
observed in the gestation day (GD) 5-9 and the GD 6-10 exposure groups. About 90% of the fetuses had only six cervical vertebrae in the GD-9 exposure group, while the GD-I0 exposure group showed missing thoracic and lumbar vertebrae in 60% of the fetuses. The findings of Narotsky et al. show the critical periods for axial development in the rat (at least at these high dose levels), and provide an experimental model for the study of rib and vertebral changes (N arotsky et aI., 1998). Since a developmental NOAEL was not obtained in the original rat study (Heindel et aI., 1992), and the rat was found to be the most sensitive mammal (based on reduced fetal body weight), a second study was carried out (Price et aI., 1996b). This second study was carried out in the same laboratory as the first study and was designed to repeat the first study as closely as possible; but additional, lower-dose exposures were included in order to tightly bracket the NOAEL dose. The average oral doses reported in the feed were 19,36,55,75, and 144 mg boric acid/kg/day. The upper two doses were identical to those from the previous study (Heindel et aI., 1992). This new study (Price et aI., 1996b) consisted of two phases, allowing for one group to survive for postnatal observations. The study was successful in reproducing the earlier skeletal effects at the overlapping doses and reported a developmental NOAEL for rats of 55 mg boric acid/kg/day on gestation day 20 and of 74 mg boric acid/kg/day on postnatal day 21. Some reversibility of skeletal abnormalities was also reported; the incidence of short rib XIII was increased at 76 mg/kg on gestation day 20. On postnatal day 21, short rib XIII was only increased at 144 mg/kg/day. Wavy ribs, however, were completely reversed by postnatal day 21 (Price et aI., 1996b). The data from this last study were utilized in the determination of a Benchmark Dose for boric acid in rats, which was reported to be 59 mg boric acid/kg/day, in good agreement with the reported NOAEL (AlIen et aI., 1996). Blood boron concentrations from the animals used in the NOAEL study have been reported (Price et aI., 1997). Increasing dietary doses of boric acid were positively correlated with blood boron concentrations in pregnant rats. Blood boron concentrations of 1.27 ±0.298 and 1.53±0.546Il-g boron/g blood were associated with the NOAEL of 55 mg boric acid/kg/day and the LOAEL of 75 mg boric acid/kg/day reported above, respectively. 62.3.2 ACCIDENTAL POISONINGS IN ANIMALS 62.3.2.1 Cattle
One should not become complacent in handling borates just because these materials are considered to be of low toxicity. A herd of 85 mixed-breed, 350- to 400-kg beef cows were allowed to forage in a recently harvested peanut field along the edge of which a bag of borate fertilizer had been inadvertently left. The fertilizer bag was completely tom apart and 26 cows subsequently died (Sisk et al., 1988). Clinical signs were weakness, mild ataxia, general depression, and mild shivering in the neck, shoulder, and hindquarter muscles. Two cows
suffered seizures. Greenish diarrhea and dehydration were observed in most affected cattle before death. A group of Herford heifers was investigated for tolerance to elevated boron levels in drinking water (Green and Weeth, 1977). When given a choice, the heifers discriminated against concentrations greater than 29 ppm boron (added as borax) and rejected concentrations above 95 ppm. None chose boron water over tap water. The safe tolerance was not determined in this study, but was proposed to be between 40 and 150 ppm boron. In a study in yearling heifers, the boron status of cattle was found to correlate with plasma and urine concentrations (Weeth et aI., 1981). 62.3.2.2 Sheep
A study was designed to produce toxicosis in goats and to study blood and cerebrospinal fluid parameters at selected dose periods (Sisk et aI., 1990). Numerous blood factors were affected. There was evidence of CNS activity, including seizures.
62.4 TOXICITY TO HUMANS A review of the health effects in humans from exposure to inorganic borates is available in which the purpose was to provide information to assist in risk assessment and in to provide helpful guidance for clinical judgment (Culver and Hubbard, 1996). The chronic "No Effect" levels for humans was reported to be about 1 g/day of boric acid (2.5 mg B/kg/day), and the chronic adverse effect level (based on anorexia, indigestion, and exfoliative dermatitis) was 5 mg B/kg/day. 62.4.1 ACUTE EXPOSURES IN INFANTS AND CHILDREN
Numerous instances of infant deaths from accidental poisonings and medical misuse have been reported in the early literature (Goldbloom and Goldbloom, 1953; Valdes-Dapena and Arey, 1962; Wong et aI., 1964). Dose-response data from these reports are difficult to obtain due to lack of information on the initial dose, as well as on the actual quantity absorbed and on complications from vomiting, plus the rapid excretion over time. Reported blood and urine boron levels must be evaluated carefully, considering the time lapse between intake and sampling. Analytical values reported in the earlier literature tend to be significantly higher (by an order of magnitude) than values reported by current inductively coupled plasma analytical techniques. Most of the infant poisonings were due to accidental ingestion of boric acid solutions in hospitals (usually from mistaking saturated boric acid solutions intended for disinfectant use and using them in infant formulations instead of water). Also, 100% boric acid powders were used to treat diaper rash. Both of these use patterns have been discontinued, and the serious poisoning incidents have essentially disappeared. Reported symptoms included diarrhea, vomiting, erythema, exfoliation, desquamation of the skin, and central nervous system irritation.
62.4 Toxicity to Humans
A more recent incident in Malasia mayor may not involve boric acid poisoning (Chao et aI., 1991a, b). An outbreak of food poisoning resulting in the deaths of 13 children in Malasia was attributed to aflatoxin or boric acid. In some countries in the Far East, boric acid is sometimes illegally used as a food preservative in noodles. All of the victims had consumed the same type of noodles. Boric acid (boron-current analytical methods do not report the boron chemical species) was found in the blood, urine, and liver of some patients, but the level was not reported. Boron would be found normally in all tissues at low levels. Aflatoxin, the likely cause, was found at toxic levels in the tissues of victims. 62.4.2 ACUTE EXPOSURES IN ADULTS Recent reports indicate that acute boric acid accidental ingestions are not as serious as is indicated by some of the earlier literature. Two publications, summarizing borate poisoning experience from three national poison control centers, report that acute boric acid ingestions produce minimal or no toxicity, and that aggressive treatment is not necessary for most patients (Linden et aI., 1986; Litovitz et aI., 1988). At the Rocky Mountain Poison Control Center in 1983-1984, of 364 cases of boric acid exposure, there was only one fatality, and that was a suspected chronic exposure. No systemic effects were reported from the acute exposures, including four patients who had ingested 10 to 297 grams (Linden et aI., 1986). Of a total of 782 acute ingestions studied during 1981 to 1985 at two Poison Centers, no patients developed severe manifestations of toxicity. The most common symptoms were vomiting, abdominal pain, and diarrhea, with less frequent symptoms of lethargy, headache, lightheadedness, and pain. A half-life for boric acid was reported as 13.4 hours. Hemodialysis shortened the half-life (three patients). These authors suggest that aggressive treatment is not necessary in most patients (Litovitz et aI., 1988). A Japanese clinical report on one incident of ingestion of 21 g of boric acid concluded that hemodialysis was helpful and reported a serum half-life decrease from 13.46 hr to 3.76 hr. The total body clearance was 0.99literlhr and increased to 3.53 literlhr by hemodialysis (Teshima et aI., 1992). 62.4.3 CHRONIC EXPOSURES IN INFANTS Two reports of chronic borate toxicity in infants (Gordon et aI., 1973; O'Sullivan and Taylor, 1983) both involve borax and honey used on infant pacifiers as a soothing agent. These reports are reviewed by Culver and Hubbard (1996). The symptoms of chronic exposure are different than those described for acute exposures. Nine infants had suffered various seizure episodes; each had been exposed to the borax and honey pacifiers for more than four weeks. The seizures stopped when the honey and borax treatments were removed. The mean intake was calculated to be 98 mg boric acid/kg/day (Culver and Hubbard, 1996). Some vomiting and diarrhea were also involved.
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62.4.4 CHRONIC EXPOSURES IN ADULTS Some very high chronic borax and boric acid exposures over many years were reported as a result of an epilepsy treatment (Kliegel, 1980). Exposures were on the order of one to two grams per day. Symptoms included skin rash, indigestion, anorexia, exfoliative dermatitis, and alopecia. When the patients were removed from the borate treatments, the symptoms stopped and no remaining effects were noted. In one study involving adult males receiving 500 mg boric acid per day for 50 days, no adverse symptoms were reported (Wiley, 1904). Occupation-related chronic exposures at considerably lower levels have also been reported. An extensive study of respiratory effects from sodium borate exposures over a 7-year period in a California mining operation has been described (Wegman et aI., 1994). No significant adverse effects from these exposures were reported. This study also evaluated acute sensory irritant effects (eye, nose, and throat). Most of the incidences reported were nasal irritation, but all three types of irritation were reported. Severity was rated on a 13-point scale, with zero equal to "not at all," 1 as "very little," 2 as "fairly little," 3 as "moderate," 4 as "pretty much," 5 as "a lot"- ... 7 as "very much," ... , and 10 as "very, very much." Among all symptoms, 91 % were reported as ::;3 and 96% were ::;4. Irritation was described as "moderate" for borate respiratory exposures (Wegman et al., 1994). Because of the numerous reports of reproductive effects in animal studies, an epidemiology study was initiated at the same mining facility as the Wegman study above, utilizing the same exposure data (Whorton et aI., 1994a, b). The methodology involved both questionnaires and interviews and utilized the "Standard Birth Ratio" for statistical comparisons. The measured end point was "live births." No adverse reproductive effects were found. There was a nonstatistically significant excess of female offspring, which was subsequently reported to be not dose-related. A third study on the same workforce investigated the relationship of blood and urine boron content to dust exposure levels at the mine packaging and shipping facilities, where the exposures were the highest (Culver et aI., 1994a). Dust concentrations in the air ranged from 3.3 mg borax/m3 to 18 mg borax/m3 . Mean blood boron concentrations varied from 0.11 to 0.26 f.!g/g, and mean urine concentrations ranged from 3.16 to 10.72 f.!g/mg creatinine. No increase in boron accumulation was observed during the course of the workweek. The highest worker exposure, which included both dietary boron and inhaled borate dust, was 0.38 mg boron/kg/day. The mean boron level measured for these workers was 0.26 f.!g boron/g blood. This value was reported to be a factor of 10 lower than corresponding boron blood values found in the rat and dog at NAOEL exposure levels (Culver et aI., 1994a). There have been some epidemiology studies in Turkey comparing family birth rates in boron rich areas with those in lowerboron areas (Sayli et aI., 1998; Sayli, 1998; Tuccar et aI., 1998). Some village drinking waters are reported with boron levels as high as 29 ppm B. No evidence of reproductive toxicity was found in this population. Two Russian studies report adverse
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CHAPTER 62
Boric Acid and Inorganic Borates
reproductive effects in humans, one from drinking-water exposures to 0.3 mg B/kg (Krasovskii et aI., 1976) and the other from occupational exposures to high boric acid dust concentrations (Tarasenko et aI., 1972). Both studies utilize a similar epidemiology study questionnaire that is not fully described, and insufficient details are reported to adequately evaluate the studies. Krasovskii reports a limited study on the sexual function of some men with no actual numbers reported. Tarasenko's study involves 28 men between the ages of 30 and 40 years who had worked at a boric acid plant for more than 10 years, and a control group of 10 men. They report a reduction in "sexual activity" (undescribed) and gave results of semen analysis on six men. The semen analysis (details not reported) showed reduced volume, fewer spermatozoa, and a lower percentage of motile sperm.
62.5 PHARMACOKINETICS The pharmacokinetics of boric acid and inorganic borates have been reviewed previously in the literature (ECETOC, 1995; IPCS, 1998; Moore, 1997; Murray, 1998).
62.5.1.2 Dermal Absorption Because of conflicting reports in the literature regarding the degree of absorption of borates through intact human skin, partially due to difficulties in analytical sensitivity in earlier studies, a recent study was conducted, utilizing both in vivo and in vitro methods. In addition, a sensitive inductively coupled plasma mass spectrometry analytical technique involving enriched boron-lO isotope was included (Wester et aI., 1998a, b, c). Results from the in vivo clinical study suggest that dermal absorption of boric acid, borax, and disodium octaborate tetrahydrate, in intact human skin, is low and is significantly less than the normal dietary intake. The in vitro results agreed best with the in vivo study when the available dose levels were equivalent. Another human study that also reports minimal dermal absorption involved 18 females and 4 males performing a lazzercise™ routine on nylon carpet to which disodium octaborate tetrahydrate had been applied (Krieger et at., 1996). Studies are needed to access dermal absorption in damaged skin. It is known from early reports of toxicity (including death) that significant absorption occurs from exposures through damaged skin, that is, bums, injuries, and extensive diaper rash (Goldbloom and Goldbloom, 1953; Pfeiffer et aI., 1945; ValdesDapenaand Arey, 1962).
62.5.1 ABSORPTION 62.5.1.1 Oral Absorption Boric acid and borates are reported to be 100% absorbed by the oral route in both animals and humans. Six adult men given 750 mg of boric acid in water solution excreted 94% of the dose in 96 hours (Schou et aI., 1984). Eight young adult males infused with 600 mg of boric acid excreted 98.7% of the dose in 120 hours; total clearance was 54.6 ml/min/1.73 m2 ; and tl/2 = 21 hr (Jansen et aI., 1984). In another experiment, 10 students who drank spa water containing 102 mg boron excreted 92% of the boron, in the urine within a 4-day period (Job, 1973). A I-ml oral dose of borax solution was given orally to male Wistar rats (n = 20 per group) at eleven concentrations (0, 100, 200, 300, 400, 500, 600, 700, 800, 900, and 1000 mg boron/liter). Twenty-four-hoururine samples were analyzed and the reported recovery was 99.6±7.9% (Usada et aI., 1998). There was a linear correlation between dose and excretion (r-0.999) suggesting complete oral absorption. In the same study, a group of rats (n = 10) was dosed intravenously with borax at 0.4 mg boron/lOO g of body weight. The following kinetic parameters were reported: absorption tI/2 = 0.608 hr; elimination tI/2 = 4.64 hr; total clearance = 0.359 mllmin per 100 g body weight. Serum boron concentrations also indicate that borate was rapidly and completely absorbed. This study suggests that borates are not strongly bound by body proteins or tissues, and that these materials are eliminated by glomerular filtration. A study with Sprague-Dawley rats reports a positive correlation between dietary boron concentrations and whole blood concentrations (Price et aI., 1997).
62.5.2 DISTRIBUTION Boric acid is rapidly distributed throughout the body water in both animals and humans. The only tissue that appears to accumulate boron significantly above blood levels is bone. In humans, elevated levels in bone have been reported (Alexander et al., 1951; Ward, 1987). In rats fed boric acid in the chow, boron was reported in bone at levels four times those in blood; all other tissues were not significantly different than blood (Ku et aI., 1991). Adipose tissue contained 20% of the level found in rat blood. In another rat study, steady-state tissue levels were obtained three to four days after dosing (Treinen and Chapin, 1991).
62.5.3 METABOLISM Boric acid and borates are not metabolized beyond the boric acid structure B(OH3). Metabolism in biological systems is not feasible due to the high-energy requirements of breaking the boron-oxygen bond (Emsley, 1989). However, boric acid (the primary chemical structure in dilute solution at physiological pH, whether boric acid or borax is ingested) does form reversible coordination bonds with many biological chemicals, particularly those with adjacent hydroxyl groups. Much of its biological activity is thought to arise from these types of complexes (Woods, 1994).
References
62.5.4 EXCRETION Boric acid is excreted primarily in the urine in both animals and humans. In humans, the reported excretion half-life is between 13 and 21 hr (Jansen et aI., 1984; Litovitz et aI., 1988; Schou et aI., 1984; Teshima et aI., 1992). The elimination half-life reported for rats is 4.6 hr (Usada et aI., 1998).
62.6 BENEFICIAL EFFECTS OF BORIC ACID AND INORGANIC BORATES It has been known since the 1920s that boron is a required nutrient for healthy plants. Its use in fertilizers is well known. More recently, numerous publications from the USDA Human Nutrition Research Center in Grand Forks, North Dakota have reported beneficial effects from dietary boron in both animals and humans (Hunt, 1994, 1998; Hunt and Neilsen, 1981; Nielsen, 1991, 1994, 1998). Reports on the nutritional essentiality for boron in animals have been published for frogs (Fort et aI., 1998, 1999) and for fish (Eckhert, 1998; Eckhert and Rowe, 1999; Rowe et aI., 1998). In both frogs and fish, low boron environments have led to death of the embryo at very early developmental stages. Typical of other nutritionally essential elements, the dose-response curves are U -shaped. Similar studies are underway in rats and mice (Lanoue et aI., 1998, 1999). Adverse effects on blastocyst development were observed in mice.
REFERENCES Agency for Toxic Substances and Disease Registry (ATSDR) (1992). "Toxicological Profile for Boron." U.S. Department of Health and Human Services, Washington, DC. Alexander, G. v., Nusbaum, R. E., and MacDonald, N. S. (1951). The boron and lithium content of humen bones, J. Bio!. Chem. 192, 489-496. Alien, B. C., Strong, P. L., Price, C. J., Hubbard, S. A., and Daston, G. P. (1996). Benchmark dose analysis of developmental toxicity in rats exposed to boric acid. Fund. Appl. Toxico!. 32, 194-204. Bakke, J. P. (1991,1992). "Evaluation of the Potential of Boric Acid to Induce Unscheduled DNA Repair Assay using the Male F-344 Rat," Study No. 2389-V500-91 and Amendment 1 to the original report. SRI International, Menlo Park, CA. (Unpublished report to U.S. Borax Inc.) Benson, W. H., Birge, W. J., and Dorough, H. W. (1984). Absence of mutagenic activity of sodium borate (borax) and boric acid in the Salmonella preincubation test. Environ. Toxico!. Chem. 3,209-214. Caujolle, E, Familiades, c., Souard, J., and Gout, R. (1962). Limite de tolerance du rat a I'acide borique. Compt. Rend. Acad. Sci. 254, 3449-3451. Chao, T. C., Maxwell, S. M., Lyen, K, Wang, D., and Chia, H. K (199Ia). Mass poisoning in Perak, Malaysia or the tale of the nine emperor gods and rat tail noodles. J. Forensic Sci. Soc. 31, 283-288. Chao, T. C., Maxwell, S. M., and Wong, S. Y. (1991 b). An outbreak of aflatoxicosis and boric acid poisoning in Malaysia: A clinicopathological study. J. Pathology 164, 225-233. Chapin, R. E, and Ku, W W (1994). The reproductive toxicity of boric acid. Environ. Health Perspect. 102(Suppl. 7), 87-91. Culver, B. D., and Hubbard, S. A. (1996). Inorganic boron health effects in humans: An aid to risk assessment and clinical judgment. 1. Trace Elem. Exp.AJed 9, 175-184.
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Culver, B. D., Shen, P. T., Taylor, T. H., Lee-Feldstein, A., Anton-Culver, H., and Strong, P. L. (1994a). The relationship of blood- and urine-boron to boron exposure in borax workers and the usefulness of urine-boron as an exposure marker. Env. Health Perspect. 102(Suppl. 7), 133-137. Culver, B. D., Smith, R. G., Brotherton, R. J., Strong, P. L., and Gray, T. M. (1994b). Boron. In "Patty's Industrial Hygiene and Toxicology" (G. Clayton and F. Clayton, eds.), Vol. 2, Part F, pp. 4411-4448. John WiIey & Sons, Inc., New York. Dixon, R. L., Lee, I. P., and Sherins, R. J. (1976). Methods to assess reproductive effects of environmental chemicals: Studies of cadmium and boron administered orally. Env. Health Perspect. 13,59-67. Dixon, R. L., Sherins, R. J., and Lee, I. P. (1979). Assessment of environmental factors affecting male fertility. Env. Health Perspect. 30, 53-68. Eckhert, C. D. (1998). Boron stimulates embryonic trout growth. J. Nutr. 128, 2488-2493. Eckhert, C. D., and Rowe, R. I. (1999). Embryonic dysplasia and adult retinal dystrophy in boron deficient zebrafish. J. Trace Elem. Exp. AJed. 12(3), 213220. Emsley,1. (1989). "The Elements," p. 32. Oxford University Press (Clarendon Press), New York. European Center for Ecotoxicology and Toxicology of Chemicals (ECETOC) (1995). "Reproductive and General Toxicology of Some Inorganic Borates and Risk Assessment for Human Beings." Tech. Rep. No 63, ECETOC, Brussels. Fail, P. A, Chapin, R. E., Price, C. J., and Heindel, J. J. (1998). General, reproductive, developmental, and endocrine toxicity of boronated compounds. Repro. Toxico!. 12(1), 1-18. Fail, P. A, George, J. D., Sauls, H. R., Dennis, S. W, and Seely, J. C. (1989). Effect of boric acid on reproduction and fertility of rodents. Adv. Contracept. Delivery Syst. 5(3-4), 324-333. Fail, P. A, George, J. D., Seely, J. c., Grizzle, T. B., and Heindel, J. J. (1991). Reproductive toxicity of boric acid in swiss (CD-I) mice: Assessment using the continuous breeding protocol. Fund. App!. Toxico!. 17,225-239. Fort, D. J., Propst, T. L., Stover, E. L., Murray, E J., and Strong, P. L. (1999). Adverse effects from low dietary and environmental boron exposure on reproduction, development, and maturation in Xenopus laevis. J. Trace Elements in Exp. Med. 12(3),75-185. Fort, D. 1., Propst, T. L., Stover, E. L., Strong, P. L., and Murray, F. J. (1998). Adverse reproductive and developmental effects in Xenopus from insufficient boron. Biological Trace Element Research 66(1-3),237-260. Goldbloom, R. B., and Goldbloom, A. (1953). Boric acid poisoning: Report of four cases and a review of 109 cases from the world literature. J. Pediatrics 43,631-643. Gordon, A. S., Prichard, J. S., and Freedman, M. H. (1973). Seizure disorders and anemia associated with chronic borax intoxication. Can. AJed. Assoc. J. 108,719-721; 108,724. Green, G. H., and Weeth, H. J. (1977). Responses of heifers ingesting boron in water. J. Animal Sci. 45(4), 812-818. Hayworth, S., Lawlor, T., Mortelmans, K, Speck, W., and Zeiger, E. (1983). Salmonella mutagenicity test results for 250 chemicals. Environ. AJutagen l(suppl.),3-142. Heindel, J. J., Price, C. J., Field, EA., Marr, M. C., Myers, C. B., Morrissey, R. E., and Schwetz, B. A (1992). Developmental toxicity of boric acid in mice and rats. Fund. Appl. Toxico!. 34,266-277. Hubbard, S. A. (1998). Comparative toxicology of borates. Biological Trace Element Research 66(1-3), 343-357. Hubbard, S. A., and Sullivan, EM. (1996). Toxicological effects of inorganic boron compounds in animals: A review of the literature. 1. Trace Elements in Exp. Med. 9, 165-173. Hunt, C. D. (1994). The biochemical effects ofphysiologic amounts of dietary boron in animal nutrition models. Env. Health Perspect. 102(Suppl. 7), 3543. Hunt, C. D. (1998). Regulation of enzymatic activity; one possible role of dietary boron in higher animals and humans. Biological Trace Element Research 66(1-3), 205-225.
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Boric Acid and Inorganic Borates
Hunt, C. D., and Nielsen, E H. (1981). Interaction between boron and cholecalciferol in the chick. Trace Element Metabolism in Man and Animals 4, 597-600. International Program on Chemical Safety (IPCS) (1998). "Boron." Environmental Health Criteria 204. World Health Organization, Geneva. Jansen, J. A., Anderson, J., and Schou, J. S. (1984). Boric acid single dose pharmacokinetics after intravenous administration to man. Arch. Toxico!. 55,64-67. Job, C. (1973). Resorption und ausscheidung von peroral zugefuhrtem Bor. Zeitschrift fur augewandte Bader und Kleinaheilkunde 20, 137-142. Kliegel, W (1980). "BOR in Biologie, Medizin und Pharmazie." SpringerVerlag, Berlin, Heidelberg, New York. Krasovskii, G. N., Varshavskaya, S. P., and Borisov, A I. (1976). Toxic and gonadotropic effects of cadmium and boron relative to standards for these substances in drinking water. Env. Health Perspect. 13,69-75. Krieger, R. I., Dinoff, T. M., and Peterson, J. (1996). Human disodium octaborate tetrahydrate exposure following carpet flea treatment is not associated with significant dermal absorption. J. Exposure Anal. Env. Epidemiology 6(3), 279-288. Ku, W W, and Chapin, R. E. (1994). Mechanism of the testicular toxicity of boric acid in rats: In Vivo and In Vitro studies. Environ. Health Perspect. 102(Suppl. 7),99-105. Ku, W W, Chapin, R. E., Moseman, R. E, Brink, R. E., Pierce, K D., and Adams, K. Y. (1991). Tissue disposition of boron in male Fischer rats. Toxicol. Appl. Pharmacol. 111,145-151. Ku, W W, Chapin, R. E., Wine, R. N., and Gladen, B. C. (1993a). Testicular toxicity of boric acid (BA): Relationship of dose to lesion development and recovery in the F344 rat. Repro. Toxicol. 7,305-319. Ku, W W, Shih, L. M., and Chapin, R. E. (1993b). The effects of boric acid (BA) on testicular cells in culture. Repro. Toxicol. 7,321-331. Landolf, J. R. (1985). Cytotoxicity and negligible genotoxicity of borax and borax ores to cultured mammalian cells. Am. J. Ind. Med. 7, 31--43. Lanoue, L., Strong, P. L., and Keen, C. L. (1999). Adverse effects of a low boron environment on the preimplantation development of mouse embryos in vitro. J. Trace Elem. Exp. Med. 12(3), 235-250. Lanoue, L., Taubeneck, M. W., Muniz, J., Hanna, L. A., Strong, P. L., Murray, E J., Nielsen, E H., Hunt, C. D., and Keen, C. L. (1998). Assessing the effects of low boron diets on embryonic and fetal development in rodents using in vitro and in vivo model systems. Biological Trace Element Research 66(1-3), 271-298. Lee, I. P., Sherins, R. J., and Dixon, R. L. (1976). Evidence for induction of germinal aplasia in male rats by environmental exposure to boron. Toxicol. Appl. Pharmacol. 13,59-67. Linden, C. H., Hall, A H., Kulig, K W, and Rumack, B. H. (1986). Acute ingestions of boric acid. Clin. Toxicol. 24(4),269-279. Litovitz, T. L., Klein-Schwartz, W, Oderda, G. M., and Schmitz, B. E (1988). Clinical manifestations of toxicity in a series of 784 boric acid ingestions. Am. J. Emerg. Med. 6(3),209-213. Magour, S., Schramel, P., Ovcar, J., and Maser, H. (1982). Uptake and distribution of boron in rats: Interaction with ethanol and hexobarbital in the brain. Arch. Environ. Contam. Toxicol. 11,521-525. McGregor, D. B., Brown, A., Cattanach, P., Edwards, I., McBride, D., Riach, c., and Caspari, W J. (1988). Responses of the L5178Y tk-/tk- mouse lymphoma cell forward mutation assay: Ill. 72 coded chemicals. Environ. Mol. Mutagen 12, 85-154. Moore, J. A. (1997). An assessment of boric acid and borax using the IEHR evaluative process for assessing human developmental and reproductive toxicity of agents. Repro. Toxicol. 11(1), 123-160. Murray, E J. (1995). A human health risk assessment of boron (boric acid and borax) in drinking water. Regul. Toxicol. and Pharmacol. 22,221-230. Murray, E J. (1996). Issues in boron risk assessment: Pivotal study, uncertainty factors, and ADI's. 1. Trace Elem. Exp. Med. 9,231-243. Murray, E J.(1998). A comparative review of the pharmacokinetics of boric acic in rodents and humans. Biological Trace Element Research 66(1-3), 331-342.
Narotsky, M. G., Schmid, J. E., Andrews, J. E., and Kavlock, R. J. (1998). Effects of boric acid on axial skeletal development in rats. Biological Trace Element Research 66(1-3),373-394. National Toxicology Program (1987). Toxicology and Carcinogenesis Studies of Boric Acid (CAS No. 10043-35-3) in B6C3F Mice (Feed Studies). Technical Report No. 324, National Institute of Health. U.S. Department of Health and Human Services, Washington, DC. Nielsen, E H. (1991). The saga of boron in food: From a banished food preservative to a beneficial nutrient for humans. Curr. Top. Plant Biochem. Physiol. 10,274--286. Nielsen, E H. (1994). Biochemical and physiologic consequences of boron deprivation in humans. Env. Health Perspect. 102(Suppl. 7), 59-63. Nielsen, F. H. (1998). The justification for providing dietary guidance for the nutritional intake of boron. Biological Trace Element Research 66(1-3), 319-330. O'Loughlin, K. G. (1991). "Bone Marrow Erythrocyte Micronucleus Assay of Boric Acid in Swiss-Webster Mice," Study No. 2389-C400-91. SRI International, Menlo Park, CA (Unpublished report to U.S. Borax Inc.) O'Sullivan, K, and Taylor, M. (1983). Chronic boric acid poisoning in infants. Arch. Disease in Childhood 58, 737-739. Pfeiffer, C. c., Hallman, L. E, and Gersh, I. (1945). Boric acid ointment: A study of possible intoxication in the treatment of bums. J. Am. Med. Assoc. 128(4),266--274. Price, C. J., Bates, H. K., Kebede, G. A, Marr, M. c., Myers, C. B., Heindel, J. J., and Schwetz, B. A (1994a). "Final Report on the CNS Developmental Toxicity of Boric Acid (CAS #10042-35-3) in Sprague-Dawley CD Rats Exposed on Gestation Days 14--17," National Toxicology Program Report Ko. TER90123. Price, C. J., Hunter, E. S., Shelby, M. D., and Schwetz, B. A. (1994b). "Final Report on the CNS Developmental Toxicity of Boric Acid (CAS #1004335-3) in Sprague-Dawley CD Rats Exposed on Gestation Days 6-15," National Toxicology Program Report No. TER93138. Price, C. J., Marr, M. c., Myers, C. B., Seely, J. c., Heindel, J. J., and Schwetz, B. A. (1996a). The developmental toxicity of boric acid in rabbits. Fund. Appl. Toxicol. 34, 176--187. Price, C. J., Strong, P. L., Marr, M. C., Myers, C. B., and Murray, E J. (1996b). Developmental toxicity NOAEL and postnatal recovery in rats fed boric acid during gestation. Fund. Appl. Toxicol. 32, 179-193. Price, C. J., Strong, P. L., Murray, E J., and Goldberg, M. M. (1997). Blood boron concentrations in pregnant rats fed boric acid throughout gestation. Repro. Toxicol. 11(6),833-842. Rowe, R. I., Bouzan, C., Nabili, S., and Eckhert, C. D. (1998). The response of trout and zebrafish embryos to low and high boron concentrations is U-shaped. Biological Trace Element Research 66(1-3),261-270. Sayli, B. S. (1998). An assessment of fertility in boron-exposed Turkish subpopulations: 2: Evidence that boron has no effect on human reproduction. Biological Trace Element Research 66(1-3), 409--422. Sayli, B. S., Tuccar, E., and Elhan, A. H. (1998). An assessment of fertility in boron-exposed Turkish subpopulations. Repro. Toxicol. 12(3),297-304. Schou, J. S., Jansen, J. A., and Aggerbeck, B. (1984). Human pharmacokinetics and safety of boric acid. Arch. Toxicol. Suppl. 7,232-235. Silaev, A. A., Kasparov, A. A, and Korolev, V. V. (1977). Ultrastructure of the spermatogenic epithelium following boric acid poisoning. Gigiena Truda i Professionalnye Zabolevaniya 2, 34--37. Sisk, D. B., Colvin, B. M., and Bridges, C. R. (1988). Acute, fatal illness in cattle exposed to boron fertilizer. J. Am. Vet. Med. Assoc. 193(8), 943-945. Sisk, D. B., Colvin, B. M., Merrill, A, Bondar, K., and Bowen, J. M. (1990). Experimental acute inorganic boron toxicosis in the goat: Effects on serum chemistry and CSF biogenic amines. Vet. Hum. Toxicol. 32(3), 205-211. Smyth, H. E, Jr., Carpenter, J. P., Weil, C. S., Pozzanni, U. c., Streigel, J. A, and Nycum, J. S. (1969). Range finding toxicity data: List VII. Am. Ind. Hyg. Assoc. J. 30,470--476. Stewart, K R. (1991). "SalmonellalMicrosome Plate Incorporation Assay of Boric Acid," Study No. 2389-A200-91. SRI International, Menlo Park, CA (Unpublished report to U.S. Borax Inc.)
References
Tarasenko, N. Y., Kasparova, A. A., and Strongina, O. M. (1972). Effect of boric acid on the generative function in males. Gigiena Truda i Professionalnye Zabolevaniya 11, 13-16. Teshima, D., Morishita, K., Ueda, Y., Futagami, K., Higuchi, S., Komoda, T., Nanishi, F., Taniyama, T., Yoshitake, J., and Aoyama, T. (1992). Clinical management of boric acid ingestion: Pharmacokinetic assessment of efficacy of hemodialysis for treatment of acute boric acid poisoning. Treinen, K. A., and Chapin, R. E. (1991). Development of testicular lesions in F344 rats after treatment with boric acid. Toxicol. Appl. Pharmacol. 107, 325-335. Truhaut, R., Phu-Lich, N., and Lousillier, F. (1964). Sur les effets de I' ingestion repetee de petites doses de derives du bore sur les fonctions de reproduction du rat. Compt. Rend. Acad. Sci. 258,5099-5102. Tuccar, E., Elhan, A. H., Yavuz, Y., and Sayli, B. S. (1998). Comparison of infertility rates in communities from boron-rich and -poor territories. Biological Trace Element Research 66(1-3), 401-408. United States Environmental Protection Agency (U.S. EPA) (1993). "Reregistration Eligibility Decision (RED) Boric Acid and its Sodium Salts," EPA 738-R-93-0l7, Washington, DC. United States Environmental Protection Agency (U.S. EPA) (1998). "Boron. Integrated Risk Information System (IRIS)," Washington, DC. Usada, K., Kono, K., Orita, Y., Dote, T., Iguchi, K., Nishiura, H., Tominaga, M., Tagawa, T., Goto, E., and Shirai, Y. (1998). Serum and urinary boron levels in rats after single administration of sodium tetraborate. Arch. Toxicol. 72, 468-474. Valdes-Dapena, M. A., and Arey, J. B. (1962). Boric acid poisoning; Three fatal cases with pancreatic inclusions and a review of the literature. J. Pediatrics 61(4),531-546. Ward, N. L. (1987). The determination of boron in biological materials by neutron irradiation and prompt gamma-ray spectrometry. J. Radioanalytical and Nuclear Chemistry 110(2), 633-639. Weeth, H. J., Speth, C. F., and Hanks, D. R. (1981). Boron content of plasma and urine as indicators of boron intake in cattle. Am. J. Vet. Res. 42(3), 474477.
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Wegman, D. H., Eisen, E. A., Xiaohank, H., Woskie, S. R., Smith, R. G., and Garabrant, D. H. (1994). Acute and chronic respiratory effects of sodium borate particulate exposures. Env. Health Perspect. 102(Suppl. 7), 119-128. Weir, R. J., and Fisher, R. S. (1972). Toxicologic studies on borax and boric acid. Toxicol. Appl. Pharmacol. 23,351-364. Wester, R. c., Hartway, T., Maibach, H. 1., Schell, M. J., Northington, D. J., Strong, P. L., and Culver, B. D. (1998c). Summary: In vitro percutaneous absorption of boron as boric acid, borax and disodium octaborate tetrahydrate in human skin. Biological Trace Element Research 66(1-3), 111-120. Wester, R. C., Hui, X., Hartway, T., Maibach, H. 1., Bell, K., Schell, M. J., Northington, D. J., Strong, P. L., and Culver, B. D. (1998a). In vivo percutaneous absorption of boric acid, borax, and disodium octaborate tetrahydrate in humans compared to in vitro absorption in human skin from infinite and finite doses. Toxicol. Sci. 45,42-51. Wester, R. c., Hui, X., Maibach, H. 1., Bell, K., Schell, M. J., Northington, D. J., Strong, P., and Culver, B. D. (l998b). Summary: In vivo percutaneous absorption of boron as boric acid, borax and disodium octaborate tetrahydrate in humans. Biological Trace Element Research 66(1-3), 101-110. Whorton, D., Haas, J., and Trent, L. (1994a). Reproductive effects of inorganic borates on male employees: Birth rate assessment. Env. Health Perspect. 102(Suppl. 7), 129-131. Whorton, M. D., Haas, J. L., Trent, L., and Wong, W. (1994b). Reproductive effects of sodium borates on male employees: Birth rate assessment. Occup. Env. Med. 51,761-767. WiIey, W. W. (1904). "Influence of Food Preservatives and Artificial Colors on Digestion and Health. I. Boric Acid and Borax," Bulletin No. 84, Part I, Bureau of Chemistry, U.S. Department of Agriculture, Washington, DC. Wong, L. c., Heimbach, M. D., Truscott, D. R., and Duncan, B. D. (1964). Boric acid poisoning: Report of 11 cases. Canad. Med. Assoc. J. 90, 10181023. Woods, W. G. (1994). An introduction to boron: History, sources, uses, and chemistry. Env. Health Perspect. 102(Suppl. 7), 5-11.
CHAPTER
63 DEET Gerald P. Schoenig ToxicologylRegulatory Services, Inc.
Thomas G. Osimitz S.c. lohnson and Son, Inc.
63.1 INTRODUCTION DEET is a personal insect repellent that has been shown to repel biting flies, biting midges, black flies, chiggers, deer flies, fleas, gnats, horse flies, mosquitoes, no-see-ums, sand flies, small flying insects, stable flies, and ticks (U.S. EPA, 1998). In addition to its nuisance prevention value, DEET has public health benefits because of its efficacy against mosquitoes that can transmit malaria, yellow fever, denque fever, and encepholitis and ticks that carry Rocky Mountain Spotted Fever and Lyme Disease. DEET was first developed in 1946 for use by the military and was registered with the U.S. Environmental Protection Agency for use by the general public in 1957 (U.S. EPA, 1980). Unlike most pesticides, it has no food uses. DEET is applied directly to the human body and to clothing. It also is used to repel insect pests on cats, dogs, and horses and in pet living quarters. DEET is sold in a variety of product forms including liquids (including pump sprays), pressurized and nonpressurized aerosols, formulated lotions, and impregnated materials (such as towelettes and wrist bands). Recently, products combining DEET with sunscreens have become available. The concentration of DEET in these products ranges from 4 to 100%. The most commonly used solvents in formulated products are ethanol and isopropanol. DEET is used in approximately 21 % of households in the United States, or by about 30% of the U.S. population annually (U.S. EPA, 1998). This corresponds to about 27% of adult males, 31 % of adult females, and 34% of children.
63.2 CHEMISTRY DEET (CAS Registry No. 134-62-3) is a member of the N,N -dialky lamide family of chemicals. Its chemical name is N,N -diethyl-m-toluamide. Its chemical structure is presented in Fig. 63.1. The empirical formula and molecular weight of DEET are C12H17NO and 191.26 grams/mole, respectively. Handbook of Pesticide Toxicology Volume 2. Agents
Technical-grade DEET is a liquid with color ranging from water-white to amber and a faint characteristic odor. Its boiling point is 160°C at 19 mm Hg. The specific gravity at 20°C is 0.996 and its viscosity is 13.3 mPa/s at 30°C. The vapor pressure is 0.0017 and 0.0056 mm Hg at 20 and 25°C, respectively. The octanol/water partition coefficient (log Kow) is 2.00 to 2.02. DEET is practically insoluble in water and glycerin, very soluble in alcohol, ether and benzene, and sparingly soluble in petroleum ether. DEET is miscible in most petroleum hydrocarbons, alcohols, chlorinated solvents, and cottonseed oil.
63.3 TOXICOLOGY Toxicology data have been generated on DEET for the past 50 years. However, because the state of the art in toxicology testing has changed over these 50 years, most of the early studies do not meet current standards. As a result, during the reregistration process initiated by U.S. EPA (U.S. EPA, 1980), a new stateof-the-art safety database was developed by the major manufacturers and formulators of DEET. This work was conducted under the auspices of the DEET Joint Venture (DJV) and the Chemical Specialties Manufacturers Association (Washington, DC). The discussions presented below focus on the studies that were conducted as part of this comprehensive data development program. Other studies from the open literature are discussed briefly and referenced at the end of each section. All of the studies conducted by the DJV followed applicable U.S. EPA Testing Guidelines (U.S. EPA, 1984), where applicable, and were conducted in accordance with U.S. EPA Good Laboratory Practice Standards (40 CFR Part 160). Even though the principal route of exposure in humans is dermal, most of the studies conducted by the DJV were conducted by the oral route of exposure. The reason the oral route of exposure was selected for many of these studies is because little or no toxicity can be demonstrated in laboratory animals at the maximum dose level that can be applied by the dermal route of exposure, i.e., 1000 mg/kg/day. Therefore, in order to
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DEET
in the Persian Gulf War potentially could have been exposed simultaneously to these chemicals. While the findings presented in these publications provide suggestive evidence for some sort of interaction between DEET and these other chemicals, the evidence for actual chemical synergy is far from conclusive. Also, because DEET was administered orally, subcutaneously or intraperitoneally at lethal or near lethal dose levels in the laboratory studies, the relevance of these findings to normal human exposure is limited. Figure 63.1
Molecular structure of DEET.
63.3.2 SUBCHRONIC TOXICITY STUDIES
satisfy the criteria for evaluating a maximum tolerated dose, the oral route of exposure was selected in order to evaluate the critical endpoints of developmental effects, reproductive effects, neurotoxicity, chronic toxicity, and oncogenicity. The exposure pattern that most closely simulates human exposure, i.e., subchronic exposure, was evaluated by both the oral and dermal routes of exposure. The absorption, distribution, metabolism, and excretion of DEET also was examined by both of these routes of administration. 63.3.1 ACUTE TOXICITY STUDIES
Technical grade DEET has a low order of acute toxicity by the oral, dermal, and inhalation routes of exposure. In a recent set of full guideline GLP studies conducted on typical production grade DEET by one of the major DEET manufacturers (Moore, 2000), the rat oral LD50 was found to be 1892 mg/kg, the rat dermal LD50 was > 5000 mg/kg, and the rat 4-hr LC50 was > 2.0 mg/liter. In this same set of studies, technical grade DEET was shown to produce moderate erythema and edema to the skin of albino rabbits following a 4-hr occluded exposure. All skin irritation subsided within seven days. When instilled into the rabbit's eye, slight corneal opacity and slight to moderate conjunctive irritation in the form of redness, swelling, and discharge were observed. All ocular irritation cleared within seven days. In a skin sensitization study conducted using the Buehler method, no evidence of skin sensitization was observed. The acute toxicity of DEET also is discussed in a number of publications in the open literature (Ambrose et aI., 1959; Macko and Weeks, 1980; Weil, 1973). The findings reported in these publications are consistent with those described above. In an acute oral toxicity study conducted in rats of different ages, it was shown that the acute oral LD50 of DEET increased four- to fivefold between the ages of 11 days and 47-56 days of age (Verschoyle et aI., 1992). Two publications describe findings in cats and dogs that were simultaneously administered DEET and fenvalerate, a combination of ingredients used in a commercial tick and flea spray (Dorman et aI., 1990; Mount et aI., 1991). Three other publications describe studies in which potential interactions between DEET, permethrin, and/or pyridostigmine were evaluated (Abou-Donia and Wilmarth, 1996; Chaney et aI., 1999; McCain et aI., 1995). The interest in this combination of chemicals arose from the fact that servicemen
A number of subchronic studies were conducted as part of the DJV Data Development Program. These studies were conducted to develop data that could be used to select dose levels for longer term studies, to satisfy regulatory requirements for 90-day oral and dermal toxicity studies, to determine the palatability of DEET in the diet, to explore other ways of administering DEET orally when palatability in the diet was a problem, and to further address findings in other subchronic studies. 63.3.2.1 Rat 90-Day Dietary Toxicity (Johnson, 1987a)
This study was conducted in order to assess the potential for DEET to produce subchronic oral toxicity and to develop data that could be used to select dose levels for a rat two-generation reproduction study and a rat chronic toxicity/oncogenicity study. DEET was incorporated into the diet and administered to Charles River CD® rats at concentrations such that dose levels of 0, 100, 500, 1000, 2000, and 4000 mg/kg/day could be evaluated. Fifteen male and 15 female rats were evaluated at each level. Due to rejection of the treated diets and subsequent body weight depression and mortality, the 4000 mg/kg/day group was discontinued after 7 days. Animals in the remaining groups continued through to the end of the 90-day test period. At the dose levels that were evaluated for 90 days, no treatment-related clinical signs or effects on hematology, clinical chemistry, or gross pathology were observed. Decreased body weight and food consumption were observed at dose levels ::::500 mg/kg/day and signs of inanition were evident at 2000 mg/kg/day. Nonspecific liver weight increases without correlative microscopic findings occurred in all treatment groups except the 100 mg/kg/day females. Renal lesions were observed microscopically in males at all concentrations and included granular cast formation, multifocal chronic inflammation, regenerative tubular epithelium, and hyaline droplets. These renal lesions were considered to be reflective of
The purpose of this study was to develop data that could be used to select dose levels for an 18-month mouse oncogenicity
63.3 Toxicology
study. DEET was incorporated into the diet and administered to Charles River CD® -1 mice at concentrations such that dose levels of 0, 300, 1000, 3000, 6000, and, 10,000 mg/kg/day could be evaluated. Fifteen male and 15 female mice were evaluated at each level. Due to rejection of the treated diets, the 6000 and 10,000 mg/kg/day dose groups were discontinued after two weeks. Animals in the remaining groups continued through to the end of the 90-day test period. At the dose levels that were evaluated for 90 days, no treatment-related clinical signs or effects on food consumption or gross pathological observations were noted in any dose group. Decreased defecation early in the study and decreased body weight gain were observed in animals in the 3000 mg/kg/day dose group. Absolute and relative liver weights were increased in both males and females at dose levels :::: 1000 mg/kg/day and in females at 300 mg/kg/day. Multifocal hepatocellular hypertrophy was observed at a high incidence in males and females at 3000 mg/kg/day and at a lower incidence in females at 1000 mg/kg/day. The increased liver weights and the corresponding hypertrophy were considered to be adaptive changes rather than an indication of systemic toxicity. No other gross or microscopic changes were observed. Hematology and clinical chemistry evaluations were not included in this dose range-finding study. 63.3.2.3 Hamster 90-Day Dietary Toxicity (Goldenthal,1989a)
This study was conducted to evaluate further the renal lesions observed in the rat 90-day dietary toxicity study and to develop data that could be used to set dose levels for a chronic toxicity/oncogenicity study in hamsters in the event that it was determined that the rat was not the most suitable animal model for evaluating chronic effects. DEET was incorporated into the diet and administered to Golden Syrian VAFlPlus® hamsters at 0, 1000, 5000, 10,000, and 15,000 ppm. Fifteen male and 15 female hamsters were evaluated at each dietary concentration. Clinical signs including labored breathing, decreased defecation, decreased activity, pale skin, and mortality were observed at 15,000 ppm. Decreased body weights and food consumption were observed in males at 5000 ppm and in males and females at :::: 10,000 ppm. The testes and epididymides appeared smaller and decreased testis weights were observed at :::: 10,000 ppm. These observations were accompanied by an increased incidence of tubular degeneration in the testes and an associated accumulation of cellular lumenal debris in the epididymides. Blood potassium levels were elevated at 15,000 ppm. No other effects on the hematological or clinical chemistry parameters were observed. The renal lesions observed in the rat 90-day dietary toxicity study were not observed in this study. 63.3.2.4 Dog Dietary Palatability Studies (Goldenthal, 1994a, 1994b, 1995a)
Three dog dietary palatability studies were conducted in order to determine the maximum concentration of DEET that dogs would consume as part of the diet and to develop data that potentially could be used to set dose levels for a chronic toxicity
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study. The first study was two weeks in duration in which DEET dietary concentrations corresponding to 0, 300, 1000, 3000, and 10,000 ppm were evaluated. One male and one female dog were evaluated at each concentration. Clear evidence of diet rejection was observed at a dietary concentration of 10,000 ppm. The findings with regard to palatability at dietary concentrations ::s3000 ppm were inconclusive. The second palatability study was eight weeks in duration and dietary concentrations of 0, 300, 1000, 3000, and 6000 ppm were evaluated. Two male and two female dogs were evaluated at each dietary concentration. Due to rejection of the treated diet, the dogs in the 6000 ppm dose group were administered diet containing DEET at a concentration of 4500 ppm during study weeks 4 and 5 and 3000 ppm during study weeks 7 and 8. Control diet was offered to dogs in this group during study weeks 3 and 6. No treatment-related clinical signs or effects on body weight, food consumption, hematology, clinical chemistry, organ weights, or gross or microscopic pathology were observed at dietary concentrations ::s 3000 ppm. Clinical signs including thinness throughout the study and decreased activity during study weeks 3 and 5 were observed in all dogs in the 6000/4500/3000 ppm dose group. Decreased body weights and food consumption also were observed in this group throughout the study except during study weeks 3 and 6 when the control diet was administered. Food consumption by the dogs in this group was much lower when offered diet at 3000 ppm during study week 8 compared to that of the dogs offered diet at 3000 ppm throughout the study, indicating that these dogs had developed an aversion to the taste of the test substance by exposure to diets containing 6000 and 4500 ppm DEET. On the basis of the results of this study, it was determined that DEET does not produce toxicity in dogs when administered in the diet at concentrations up to 3000 ppm (approximately 75 mg/kg/day), while diets containing greater than 3000 ppm are not palatable to the beagle dog. The third palatability study was three weeks in duration and evaluated dietary concentrations of 0 and 4000 ppm. Two male and two female dogs were evaluated at each level. Depressed food consumption, decreased body weight, decreased defecation, thin and/or dehydrated appearance, and emesis were observed in the 4000 ppm treatment group. The results of this study confirmed that the highest concentration of DEET in the diet that is palatable to dogs is 3000 ppm (approximately 75 mg/kg/day). 63.3.2.5 Dog Oral Toxicity Studies Using Gelatin Capsule Administration (Goldenthal, 1994c, 1995b, 1997)
Because palatability limited the dose level of DEET that could be administered in the diet to approximately 75 mg/kg/day (3000 ppm), three studies were conducted in order to determine if higher dose levels could be administered to dogs as bolus doses using gelatin capsules. In the first of these studies, DEET was administered daily as a single bolus dose, at levels of 0, 62.5, 125, 250, and 500 mg/kg/day for a period of two weeks.
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One male and one female dog were evaluated at each dose level. Food was allowed ad libitum. No treatment-related effects on body weight or food consumption were observed. At 250 and 500 mg/kg/day emesis, ptyalism and nodding or twitching of the head and neck were observed occasionally. The male dog at 500 mg/kg also exhibited ptosis, ataxia, and convulsions. Clinical signs were observed shortly after dosing and the affected animals fully recovered shortly after their occurrence. No treatment-related clinical signs were observed at the two lower dose levels. In the second gelatin capsule study, DEET was scheduled to be administered daily as a single bolus dose at levels of 0, 75, 125, 175, and 225 mg/kg/day for a period of eight weeks. Two male and two female dogs were evaluated at each dose level. Food was allowed ad libitum. At dose levels ::::125 mg/kg/day, emesis, ptyalism, abnormal biting and scratching, and abnormal head movements were observed in one or more animals in each group during the first five days of the study. Ataxia and ptosis also were observed in some dogs at 175 and 225 mg/kg/day. In addition, convulsions were observed following the first dose in a female dog in the 225 mg/kg/day dose group and following the third dose in a male dog in the 125 mg/kg/day dose group. Clinical signs occurred shortly after dosing, after which recovery was observed. Because of the unexpected nature and severity of the clinical signs, this study was terminated after five days. The results of the first two gelatin capsule studies demonstrated that oral administration of DEET as a single bolus dose is not a suitable method of dose administration for a chronic toxicity study. Therefore, in a third study this procedure was modified by allowing the dogs access to food only for a I-hr period prior to dose administration (thereby assuring that the dogs would not be receiving the bolus dose of DEET on an empty stomach) and dividing equally the daily dose into one morning and one afternoon dose. In addition to allowing a higher daily dose of DEET to be administered, this dosing procedure was considered to have the added advantage of more closely simulating the time frame of exposure humans receive under normal conditions of use. In this study, DEET was administered orally to beagle dogs via gelatin capsules at dosage levels of 0, 50, 100, 200, and 400 mg/kg/day for a period of eight weeks. Two male and two female dogs were evaluated at each dose level. No treatment-related clinical signs or effects on body weight, food consumption, hematology, clinical chemistry, organ weights, or gross and microscopic pathology were observed at dose levels of 50 and 100 mg/kg/day. Clinical observations including abnormal head movements and ptyalism were observed at 400 mg/kg/day. Ptyalism also was observed occasionally at 200 mg/kg/day. These effects generally were observed within 1 hr of dosing and were considered to be related to the administration of DEET. Additional treatment-related effects observed at 400 mg/kg/day included decreased body weight gain in dogs of both sexes, decreased food consumption for females, and a slight decrease in cholesterol levels for males. The results of this study demonstrated that the oral administration of DEET in divided daily doses via gelatin capsules in nonfasted dogs
offered advantages over dietary administration and represented an appropriate means of orally administering DEET in a dog chronic toxicity study. Oral dose levels of 0, 30, 100, and 400 mg/kg/day were selected for the I-year chronic dog toxicity study. 63.3.2.6 Rat 90-Day Dermal Toxicity (Johnson, 1987c) The purpose of this study was to evaluate the potential for DEET to produce toxicity by the dermal route of exposure in a definitive 90-day study. DEET was applied dermally to the shaven backs of Charles River CD® rats five days per week for 13 weeks at dosage levels of 0, 100, 300, and 1000 mg/kg/day. The 1000 mg/kg/day dose level represented the maximum dose of DEET that could be applied dermally without significant runoff. Fifteen male and 15 female rats were evaluated at each dosage level. Treatment-related effects included dermal irritation, body weight depression in males at the 1000 mg/kg/day dose level, and renal lesions that were observed in males at all dose levels. Microscopically, these lesions were described as granular cast formation, multifocal inflammation, regenerative tubular epithelium, and hyaline droplets. The dermal irritation was confirmed microscopically in the form of acanthosis and/or hyperkeratosis. The renal lesions were accompanied by elevated kidney weights and slightly increased urea nitrogen levels at l3 weeks at the 300 and 1000 mg/kg/day dose levels. Increased liver weights also were observed in male and female rats in the 1000 mg/kg/day dose level. No treatmentrelated clinical signs or effects on food consumption, hematology, or ophthalmology were observed in any of the treatment groups. As was the case in the 90-day oral toxicity study, the renal lesions observed in the male rats were attributed to alpha2u-globulin nephropathy. 63.3.2.7 Micropig® 90-Day Dermal Toxicity (Goldenthal, 1991) The purpose of this study was to evaluate the potential for DEET to produce toxicity by the dermal route of exposure in a nonrodent species and to develop data to demonstrate that the renal lesions observed in the rat 90-day oral and dermal toxicity studies do not occur in a nonrodent species. DEET was applied dermally to the shaven backs of Charles River micropigs® five days per week for 13 weeks at dosage levels of 0, 100, 300, and 1000 mg/kg/day. Four male and four female micropigs® were evaluated at each level. Parameters evaluated included observations for clinical signs, dermal irritation, body weight, food consumption (approximated), hematology, clinical chemistry, organ weights, and gross and microscopic pathology. With the exception of slight skin irritation at the application site in all treatment groups, no treatment-related effects were noted in this study. The grossly observed skin irritation was confirmed microscopically in the form of acanthosis and/or hyperkeratosis.
63.3 Toxicology 63.3.2.8 Subchronic Studies to Evaluate Further Male Rat Kidney Lesions
The findings from the rat 90-day oral and dermal toxicity studies discussed above indicated that DEET produces kidney lesions in male Charles River CD® rats that are characteristic of the renal lesions produced by a wide range of chemicals that induce alpha2u-globulin accumulation in the epithelial cells of renal proximal tubules. These lesions were not observed in female CD® rats nor in animals of any other species in which 90-day toxicity studies were conducted. In order to investigate further the correlation between the renal findings observed in the rat subchronic toxicity studies and alpha2u-globulin, two additional 90-day studies were undertaken by the DJY. The first study investigated renal effects of DEET in three different strains of male rats, two strains that produce alpha2u-globulin and one that does not produce alpha2u-globulin. The second study investigated the effect of castration on the renal toxicity of DEET to male rats. This study was of interest because the synthesis of alpha2u-globulin is influenced by the level of circulating androgens. Rat 90-Day Oral Toxicity: Multistrain Study (Goldenthal, 1992) DEET was incorporated into the diet and administered to three different strains of male rats such that dose levels of o and 400 mg/kg/day could be evaluated. The three rat strains utilized were CD®, Fischer 344, and NCI Black-Reiter (NBR). CD® and Fischer rats produce high levels of alpha2u-globulin, while NBR rats do not produce alpha2u-globulin. Ten male rats of each strain were evaluated at the 0 and 400 mg/kg/day dose levels. A limited set of parameters was evaluated, including clinical observations, body weight and food consumption measurements, kidney weight measurements, and gross and microscopic examination of the kidneys. Hematoxylin and eosin and Mallory-Heidenhain stained slides of the kidney of each rat were examined microscopically. Treatment-related clinical findings in this study included slightly decreased body weights for CD® and NBR treated rats as compared with their respective control groups. In addition, slightly decreased food consumption was observed for treated CD® rats as compared with the CD® rat control group. An increase in microscopic kidney lesions, including granular cast formation (CD® only), chronic inflammation, regenerative tubular epithelium and hyaline droplets, was observed in the CD® and Fischer 344 rats but not in the NBR rats. The distribution of microscopic kidney lesions in the three strains of rats supports the correlation of DEET kidney toxicity in male rats with alpha2u-globulin nephrotoxicity. Castrated Rat 90-Day Dermal Toxicity (Goldenthal, 1989b) DEET was applied to the shaven backs of castrated Charles River CD® rats at dose levels of 0 or 1000 mg/kg/day. In addition, noncastrated CD® rats were administered DEET dermally at a dose level of 1000 mg/kg/day. Fifteen rats were evaluated in each dose group for parameters including clinical observations, dermal irritation, body weight and food consumption
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measurements, kidney weight measurements, and microscopic examination of the kidney. One kidney from each rat was processed for alpha2u-globulin analysis by immunocytochemistry using the method described in Olson et al. (1987). Microscopic examination of the kidney revealed renal lesions in both the castrated and noncastrated DEET-treated rats. These lesions included hyaline and granular cast formation, chronic inflammation, regenerative tubular epithelium, and hyaline droplets. The incidence and severity of these lesions was greater in the noncastrated treated group. No treatment-related lesions were noted in the castrated control group. Immunocytochemical techniques confirmed the presence of hyaline droplets containing alpha2u-globulin in the kidneys of the noncastrated and castrated treated animals. No alpha2u-globulin was observed in kidneys of the castrated control group. The results of this study showed that while castration reduced the renal effects of DEET, it did not completely eliminate them. The most likely explanation for the incomplete modulation of the renal effects by castration is that alpha2u-globulin synthesis is regulated by other hormones besides androgens (Roy, 1973; Sippel et aI., 1975). Conclusions Regarding Male Rat Kidney Lesions Considerable evidence currently is available to support the position that the renal lesions observed in male rats following DEET administration are associated with accumulation of alpha2uglobulin. First, the microscopic changes that were observed are characteristic of the lesions produced by other chemicals that produce these effects via this mechanism. Second, the lesions were not observed in DEET treated female rats or in mice, hamsters, dogs, or micropigs of either sex. Third, it has been shown that castration modifies the response and that similar lesions are not observed in a strain of male rat that does not produce alpha2u-globulin. It also was shown that alpha2uglobulin was present in the hyaline droplets that occur in the renal tubule cells of DEET treated rats. Since alpha2u-globulin induced nephropathy is not observed in humans, the renal lesions observed in male rats in several DEET studies are not considered to be relevant to human risk assessment. This position was supported by the U.S. EPA Risk Assessment Forum in 1991. 63.3.2.9 Subchronic Studies Referenced in the Open Literature
Three other studies in which the subchronic toxicity of DEET was evaluated by dietary incorporation in the rat (Ambrose et aI., 1959), by inhalation exposure in the rat (Macko and Bergman, 1980), and by oral gavage in the rabbit (Haight et aI., 1980) are described in the open literature. None of these studies comes close to meeting current testing standards and the findings, for the most part, are similar to those observed in the DJV studies.
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63.3.3 TERATOLOGY STUDIES 63.3.3.1 Rat Teratology (Schoenig et aI., 1994) Dose range-finding and definitive teratology studies were conducted with Charles River CD@ rats. In both studies, DEET was administered undiluted by oral gavage on gestation days 6-15. Control animals received corn oil at the volume used in the highest dose group in each study. In the dose range-finding study, five mated female rats per group were evaluated at levels of 0, 62.5, 125, 250, 500, and 1000 mg/kg/day. In the definitive study, 25 mated female rats per group were evaluated at levels of 0, 125,250, and 750 mg/kg/day. Parameters evaluated in both studies included observations for clinical signs, body weights, food consumption, maternal liver and gravid uterine weights, maternal ovarian and uterine exams, and fetal external examinations. In the definitive study, the fetuses also were given detailed internal soft-tissue and skeletal examinations. In the dose range-finding study, maternal toxicity in the form of mortality, decreased body weights, decreased food consumption, hypoactivity, ataxia, prostration, unkempt appearance, urine stains, and perioral wetness was observed at 1000 mg/kg/day. No maternal effects were observed at levels below 1000 mg/kg/day and no evidence of developmental toxicity was observed at any dose. On the basis of these results, dose levels of 0, 125,250, and 750 mg/kg/day were selected for the definitive study. In the definitive study, maternal toxicity, including mortality, decreased body weight gain, decreased body weights, decreased food consumption, hypoactivity, ataxia, decreased muscle tone, urine stains, foot splay, perinasal encrustation, perioral wetness, and increased liver weights, was observed at 750 mg/kg/day. Remarkably, the only fetotoxic effect at this clearly toxic maternal dose level was a reduction in fetal body weights per litter. The incidence of external, visceral, and skeletal variations and/or malformations were comparable in the control and treatment groups. 63.3.3.2 Rabbit Teratology (Schoenig et al., 1994) Dose range-finding and definitive teratology studies were conducted in female New Zealand White rabbits. In both studies, DEET was administered undiluted by oral gavage on gestation days 6-18. Control animals received corn oil at the volume used in the highest dose group in each study. In the dose rangefinding study, five mated female rabbits per group were evaluated at levels of 0,62.5, 125,250,500, and 1000 mg/kg/day. In the definitive study, 16 mated female rabbits per group were evaluated at levels of 0, 30, 100, and 325 mg/kg/day. Parameters evaluated in both studies included observations for clinical signs, body weights, food consumption, maternal liver and gravid uterine weights, maternal ovarian and uterine exams, and fetal external examinations. In the definitive study, the fetuses also were given detailed internal soft-tissue and skeletal examinations. In the dose range-finding study, clinical signs of maternal toxicity in the form of mortality, hypoactivity, ataxia, slow or rapid respiration, and gasping were observed at 1000
mg/kg/day. Mortality and rapid respiration also were noted at 500 mg/kg/day and rapid respiration was observed at 250 mg/kg/day. Decreased body weight gain and food consumption were observed at :::500 mg/kg/day. There were no signs of maternal toxicity at 125 mg/kg/day and below and no signs of fetotoxicity or developmental toxicity at any dose level. On the basis of these results, dose levels of 0, 30, 100, and 325 mg/kg/day were selected for the definitive study. Maternal toxicity in the form of decreased body weight gain and food consumption was observed in the 325 mg/kg/day group. No evidence of fetotoxicity was observed and the incidence of external, visceral, and skeletal variations and/or malformations were comparable in the control and treatment groups. 63.3.3.3 Teratology Studies Referenced in the Open Literature Four other developmental toxicity studies are described in the open literature. Three of these studies were conducted in rats by the oral (Sterner, 1977), dermal (Gleiberman et aI., 1975), and subcutaneous (Wright et aI., 1992) routes of administration, respectively. The fourth study was conducted in rabbits by the dermal route of administration (Angerhofer and Weeks, 1980). None of these studies came close to meeting today's standards for developmental toxicity studies. While no increases in malformations or variations were described, evidence of fetotoxicity at maternally toxic doses was noted in three studies (Angerhofer and Weeks, 1980; Gleiberman et aI., 1975; Sterner, 1977). 63.3.4 REPRODUCTIVE TOXICITY STUDIES 63.3.4.1 Rat Two-Generation Reproduction Study (Schardein, 1989) A two-generation reproduction study was conducted in Charles River CD@ rats. DEET was incorporated into the diet and administered to the rats at concentrations of 0, 500, 2000, and 5000 ppm. The Fo parental generation consisted of 28 males and 28 females per group which were administered treated or control diet for at least 80 days prior to mating. Twenty-eight male and 28 female offspring per group from the F 1 generation were selected randomly to become the parents of the F2 generation. These animals were treated for at least 93 days prior to mating. For both parental groups, treatment was continued through gestation and lactation. Parameters evaluated in the parental rats included observations for clinical signs, body weight and food consumption measurements, gross necropsy, and microscopic examination of gross lesions and reproductive tract organs. Reproductive and litter parameters that were evaluated included male and female fertility indices, events at parturition, gestation duration, litter size, numbers of viable and stillborn pups, and pup survival and growth during lactation. Parental toxicity in the form of decreased body weight and food consumption was noted for males and females in the Fo and FI generations at 5000 ppm, and decreased body weight
63.3 Toxicology was noted for males in the Fo generation at 2000 ppm. A slight increase in hair loss was observed in the Fo and FJ females at 5000 ppm. In the F 1 adult males, kidney lesions including mottled kidneys, hyaline droplets, chronic inflammation, regenerative tubules, and renal granular cast formation were observed. These kidney effects occurred in a dose-related manner in all treatment groups and were characteristic of alpha2u-globulin nephropathy. No other treatment-related effects were observed in the parental generations at 500 ppm. Neonatal toxicity as evidenced by reduced pup sizes and weights in both generations was noted for males and females in the 5000 ppm group. No treatment-related effects were observed in pups at .::::2000 ppm. No treatment-related effects on reproduction or fertility were observed at any of the dose levels evaluated in this study. 63.3.4.2 Reproductive Toxicity Studies Referenced in the Open Literature Three other nonguideline, non-GLP studies that address reproductive toxicity are described in the open literature. One study was a subchronic study conducted in rats by the subcutaneous route of exposure that was designed to evaluate dominant lethal effects and male fertility (Wright et aI., 1992). A second was a rat subchronic inhalation study in which testicular weights and sperm morphology were evaluated (Macko and Bergman, 1980). The third was a rat subchronic dermal toxicity study in which sperm morphology was evaluated (Brusick, 1980). An increase in the incidence of abnormal sperm was reported in the latter two studies; however, the evidence for this finding is very weak. 63.3.5 CHRONIC TOXICITY AND ONCOGENICITY STUDIES 63.3.5.1 Dog Chronic Toxicity (Schoenig et al., 1999) DEET was administered orally for one year to purebred beagle dogs via gelatin capsules at 0, 30, 100, and 400 mg/kg/day. The daily dose of DEET was divided equally into two doses administered in the morning and afternoon following a I-hr period of food availability. Four male and four female dogs were evaluated at each dose level. Parameters evaluated in this study included observations for clinical signs, body weight and food consumption measurements, hematology, clinical chemistry, urinalysis, ophthalmology, organ weights, and gross and microscopic pathology. Treatment-related effects were observed only in the 400 mg/kg/day group and included emesis, ptyalism, decreased body weights, and decreased food consumption for both males and females. One male in the 400 mg/kg/day group also exhibited occasional ataxia, tremors, abnormal head movements, and convulsions. These clinical signs generally occurred shortly after dosing and were followed by complete recovery before the succeeding dose. Other treatment-related effects observed in the 400 mg/kg/day group included transient reduction
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in hemoglobin and hematocrit levels, increased alkaline phosphatase levels (males only), decreased cholesterol levels, and increased potassium levels (males only). 63.3.5.2 Mouse Oncogenicity Study (Schoenig et al., 1999) DEET was incorporated into the diet and administered for 78 weeks to Charles River CD® -1 mice at concentrations such that dose levels of 250, 500, and 1000 mg/kg/day could be evaluated. Sixty male and 60 female mice were evaluated at each dose level. In addition, two independent untreated control groups, each consisting of 60 male and 60 female mice, were included in the study. Parameters evaluated included observations for clinical signs and palpable masses, measurements, of body weight and food consumption, hematology, organ weight measurements, and gross and microscopic pathology. A slight decrease in body weight and food consumption was noted at the 1000 mg/kg/day dose level and an increase in liver weights was noted in male and female mice at 500 and 1000 mg/kg/day. The liver weight increases were considered to be adaptive in nature, since no corroborative microscopic findings were observed. No other treatment-related effects were observed and DEET administration had no effect on survival or tumor incidence. 63.3.5.3 Rat Chronic Toxicity/Oncogenicity (Schoenig et al., 1999) DEET was incorporated into the diet and administered for two years to Charles River CD® rats at dietary concentrations such that dosage levels of 10, 30, and 100 mg/kg/day for males and 30, 100, and 400 mg/kg/day for females could be evaluated. Lower dosage levels were selected for the males because it was felt that alpha2u-globulin induced renal lesions that were observed in male rats in the subchronic toxicity studies would jeopardize survival of the male rats at dose levels above 100 mg/kg/day. Sixty male and 60 female rats were evaluated at each dose level. In addition, two independent untreated control groups were included, each consisting of 60 male and 60 female rats. Parameters evaluated included observations for clinical signs and palpable masses, body weight and food consumption measurements, hematology, clinical chemistry, urinalysis, ophthalmology, organ weight measurements, and gross and microscopic pathology. Decreased body weight and food consumption and slightly increased cholesterol levels were observed in the female rats at 400 mg/kg/day. No other treatment-related effects were noted and DEET administration did not have any effect on survival or tumor incidence. 63.3.5.4 Chronic Toxicity/Oncogenicity Studies Referenced in the Open Literature Two other chronic toxicity/oncogenicity studies are reported in the open literature. In one of these studies, DEET was applied dermally to mice for 120 weeks (Stenblick, 1976). In the other,
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DEET was applied dermally to the ears of rabbits for 95 weeks (Stenback, 1976). No treatment-related increase in tumors was observed in either study. Both studies had many major deficiencies relative to today's standards for conducting chronic toxicity/oncogenicity studies. 63.3.6 NEUROTOXICITY STUDIES 63.3.6.1 Acute Neurotoxicity Study (Schoenig et al., 1993) The neurotoxicity potential of DEET was evaluated in Charles River Crl:CD® VAFlPlus® rats. A single dose of undiluted DEET was administered by oral gavage at dose levels of 0, 50, 200, or 500 mg/kg and the rats were observed for 14 days following dose administration. These dose levels were selected on the basis of findings in a dose range-finding study in which mortality was observed at a dose level of 1000 mg/kg and pharmacotoxic signs were evident at dose levels of 500 and 1000 mg/kg. Ten male and 10 female rats were evaluated at each dose level. Parameters evaluated included observations for clinical signs and measurements of body weight and food consumption. In addition, functional observational battery (FOB), thermal response test, and motor activity measurements were made at 1 hr, 24 hr, and 14 days following dose administration. No clinical signs of toxicity were observed in any animals and there were no treatment-related effects on body weight or food consumption. No effects related to DEET exposure were observed in the FOB. Animals in the 500 mg/kg dose group exhibited an increased response time to heat in the thermal response test and slightly decreased rearing activity in the motor activity test. These effects occurred in animals of both sexes at the 1-hr post-treatment observation interval but were not observed at 24 hr or 14 days post-treatment. No other treatmentrelated effects were observed during this study. 63.3.6.2 Chronic Neurotoxicity Study (Schoenig et al., 1993) The neurotoxicity potential of DEET was evaluated following multigeneration plus chronic administration. The animals used in this study were second generation (F2) offspring from the rat multigeneration reproductive toxicity study in which DEET was administered continuously over two generations at dietary concentrations of 0, 500, 2000, and 5000 ppm. The dietary concentration of 5000 ppm meets the criteria for a maximum tolerated dose (MTD) based upon traditional chronic toxicology assessments. Following weaning, all control offspring and two male and two female F2 generation pups from 21 to 25 litters per treatment group were selected and maintained on the treated diets for an additional 9 months. During the 9-month dietary administration period, the rats were observed for clinical signs, and body weight and food consumption measurements were made. At the end of this 9-month period, one male and one female from each of 20 litters were selected for the neurotoxicity evaluations. An additional control group of the 10 males and 10
females was selected from the rats maintained on basal diet for use as a sham group in the passive avoidance test. Neurotoxicity evaluations included a functional observational battery, motor activity testing, M-water maze, acoustic startle habituation, and passive avoidance. In addition, comprehensive neuropathological examinations were performed on 1 male and 1 female from each of 10 litters per group. The following tissues were included in this neuropathological examination: forebrain, center of cerebrum, midbrain, cerebellum, pons, medulla oblongata, proximal sciatic nerve, sural nerve, tibial nerve, spinal cord at cervical swelling (C3-C6) and lumbar swelling (LI-L4), gasserian ganglia, dorsal root ganglia (C3-C6, LI-L4), and dorsal and ventral root fibers (C3-C6, LI-L4). At the time that these neurotoxicity evaluations were performed, all animals were approximately 40 weeks of age. No clinical signs of toxicity were noted throughout the 9-month dietary administration period; however, decreased body weights relative to the control animals were observed for all treatment groups. These findings were somewhat unexpected at 500 and 2000 ppm, since decreased body weights in the original multigeneration study were observed only at 5000 ppm. The decreased body weights at 500 and 2000 ppm may have been due to random selection of the animals at study start, since the treatment groups were not balanced with respect to animal weight at the time of selection. In the in-life neurotoxicity evaluation, the only treatment-related finding was a transient increase in exploratory locomotor activity at 5000 ppm. Equivocal findings were noted in the M-water maze, in which a decrease in initial choice accuracy on reversal was observed in all of the treated groups as compared with the controls. However, the effect was not dose-dependent, and no confirmatory evidence was found for a DEET-related reversal learning effect on response times and total errors, which are the primary measures of performance recorded for this task. Therefore, this finding was not considered to be adequate evidence of an effect of DEET on learning. No treatment-related effects were observed in the comprehensive neuropathological examinations performed on the central and peripheral nervous tissues. Since the only potential neurotoxic effect that was observed in this study was observed at 5000 ppm, a dose level at which other toxic effects have been observed, the results of this study demonstrate that the nervous system does not appear to be a selective target when DEET is administered chronically at dose levels up to and including the MTD. In addition, the results of this study demonstrate that the chronic administration of DEET at a MTD dose does not result in any morphologic changes in the tissues of the nervous system. 63.3.6.3 Studies Referenced in the Open Literature that Describe Neurotoxic Effects in Laboratory Animals Six other studies that describe neurotoxic effects associated with DEET administration are described in the open literature. One study is an acute inhalation study conducted in rats (Sherman, 1980). A second is an acute oral toxicity study in rats
63.3 Toxicology (Verschoyle et aI., 1992). A third is a subchronic inhalation study in rats (Macko and Bergman, 1980). The fourth is a subchronic oral toxicity study conducted in rats (Campbell, 1986). The fifth is a developmental toxicity study conducted in rats by the subcutaneous route of administration (Wright et aI., 1992). The sixth is a study designed to evaluate dominant lethal effects and male fertility in which DEET was administered subcutaneously to male rats (Wright et aI., 1992). None of these studies came close to meeting current standards for neurotoxicity testing and most reported effects were observed at lethal or near lethal dose levels. 63.3.7 GENOTOXICITY STUDIES
63.3.7.1 SalmonellalMammalian- Microsome Plate Incorporation Assay (Ames Test) (San and Schadly, 1989) Five strains of Salmonella typhimurium, TA98, TA 100, TA1535, TA1537, and TA1538, were used. Each strain was tested in the presence and absence of metabolic activation by a rat liver S-9 system induced with Aroclor 1254. The concentrations of DEET evaluated in these studies (with and without metabolic activation) ranged from 278 to 8333 )l.g/plate. Mutagenic frequency did not increase in any of the tester strains. Results from the initial assay were confirmed in an independent repeat assay. 63.3.7.2 Chromosomal Aberrations in Chinese Hamster Ovary Cells (Putman and Morris, 1989) This assay was conducted in Chinese Hamster Ovary cells in the presence and absence of metabolic activation with a rat liver S-9 fraction after induction by Aroclor 1254. In the absence of metabolic activation, the concentrations of DEET evaluated were 0.063, 0.125, 0.25, 0.5, and I )l.lIml while, with activation, the concentrations of DEET evaluated were 0.032, 0.063, 0.125, 0.25, and 0.5 )l.lIml. No increase in chromosomal aberrations was observed with or without metabolic activation. 63.3.7.3 Unscheduled DNA Synthesis in Rat Primary Hepatocytes (Curren, 1989) This assay was conducted in rat primary hepatocytes. DEET concentrations of 0.003, 0.01, 0.03, 0.1, and 0.2 )l.lIml were evaluated in an initial assay and DEET concentrations of 0.01, 0.03,0.1,0.2, and 0.3 )l.lIml were evaluated in a confirmatory assay. No increase in DNA synthesis was observed in either assay. 63.3.7.4 Genotoxicity Studies Referenced in the Open Literature One study evaluated reverse gene mutation in Salmonella typhimurium (Zeiger et aI., 1992) and two studies evaluated both reverse gene mutation in Salmonella typhimurium and Saccharomyces cerevisiae (Brusick, 1976; Macko and Weeks, 1980).
1447
A dominant lethal assay in Swiss white mice is reported by Swentzel (1978). No significant activity was observed in any of these studies 63.3.8 PHARMACOKINETIC STUDIES 63.3.8.1 Absorption, Distribution, Metabolism, and Excretion in Rats (Schoenig et al., 1996) Absorption, distribution, metabolism, and excretion (AD ME) was evaluated in Charles River CD@ rats following oral or dermal administration of 4C]DEET. Six experiments were conducted using separate treatment groups, each consisting of five male and five female rats. Four experiments involved the evaluation of ADME patterns following administration of 4C]DEET orally as a single low dose of 100 mg/kg, orally as a single high dose of 500 mg/kg, orally as a single low dose following 13 days of oral dosing at 100 mg/kg/day with nonradiolabeled DEET, and dermally as a single low dose of 100 mg/kg. Urine and feces were collected over a 7-day posttreatment period, after which the animals were humanely sacrificed and selected tissues and organs were harvested. Urine, feces, and tissues were analyzed for 14C content and the major urinary metabolites were identified. The remaining two experiments examined the distribution of radioactivity in tissues of animals humanely sacrificed at peak 14C blood levels after receiving a single dose of 100 mg/kg [14C]DEET by the oral or dermal route of administration. In the three experiments designed to determine the ADME patterns of DEET after oral administration, 85 to 91 % of administered radioactivity was found in the urine and 3 to 5% was found in the feces. Quantitatively, the overall amount of radioactivity excreted into the urine and feces was similar for males and females in the three groups, but the rate at which radioactivity appeared in the urine differed significantly between the three dosing regimens. The fastest rate of excretion of radioactivity into the urine was observed in the rats that received the repeated oral low dose, followed by the single oral low dose and the single oral high dose groups. This pattern of excretion suggests that the rate of excretion is closely related to the rate of metabolism and that repeated dose administration induces the enzymes responsible for the metabolism of DEET. In the group of rats that received the dermal low dose, 74 to 78% of administered radioactivity was found in the urine and 4 to 7% was found in the feces, with an additional 6.5% found on the skin surface at the application site or associated with the occlusive enclosure. Slightly less radioactivity was recovered in the urine and slightly more was recovered in the feces for females than for males in this dosing regimen. Overall, the rate of excretion of radioactivity into the urine and feces was much slower after dermal administration than after any of the oral dosing regimens. Total tissue residues of 14C activity at 7 days post-treatment ranged from 0.15 to 0.67% of administered radioactivity for all four dosage regimens. At peak 14C blood levels (0.5 hr for
e
e
1448
CHAPTER 63
DEET
males and 2 hr for females), the percent of administered radioactivity reaching systemic circulation and the tissues was much higher for animals administered 4C]DEET orally than for animals administered [14C]DEET dermally. In both cases, the only tissues with 14C residues consistently higher than plasma 14C residues were the liver, kidney, and fat. DEET was metabolized completely in all treatment groups, with little or no parent compound excreted in the urine. Two major urinary metabolites were identified in the urine by mass spectroscopy as m-[(N,N-diethylamino)carbonyl]benzoic acid and m-[(ethylamino)carbonyl]benzoic acid. Both metabolites involved oxidation of the aromatic methyl substituent in the DEET molecule to a carboxylic acid moiety. In addition, one of the metabolites also had undergone N -dealkylation of an ethyl substituent on the amide moiety. Based upon the metabolites observed, a metabolic pathway was proposed for DEET as shown in Fig. 63.2.
e
63.3.8.2 Absorption, Metabolism, and Excretion in Humans (Selim et al., 1995) The absorption, metabolism, and excretion of DEET following dermal application of 4C]DEET to male human volunteers was evaluated. DEET was applied to the forearm of two groups of six volunteers either as the undiluted technical grade material or as a 15% (w/w) solution in ethanol. The material was left in contact with the skin for 8 hr and then rinsed off the skin. Serial blood samples and all urine and feces were collected for 5 days after application. The application sites also were stripped with
e
tape at 1,23, and 45 hr following rinsing. All samples collected were analyzed for total radioactivity. Urine samples also were analyzed by HPLC for metabolite characterization. Absorption of DEET as evidenced by the appearance of radioactivity in the plasma occurred within 2 hr after dermal dose administration. Elimination of radioactivity from the plasma was rapid, with measurable levels of radioactivity remaining in the plasma only up to 4 hr after the application period. Most of the excreted radioactivity was found in the urine, with an average recovery of 5.61 and 8.33% of the applied dose in the undiluted DEET and 15% DEET in ethanol groups, respectively. Less than 0.08% of the applied dose was recovered in the feces in both groups. Total excretion in the urine and feces of applied radioactivity ranged from 3 to 8% with a mean of 5.6% in the volunteers that applied undiluted DEET and from 4 to 14% with a mean of 8.4% in the volunteers that applied 15% DEET in ethanol. The recovery of radioactivity in the tape strips was very low, indicating that DEET did not accumulate in the superficial layers of the skin. Most of the applied radioactivity was recovered in the skin rinses. Absorbed DEET was metabolized completely and six major metabolites were observed in the urine. Based upon a comparison of their HPLC retention times with the two major rat urinary metabolites, the two human urinary metabolites tentatively were identified as m-[( N ,N -diethylamino )carbonyI]benzoic acid and m-[(ethylamino)carbonyl]benzoicacid. The results of this study showed that DEET is absorbed slowly and excreted rapidly when applied to human skin. It also shows that DEET is metabolized completely prior to urinary excretion and that the
METABOLITE A
DEET
1
1 O~
COOH METABOLITE B
Figure 63.2
Proposed metabolic pathway for DEET in rats (Selim et al.. 1995).
63.4 Human Aspects metabolic pattern in humans is similar to that observed in rats. The latter finding is important since it supports the relevance of the findings observed in the toxicology studies conducted in this animal model. 63.3.8.3 Studies in the Open Literature that Address the Pharmacokinetics of DEET A large number of other nonguideline, non-GLP studies that were designed to evaluate one or more of the pharmacokinetic properties of DEET are available in the open literature. These studies were conducted by the dermal route of administration in mice (Blomquist and Thorsell, 1977), rats (Moody and Nadeau, 1993; Moody et aI., 1989; Snodgrass et al., 1982), guinea pigs (Moody and Nadeau, 1993; Schmidt et aI., 1959), rabbits (Hanniba1, 1992; Snodgrass et al., 1982), dogs (Reifenrath et aI., 1980, 1981; Snodgrass et aI., 1982), cattle (Taylor et aI., 1994), monkeys (Moody et aI., 1989), and humans (B10mquist and Thorsell, 1977; Feldman and Maibach, 1970; Wu et aI., 1979), by intravenous injection in mice (Blomquist et aI., 1975), and using in vitro methods (Baynes et al., 1997; Taylor, 1986; Windheuser et aI., 1982; Yeung and Taylor, 1988). Overall, these studies, along with those conducted by the DJV, show that DEET is well absorbed by the oral route of administration in all animal species. By the dermal route of exposure, absorption ranged from 7.9 to 92.5% in the vari-
1449
ous laboratory animal species. In humans, dermal absorption ranged from 4.6 to 16.7%. The site of application, the method of application (occluded or unoccluded) and vehicle were shown to influence dermal absorption. There is little evidence for tissue accumulation and, for the most part, DEET was found to be quantitatively metabolized and excreted in the urine.
63.4 HUMAN ASPECTS 63.4.1 CLINICAL CASE REPORTS 63.4.1.1 Dermal Exposure The analysis of the available case reports on DEET is difficult because of limited clinical details provided. Nonetheless, these reports often are cited in discussions of DEET safety. Over the past 40 years, 14 reports have appeared in the medical literature associating clinical illness with dermal DEET exposure in 20 individuals (Osimitz and Murphy, 1997). Fourteen individuals had neurologic symptoms (Table 63.1), one had an acute manic psychosis (Snyder et aI., 1986), one had anaphylaxis (Miller, 1982), and one had a cardiovascular event (Clem et aI., 1993). Three were newborns who were reported to have alleged teratogenic effects from DEET (Hall et aI., 1975). Thirteen of the reported 14 patients with neurologic symptoms were less than 8 years old (Table 63.1). The DEET con-
Table 63.1 Case Studies in which Neurological Symptoms Were Reported Following the Topical Application of DEET Sex/Age %DEET
Pattern of
Patient
(y)
used
use
Symptoms
Laboratory
Outcome
diagnosis
1 (Gryboski et aI., 1961)
F/3.5
15
Daily x 2
Seizures
Normal CSF
Full
Idiopathic seizure
Possible alternative
wks 2 (Hall et aI., 1975) 3 (Hall et aI., 1975)
F/5
F!7.5
10 10
recovery
Nightly x 3
Headaches, ataxia, 185 WBC in CSF
mos
seizures, agitation
Application
Opisthotonos
F/8
15
Encephalitis, parainfectious encephalopathy
?? cells
& ingestion 4 (Zadikoff, 1979)
Death
10 occasions Headaches, ataxia, Hyperarnmonemia
Full
Encephalitis, parainfectious
recovery
encephalopathy
Death
OCT heterozygote
disoriented 5 (de Garbino and Laborde, 1983) FlU
NA
Frequent
"acute
Normal
Death
encephalopathy" 6 (Roland et aI., 1985)
F/8
15 and
Copius X4d Rash, restlessness, Abnormal EEG
100 7 (Edwards and Johnson, 1987)
FlU
20
seizures 3 mos
Recovery Idiopathic seizure nonspecific rash
Ataxia, movement 14 WBCin CSF
Recovery Encephalitis, parainfectious
disorder, drooling,
encephalopathy, myclonic
opisthotonos,
encephalopathy
opsoclonus, myoclonus 8-11 (Oransky et aI., 1989)
Ml3-7
NA
NA
Seizures
Normal
Recovery Idiopathic seizure
12 (Oransky et aI., 1989)
Ml28
NA
NA
Seizures
Normal
Recovery Idiopathic seizure
13 (Oransky, 1991)
?/8
NA
Hours
Seizure
Unknown
Unknown Idiopathic seizure
14 (Lipscomb et aI., 1992)
MI5
100 and 15 Brief
Seizures
Normal CSF, normal CT Recovery Idiopathic seizure
Key: NA, not available; WBC, white blood cells; OCT, ornithine carbamoyl transferase; EEG, electroencephalogram; CSF, cerebrospinal fluid; CT, cranial computed tomography.
1450
CHAPTER 63
DEET
centration of the products used is known for 7 of the 14 patients. Of these 7, 5 developed symptoms following use of products with less than 20% DEET. The other two had applications of both 15% and 100% DEET. The most commonly reported neurologic symptom was convulsions. This was the only symptom in eight patients (patients 1 and 8 through 14), leaving the impression that a single seizure was the only symptom. All eight of these patients recovered. One 8-year-old girl had seizures and a rash on her extremities, with spontaneous resolution of symptoms (patient 6). She is the only patient in this series with a rash. Since exanthematous illnesses in children are not infrequently associated with convulsions, an infectious disease is not excluded (Modlin, 1995). In pediatric patients, single seizures in otherwise normal children may be idiopathic events. In fact, about 2% of U.S. children have had isolated, idiopathic, nonfebrile seizures by the time they reach age 10 (Annegers, 1993). Considering that, if 2% of the 37,459,000 children in the United States that are less than 10 years old (or 750,000 children) have had single seizures by the age of 10 and, if 25% of these events (or 187,500) occur during the three summer months when DEET is most used (hence, approximately 18,750 seizures among this cohort each summer), on many occasions the use of DEET and appearance of a seizure may be temporally associated only by chance. It is interesting to note that there are no reports of DEET seizures in children with epilepsy (close to 1% of children; Annegers, 1993). These children may be exposed to DEET as frequently as children without epilepsy, and the risk in this population merits evaluation. Three patients, one of whom died, had more complex neurologic symptoms (patients 2, 3, and 7). All had cerebrospinal fluid pleocytosis, suggesting either an inflammatory response to DEET or the possibility of infectious (Whitley, 1994) or parainfectious disease (Marks et a!., 1988) processes involving the central nervous system. Patient 4 was a heterozygote for ornithine carbamoyl transferase (OCT) deficiency and suffered neurologic symptoms, coma, and death following DEET exposure. She had suffered prior hyperammonemic episodes not associated with the use of DEET. Heterozygosity for OCT deficiency is known as a potentially lethal hyperammonemic condition (Rowe et a!., 1986). The specific effect of DEET on ammonia metabolism is unknown, but DEET does not inhibit human OCT in vitro (Rej et a!., 1990). In summary, 9 of the 14 patients with neurological symptoms (patients 1, 6, and 8-14) may have had idiopathic seizures, one may have had an exanthematous illness and a convulsion, three may have had an inflammatory process affecting the central nervous system, and one was heterozygous for OCT deficiency but synergism with DEET is not excluded. Insufficient information is available on patient 5 to determine if there are alternate explanations for the patient'S encephalopathy. The authors of 7 of the 10 reports of neurologic adversity after DEET exposure were contacted in 1997 and no additional information on this cohort of patients was available (Osimitz and Murphy, 1997).
Of the six clinical cases not involving neurologic symptoms, one describes a woman with anaphylactic symptoms following brief exposure to DEET who also had very similar symptoms following re-exposure to DEET in an emergency room setting (Miller, 1982). She is thought to represent an allergic hypersensitivity to DEET. Given the unique nature of the events in the other five patients, i.e., acute manic psychosis (Miller, 1982), a cardiovascular event (Clem et a!., 1993), and fetal malformations (Hall et aI., 1975; Schaefer and Peters, 1992), as well as the possibility of alternative etiologies (e.g., gestational urinary tract infection; Schaefer and Peters, 1992) and familial predisposition (Hall et a!., 1975), it is difficult to establish a cause and effect relationship. 63.4.1.2 Oral Ingestion
In addition to the above case reports that involved dermal application, two investigators have reported adverse reactions to intentional oral ingestion of DEET by six patients (Fraser et a!., 1995; Tenenbein, 1987). Of the five patients described by Tenenbein, two died and three resolved with no sequelae. The common symptoms included coma, seizures, and hypotension observed within 1 hr of ingestion. Alcohol and/or other drugs were present in three of the five patients. The two patients who died had serum DEET levels of 1.68 mg/ml and 2.4 mg/ml. This compares with a recent report showing peak serum levels of 0.52 J.1g/ml in humans following a single dermal application of 100% DEET at the 95th percentile of human use (Osimitz et aI., 1997). One of the fatalities reported by Tenenbein had a blood ethanol level of 13.0 mg/ml and was also positive for cannabinoids. The other individual who died was found with empty bottles of chlorpromazine-HCI and hydralazine-HCl. Fraser et a!. (1995) reported the intentional ingestion of 15 to 25 ml 95% DEET by a 19-year-old woman with a history of psychiatric disorders. An EKG shortly after admission to the hospital (1 to 2 hr after ingestion) showed left and right atrial enlargement, diffuse ST-T abnormalities and a normal QT interval. The EKG was normal within 12 hr and the patient recovered completely. Analysis of the serum indicated a DEET level of 63 mg/liter about 2 hr after ingestion. Because of the intentional nature of the ingestions reported above and the fact that peak plasma levels were in excess of 1000 times higher than that seen during normal use of DEET, such cases are not relevant for the assessment of adverse effects following typical consumer use of DEET. 63.4.1.3 Occupational Exposure
The Hazards Evaluation and Technical Assistance Branch of The National Institute for Occupational Safety and Health used a survey interview to evaluate occupational exposure to DEET and self-reported health effects among adults employed at Everglades National Park in Florida (McConnell et a!., 1986). Higher exposure to DEET was associated with a higher prevalence of insomnia, muscle cramps, mood disturbances, rashes, or difficulty in micturition. The subjects with higher exposure also were more likely to work at sites distant from the park
63.5 Risk Assessment Considerations
headquarters, to use pesticides on a regular basis, to consume more alcohol, and to have had a job with prior chemical exposure. Whether any of the factors, alone or in combination with DEET, could account for the symptoms, was not explored. A follow-up comparison of the workers pre-exposure (March) and post-exposure (August) did not support the findings of the initial survey. Any relationship, if it existed at all, between DEET exposure and reported neurologic symptoms was confounded by other factors. 63.4.1.4 Poison Control Center Data Further information of the human experience with DEET comes from a review of adverse effect reports following DEET exposure from the 71 Poison Control Centers (PCCs) participating in the American Association of Poison Control Centers' National Data Collection System between 1985 and 1989 (Veltri et aI., 1994). The participating centers covered a wide area of the United States with a population of over 180 million people. Trained poison control specialists assessed the cases as minor, moderate, or major at the conclusion of patient contact. There were 9086 human exposures involving DEET-containing insect repellents reported to PCCs from 1985-1989. Of these, 98.9% experienced either no effect or had short-lived symptoms. These involved mild irritation to the skin or mucous membranes. Sixty-six patients had symptoms classified as moderate, i.e., more pronounced or more prolonged than the minor effects, but all symptoms resolved without sequelae. Five patients had symptoms classified as major. All five patients experiencing a major effect were exposed to products containing 11-50% DEET. Two of the five patients experienced eye irritation which was treated at home. One patient, a 17 -yearold male who had saturated his clothing with 17.9% DEET, was ataxic and may have had a seizure. A 33-year-old male reported diminished sensation and hypotension one week after using a DEET product. Both were discharged after emergency room evaluation and recovered fully. The fifth patient had a dystonic reaction responding to diphenhydramine. Earlier in the day, he had ingested prochlorperazine, a known cause of dystonia, but synergism with DEET is not excluded. One patient ingested 8 oz. of DEET in a successful suicide attempt. Symptoms included cardiorespiratory arrest and status epilepticus. Beginning in 1995, the DJV contracted Pegus Research, Inc. (Salt Lake City, UT) to establish the National Registry of Human Exposures to DEET. The purpose is to collect detailed information from individuals who use DEET-containing insect repellents and report serious adverse effects that are either neurologic or systemic. The Registry has been designed to overcome some of the limitations inherent in retrospective analysis of PCC data. Because of its prospective nature, the Registry allows for the quick and thorough follow-up on individual cases to determine the circumstances surrounding the case, the medical features, and to attempt to determine the relationship between the exposure and the symptoms (causality). The results of this ongoing effort, which will be published shortly in the literature, will provide important insight into the human safety of DEET.
1451
63.5 RISK ASSESSMENT CONSIDERATIONS 63.5.1 INTRODUCTION The risk assessment for DEET can be approached in a number of different ways. One is the traditional method of defining toxicity endpoints of concern from the existing toxicology database and comparing the no-effect levels for these end points to some estimate of human exposure. In the case of DEET, there are at least two problems associated with this approach. One is that little or no toxicity can be produced in laboratory animals by the principal (dermal) route of exposure, to which humans are exposed. The second is that, while toxic effects can be produced by the oral, inhalation, and subcutaneous routes of exposure, because of the large differences in how DEET would be expected to be handled in the body by these other routes of administration, it is questionable as to whether toxicity end points identified by these other routes of administration are applicable to human risk assessment. For example, it has been demonstrated that neurotoxic effects can be produced in laboratory animals when DEET is administered orally as a bolus dose. However, as data that are presented below show, if these effects are related to the peak levels of DEET in the blood, then they probably have little or no applicability to human risk by the dermal route of exposure. On the other hand, since it has been shown that DEET is not a teratogen, a reproductive toxin, or an oncogen at maximum tolerated doses by the more rigorous oral route of administration, a risk assessor can feel quite confident that developmental, reproductive, or oncogenic effects are not toxicity endpoints of concern for DEET. This is the type of reasoning that went into U.S. EPA's recent decision, as stated in the DEET Reregistration Eligibility Document, that there are no significant endpoints of toxicity concern for DEET and that a comprehensive, quantitative risk assessment is not warranted (U.S. EPA, 1998). While the DJV agreed with the approach taken by the U.S. EPA, it felt that additional data were needed to aid in the risk assessment process. Therefore, five blood level studies designed to define the profile of systemic exposure that occurs under four different exposure scenarios were undertaken. The exposure scenarios that were evaluated were: (1) in humans following single and repeated dermal applications of DEET at the 95th percentile of human exposure; (2) in dogs under the conditions in which neurotoxic signs were observed following single and repeated oral dose administration via gelatin capsules; (3) in rats under the conditions in which potential neurotoxic effects were observed following acute oral administration by gastric gavage; and (4) in rats under the conditions of the 90-day dermal toxicity study. These studies, in which systemic exposure was measured in terms of DEET plasma levels, were intended to demonstrate two points. One is that the systemic exposure laboratory animals receive following oral dose administration (especially oral bolus dosing) is very different than that which humans receive by the dermal route of exposure. This is an important consideration for the risk assessment process since treatment-related
1452
CHAPTER 63
DEET
effects have been observed in a number of oral toxicity studies. A second point that these blood level data were intended to demonstrate is that the subchronic dermal toxicity studies conducted in rats and micropigs provide the most meaningful data for human risk assessment. The design and results of these blood level studies are discussed below. 63.5.2 BLOOD LEVEL STUDIES 63.5.2.1 Blood Level Study to Define DEET Plasma Profile in Humans Following Single and Repeated Dermal Applications of DEET at the 95th Percentile of Human Exposure (Ohayon et al., 1997) In this study, three male and three female human volunteers were administered undiluted DEET by the dermal route of administration at the 95th percentile of human use (3 g/day for females and 4 g/day for males). Both single and repeated (8 hr per day for four consecutive days) applications were evaluated. DEET plasma levels were profiled on the first and fourth days of the study. The DEET plasma profiles for humans under this scenario are presented in Fig. 63.3 and are summarized in Table 63.2.
The findings from this study show that DEET does not appear in the blood of humans until 1 to 2 hr after dermal application, after which time the DEET plasma levels gradually increase until the material is washed off 8 hr after application. The DEET plasma profiles and peak plasma levels were similar in males and females and did not increase after repeated dosing. The overall mean peak plasma level was 0.45 ).tg/ml. The overall mean area under the DEET plasma concentration versus time curve (AUC) was 3.51 ).tghr/ml. 63.5.2.2 Blood Level Study to Define DEET Plasma Profile in Dogs under Conditions in which Neurotoxic Signs Were Observed Following Single and Repeated Oral Dose Administration (Badalone, 1997) As discussed previously in Section 63.3.2.5 clinical signs indicative of neurotoxicity were observed in two studies in which undiluted DEET was administered to dogs as a single oral bolus dose via gelatin capsules. In order to compare the systemic exposure that the dogs received under the dosage regimen employed in these studies to that of humans under normal conditions of use, a blood level study was conducted in dogs ad-
20 0.8
A
A 15
0.7 0.6 0.5 0.4 0.3 0.2
0.1
o
0.0
10
12
14
16
18
20
22
20 0.8
3
4
24
B
B
0.7
15 0.6 0.5 0.4 0.3 0.2
0.1 0.0
o Time After Application (hours)
Figure 63.3 DEET plasma profile in human volunteers administered DEET dermally at the 95th percentile of human use during and following the first of four (A) and fourth of four (B) daily 8-hr dermal applications. Each value represents the mean of six individual values (three males and three females).
4
Time After Administration (bours) Figure 63.4 DEET plasma profile in dogs during and following the first of four (A) and fourth of four (B) daily oral bolus doses of DEET at a level of 75 mg/kg/day. Each value represents the mean of eight individual values (four males and four females).
1453
63.5 Risk Assessment Considerations 24
A
22 20
18 16 14 12
10
~
00
+1
6
Cl
...
::;'"
4
.:s
~
0
-S
I
0
30
60
90
120
150
180
210
240
60
90
120
150
180
210
240
Cl
c
'=
'"
40
c.I
35
~...
B
Cl
c
u
30
'"
~
25
Eo<
20
"""" Q
15 10
o 0
30
ministered DEET as a daily oral bolus dose via gelatin capsules at the NOAEL for effects in the dog studies, i.e., 75 mg/kg/day. Four male and four female dogs were utilized in this four day study. DEET plasma levels were profiled on the first and last days of this study. The DEET plasma profile for dogs under this dosing regimen is presented in Fig. 63.4 and is summarized in Table 63.3. In this study, both the blood profiles and peak blood levels observed in these dogs were much different than those observed in humans. For example, peak plasma levels were observed within 30 minutes after dosing and plasma levels were back below the limit of quantitation within 3 to 4 hr after dose administration. Consistent with humans, no differences were observed between male and female dogs and there was no evidence of accumulation of DEET in the blood following repeated doses . The overall mean peak plasma level was 14.7 I-lg/ml. The overall mean AUC was 12.59 I-lg hr/ml. Since the effects that were observed in the dogs occurred shortly after dosing, they appear to be closely associated with the time frame in which peak plasma levels would have occurred. A comparison of the overall mean peak plasma levels between dogs and humans showed a 33-fold difference, i.e., 14.7 versus 0.45 I-lg/ml. Since these plasma level data provide a direct comparison of the true systemic exposure, all of the un-
Time After DEET Dosing (minutes)
A
Figure 63.5 DEET plasma profile over a 4-hr period in male (A) and female (B) rats following a single oral bolus dose of DEET at a dose level of 200 mg/kg. Each value represents the mean of three individual values. 12
A
10
O~~--
o
__
--,---~~--
10
__- -__ 12
14
~--
__- -__
16
18
20
--~~
22
24
"~~~~~~~~~~~~~~~~~~~~
10
11
12
24 , 48
40
B 35
B
30
25
20 15
10
~~--~~--~~------~~~~------~~~ 10
11
12
24,48
Time After DEET Application (hours)
Time After DEET Dosing (hours)
Figure 63.6 DEET plasma profile over a 48-hr period in male (A) and female (B) rats following a single oral bolus dose of DEET at a dose level of 200 mg/kg. Each value represents the mean of five individual values.
Figure 63.7 DEET plasma profile in male (A) and female (B) rats during and following the first and fifth of five consecutive dermal applications of DEET at a dose level of 1000 mg/kglday. Each value represents the mean of four individual values. (e) First application. (0) Fifth application.
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DEET
Table 63.2 DEET Plasma Levels in Human Volunteers Administered DEET Dermally at the 95th Percentile of Human Use DEET exposure metric
Single exposurea
Repeated exposure b
Male
Female
Mean
Male
Female
Mean
Plasma level ().!g/ml)
0.62
0.43
0.52
0.49
0.25
0.37
AUCc ().!g hr/ml)
5.11
3.07
4.09
3.76
2.10
2.93
Mean maximum
Table 63.3 DEET Plasma Levels in Dogs Administered DEET in Gelatin Capsules at a Dose Level of 75 mglkg/day DEET exposure metric
Single exposure
Q
Repeated exposure b
Male
Female
Mean
Male
Female
Mean
Plasma level ().!g/ml)
18.26
13.97
16.11
14.25
12.48
13.36
AUC ().!ghr/mlY
15.87
13.45
14.66
11.21
10.17
10.69
Mean maximum
a From data collected during and following the first of four daily dermal applications of technical grade DEET at dose levels of 4 g/day for males and 3 g/day for females. bFrom data collected during and following the fourth of four daily dermal applications of technical grade DEET at dose levels of 4 g/day for males and 3 glday for females. C AUC = area under the DEET plasma concentration versus time curve.
certainties associated with making similar comparisons based on whole body exposure on a mg/kg basis have been eliminated. Because of this reduction in uncertainty, a lO-fold difference in plasma levels generally is considered to be at least equivalent to a lOO-fold safety factor or margin of exposure (MOE) based on whole body exposure on a mg/kg basis. 1 Therefore, the observed 33-fold difference in overall mean peak plasma levels is at least equivalent to a 300-fold safety factor or MOE developed on the basis of whole body exposure. The data obtained in this dog blood level study also are applicable to the eight-week and chronic dog studies since peak blood levels were observed within 30 minutes after dosing and blood levels returned to baseline within 3 to 4 hr in the blood level study. Because the dogs were fed just prior to dose administration in the subchronic and chronic studies, the peak blood levels after the a.m. dose may have been slightly lower due to the influence food may have had on absorption. However, if this were the case, the peak blood levels following the p.m. dose probably would have been as high or higher because the blood levels probably would not have returned to baseline prior to the p.m. dose. 63.5.2.3 Blood Level Study to Define Plasma Profile in Rats Administered DEET under the Conditions in which Potential Neurotoxic Effects Were Observed Following Acute Oral Administration by Gastric Gavage (Goldenthal, 1999; Laveglia, 1998a) As discussed in Section 63.3.6.1 an increase in time to respond in a thermal response test and decreased rearing activity detected by two measures of vertical movement (vertical activity I While these types of blood level data are not usually available for pesticides and insect repellenys, they are routinely developed and used for risk assessment for pharmaceutical products by U.S. FDA. A lO-fold difference in plasma levels is used as the benchmark for safety for pharmaceutical products in much the same manner as other regulatory agencies use an MOE of 100 for nondrug products such as pesticides.
aFram data collected during and following the first of four daily oral doses of
technical grade DEET. bFrom data collected during and following the fourth of four daily oral doses of technical grade DEET. C AUC = area under the DEET plasma concentration versus time curve.
and vertical time) in a motor activity test were observed in a rat acute neurotoxicity study. These findings were observed in animals of both sexes but only at the l-hr post-treatment time point and only at the 500 mg/kg dose level. Other time points at which these parameters were evaluated were 24 hr and 14 days post-treatment. The NOEL in this study was 200 mg/kg. In order to compare the systemic exposure to DEET that rats received under this dosage regimen to that of humans under simulated conditions of human use, a blood level study was conducted at the highest level at which no effects were observed in the rat study, i.e., 200 mg/kg. Eight groups, each consisting of three male and three female rats, were employed in this study. One group of rats was humanely sacrificed for blood collection and subsequent DEET plasma analysis at each of eight different time intervals over a 4-hr time period following oral dose administration. The DEET plasma profiles for male and female rats under the scenario described above are presented in Fig. 63.5 and are summarized in Table 63.4. As was the case with dogs, the blood profiles and peak blood levels observed in these rats were much different than those observed in humans under normal use conditions. For example, peak plasma levels were observed within 15 to 45 minutes after dosing, after which time a plateau level or gradual decrease in plasma levels was observed during the 4-hr time period in which the plasma levels were measured. Average peak plasma levels were 9.58 !l-g/ml in male rats and 13.61 !l-g/ml in female rats. AUC values could not be determined because DEET plasma levels did not return to baseline over the 4-hr period in which they were measured in this study. However, since the effects that were observed in the rats occurred shortly after dosing, they appear to be closely associated with the time frame during which peak plasma levels would have been expected to occur in the rat acute neurotoxicity study rather than AUC. A comparison of peak plasma levels between the rats and humans shows 21- to 30-fold differences. In order to characterize more fully the time course of elimination of DEET from the plasma following an oral bolus dose, a second oral blood level study was conducted in rats. In the second study, the same dose level (200 mg/kg) was adminis-
63.5 Risk Assessment Considerations Table 63.4 DEET Plasma Levels in Rats Administered DEET by Gastric Gavage at a Dose Level of 200 mg/kg
Table 63.6 DEET Plasma 1000 mg/kg/day
DEET exposure metric
DEET
Male
Female
9.58
13.61
exposure metric
Mean maximum Plasma level a Peak
(~g/ml)Q
plasma level during a 4-hr period following dose administration.
Table 63.5 DEET Plasma Levels in Rats Administered DEET by Gastric Gavage at a Dose Level of 200 mg/kg DEET exposure metric
Levels
Male
Female
in
Rats
Administered
DEET
1455
Dermally
Single exposureQ
Repeated exposureb
Male
Male
Female
at
Female
Mean maximum plasma level AUC
(~g/ml)
(~g hr/ml)C
4.56 89.5
24.37 368.2
4.37 100.8
17.14 134.8
aFrom data collected during and following the first of four daily dermal applications. bFrom data collected during and following the fourth of four daily dermal applications. . . C AUC = area under the DEET plasma concentratIOn versus time curve.
Mean maximum Plasma level
(~g/ml)a
AUC (~ghr/ml)b
7.06
15.18
22.91
79.18
apeak plasma level during a 48-hr period following dose administration. b AUC = area under the DEET plasma concentration versus time curve (48-hr time period).
tered; however, the DEET plasma profile was evaluated over a 48-hr rather than a 4-hr period of time, and groups consisting of five male and five female rats were humanely sacrificed at each time interval. The DEET plasma profiles for rats in this study are presented in Fig. 63.6 and are summarized in Table 63.5. The data from this study showed that peak plasma levels occurred 30 minutes after dosing and that there was a distinct difference between both male and female rats in both maximum plasma levels and AUC. In this study, the elimination of DEET from plasma after reaching peak levels was biphasic, indicating that DEET was rapidly absorbed from the stomach, distributed in the body within 2 to 4 hr of administration, and subsequently eliminated from plasma within 12 hr. A comparison of peak plasma levels between rats and humans shows that the levels in male and female rats under these experimental conditions were 16 to 34 times higher than in humans, respectively. The data developed in these rat acute oral blood level studies also can be used to support the position that the clinical signs observed in the rat teratology study are not relevant to human health. Since the exposure was more severe (due to repeated daily dosing) and the NOAEL was higher in the teratology study than it was in the acute neurotoxicity study (250 versus 200 mg/kg), systemic exposure at the NOAEL in the teratology study would be expected to be greater than the acute neurotoxicity study in which DEET peak plasma levels were 16 to 34 times higher than humans at the 95th percentile of human use.
63.5.2.4 Blood Level Studies to Define the Plasma Profile in Rats Administered DEET under the Conditions of the Rat 90-Day Dermal Toxicity Study (LavegIia, 1998b) Because human exposure to DEET is by the dermal route of exposure and is, for the most part, seasonal, the rat and mi-
cropig 90-day dermal toxicity studies appear to be the most appropriate studies for human risk assessment. There also is considerable evidence that rat skin is more permeable than human skin to xenobiotics like DEET, and that the permeability of the skin of the pig is considered to be similar to that of man. Therefore, systemic exposure to DEET in these two animal models would be expected to exceed or mimic systemic human exposure by the dermal route. In these studies, undiluted DEET was administered to rats and micropigs at dose levels of 100, 300, and 1000 mg/kg/day five days per week for a period of 13 weeks. The 1000 mg/kg/day dose level represents the highest dose level that can be applied without runoff and the NOAEL for biologically meaningful effects. In order to compare the systemic exposure to DEET that rats received under this dosage regimen to that of humans under normal use conditions, a blood level study was conducted at a dose level of 1000 mg/kg/day under conditions that simulated the exposure the rats received in the 90-day dermal toxicity study. Ten groups, each consisting of four male and four female rats, were employed in this study. One group of rats was humanely sacrificed for blood collection and subsequent DEET plasma analysis at each of 10 different time intervals over a 24-hr period following a single or following the fifth daily dermal dose. The data from this study are presented in Fig. 63.7 and are summarized in Table 63.6. Unlike the data obtained in the dog and rat blood level studies in which DEET was administered orally as a bolus dose, the DEET plasma profile in the rat dermal blood level study was qualitatively similar to that obtained in the human dermal blood level study. DEET did not accumulate in the plasma and the DEET plasma levels in rats decreased to about baseline levels within 24 hr in a manner similar to that of the human volunteers, even though no attempt was made to remove the material after 8 hr of exposure. The data from the rat dermal blood level study can be compared to the data from the human blood level study in two ways. One is on the basis of peak plasma levels and the other is on the basis of the area under the DEET plasma versus time curve. In this case, the AUC data are considered to be more appropriate because no clinical signs were observed in the laboratory
1456
CHAPTER 63
DEET
animal studies and because plasma profiles were qualitatively similar over the entire 24-hr period. A comparison of the mean AUC data shows 27- and 72-fold differences between humans and male and female rats, respectively.2 63.5.3 RISK ASSESSMENT CONCLUSIONS
Because of the large differences in actual systemic exposure, the studies conducted with DEET by the oral route of administration are not amenable for defining endpoints for toxicity concern or quantitative risk assessment. However, the teratology, reproductive toxicity, neurotoxicity, and oncogenicity studies that were conducted at MTD doses by the more rigorous oral route of administration are useful from the perspective that they demonstrate that DEET is not a teratogen, reproductive toxin, selective neurotoxin, or oncogen. In the studies that are most appropriate for use in human risk assessment, i.e., 90-day dermal toxicity studies in rats and micropigs, no significant endpoints for toxicity concern were observed at the highest dose that could be applied in these animal model systems, i.e., 1000 mg/kg/day. The blood level studies that examined actual systemic exposure to DEET under key exposure scenarios support these conclusions and provide useful data for quantitative risk assessment.
63.6 CONCLUSIONS Over the past 50 years, DEET has been proven to be a safe and effective personal insect repellent that provides public health benefits. In addition to its safe history of human use, it is one of the most thoroughly studied chemicals known to man and a state-of-the-art toxicology database recently has been developed. These studies show that, by the most relevant route of human exposure, DEET is practically nontoxic in laboratory animals. Because systemic toxicity cannot be produced in laboratory animals by the dermal route of exposure, most of the key studies that are traditionally used to define endpoints of toxicity concern were conducted by the more rigorous oral route of exposure in order to satisfy the maximum tolerated dose criteria. Under these conditions, it has been shown that DEET is not a teratogen, reproductive toxin, selective neurotoxin, or oncogen. In addition, while clinical signs of toxicity can be produced by administering DEET as an oral bolus dose, blood level studies have demonstrated that the DEET plasma profile following oral bolus dosing is much different than that observed following dermal administration. Therefore, it is unlikely that the findings observed under these conditions have any relevance to human health under the conditions that DEET is intended to be used. Relative to the millions of people who use DEET each year, there have been relatively few reported cases of serious illness 2 In the human blood level study, no differences were noted between male and female volunteers or between single and repeated exposure scenarios, and the overall mean AUC value was 3.51 ~ghr/m!. For male rats, the corresponding value is 95.25 ~g hr/m!. For female rats, the corresponding value is 251.5 ~ghr/ml.
associated with its use. However, there are approximately 14 case reports involving 20 individuals in the open literature in which a purported relationship between DEET exposure and serious effects have been reported. A review of these cases provided alternative etiologies for the symptoms reported in most patients (Osimitz and Murphy, 1997). Therefore, consistent with the recent opinion of the U.S. EPA (U.S. EPA, 1998), there are no endpoints for toxicity concern for DEET and its use as an insect repellent does not pose a significant health risk to the general population.
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DEET
Moore, G. E. (2000). Acute oral, dennal and inhalation toxicity, primary skin and eye irritation and dennal sensitization studies with DEET insect repellent. Unpublished studies conducted at Product Safety Labs under the sponsorship of Morfiex, Inc. Mount, M. E., Moller, G., Cook, J., Holstege, D. H., Richardson, E. R, and Ardans, A. (1991). Clinical illness associated with a commercial tick and flea product in dogs and cats. Vet. Hum. Toxicol. 33(1), 19-27. Ohayon, A., Ehler, L., and Denver, K. (1997). A blood level study in humans following topical application of N,N-diethyl-m-toluamide (DEET). Unpublished study conducted at L.A.B. Pharmacological Research International, Inc. under the sponsorship of the DEET Joint Venture/Chemical Specialties Manufacturers Association. Olson, M. J., Garg, B. D., Murty, C. V. R, and Roy, A. K. (1987). Accumulation of a2u -globulin in the renal proximal tubules of male rats exposed to unleaded gasoline. Toxicol. Appl. Pharmacol. 90,43-51. Oransky, S., Roseman, B., Fish, D., Gentile, M. S., Melius, J., and Cartter, M. L. (1989). Seizures temporally associated with use of DEET insect repellent. M.M. WR. (New York and Connecticut) 38, 378-680. Oransky, S. (1991). Letter on file with the Department of Environment, Bureau of Pesticides Management, Albany, NY. Osimitz, T. G., Gill, M. w., Gabriel, K. L., Ouellette, R. E., and Schoenig, G. P. (1997). DEET blood level studies in humans following dennal application and in dogs following oral administration. Toxicologist 36,343. Osimitz, T. G., and Murphy, J. V. (1997). Neurological effects associated with use of the insect repellent N ,N -diethyl-m-toluamide (DEET). Clin. Toxicol. 35(5), 435-441. Putman, D. L., and Morris, M. J. (1989). Chromosome aberrations in Chinese hamster ovary (CHO) cells. Unpublished study conducted at Microbiological Associates, Inc. under the sponsorship of the DEET Joint Venture/Chemical Specialties Manufacturers Association. Reifenrath, W. G., Hill, J. A., Robinson, P. B., McVey, D. L., and Akers, W. A. (1980). Percutaneous absorption of carbon 14 labeled insect repellents in hairless dogs. J. Environ. Pathol. Toxicol. 4, 249-256. Reifenrath, W. G., Robinson, P. B., Bolton, V. D., and Aliff, R. E. (1981). Percutaneous penetration of mosquito repellents in the hairless dog: Effect of dose on percentage penetration. Food Cosmet. Toxicol. 19, 195-199. Rej, R., Loux, M., and Copeland, W. (1990). Effect of diethyl-toluamide on human ornithine carbamoyltransferase. Clin. Chem.36, 1143. Roland, E. H., Jan, C. E., and Rigg, M. J. (1985). Toxic encephalopathy in a child after brief exposure to insect repellents. Can. Med. Assoc. J. 132, 155-156. Rowe, P. C., Newman, S. L., and Brusilow, S. W. (1986). Natural history of symptomatic partial ornithine transcarbamylase deficiency. N. Engl. J. Med. 314,541-547. Roy, A. K. (1973). Androgen-dependent synthesis of a2u-globulin in the rat: Role of the pituitary gland. J. Endocrinol. 56, 295-301. San, R. H. C., and Schadly, M. B. (1989). Salmonella/mammalian-microsome plate incorporation mutagenicity assay (Ames test) with a confinnatory assay. Unpublished study conducted at Microbiological Associates, Inc. under the sponsorship of the DEET Joint Venture/Chemical Specialties Manufacturers Association. Schaefer, c., and Peters, P. W. (1992). Intrauterine diethyltoluarnide exposure and fetal outcome. Reprod. Toxicol. 6, 175-176. Schardein, J. L. (1989). Evaluation of DEET in a two generation reproduction/fertility study in rats. Unpublished study conducted at International Research and Development Corporation under the sponsorship of the DEET Joint Venture/Chemical Manufacturers Association. Schmidt, C. H., Acree, E, Jr., and Bowman, M. C. (1959). Fate of C 14 _ diethyltoluarnide applied to guinea pigs. J. Econ. Entomol. 52(5), 928-930. Schoenig, G. P., Hartnagel, R E., Jr., Schardein, J. L., and Vorhees, C. V. (1993). Neurotoxicity evaluation of N,N-diethyl-m-toluarnide (DEET) in rats. Fund. Appl. Toxicol. 21,355-365. Schoenig, G. P., Neeper-Bradley, T. L., Fisher, L. C., and Hartnagel, RE., Jr. (1994). Teratologic evaluations of N,N-diethyl-m-toluamide (DEET) in rats and rabbits. Fund. Appl. Toxicol. 23,63-69.
Schoenig, G. P., Hartnagel, R E., Jr., Osimitz, T. G., and Llanso, S. (1996). Absorption, distribution, metabolism, and excretion of N,N-diethyl-mtoluamide in the rat. Drug Metab. Disposition 24(2), 156-163. Schoenig, G. P., Osimitz, T. G., Gabriel, K. L., Hartnagel, R, Gill, M. w., and Goldenthal, E. I. (1999). Evaluation of the chronic toxicity and oncogenicity of N,N-diethyl-m-toluamide (DEEt). Toxicol. Sci. 47, 99-109. Selim, S., Hartnagel, R. E., Jr., Osimitz, T. G., Gabriel, K. L., and Schoenig, G. P. (1995). Absorption, metabolism, and excretion of N,N-diethyl-mtoluamide following dennal application to human volunteers. Fund. Appl. Toxicol. 25, 95-100. Shennan, R A. (1980). Phase 2-Behavioral effects of acute aerosol exposure to N, N -diethyl-meta-toluarnide (M-Det). Unpublished study conducted at U.S. Anny Environmental Hygiene Agency. Sippel, A. E., Fcigclson, P., and Roy, A. K. (1975). Honnonal regulation of the hepatic messenger RNA levels for a2u -globulin. Biochemistry 14, 825-829. Snodgrass, H. L., Nelson, D. c., and Weeks, M. H. (1982). Dennal penetration and potential for placental transfer of the insect repellent N, N -diethyl-mtoluamide. Am. Ind. Hyg. Assoc. J. 43(10),747-754. Snyder, J. w., Poe, R. 0., Stubbins, J. E, and Garrettson, L. K. (1986). Acute manic psychosis following the dennal application of N-N-diethylm-toluamide. J. Toxicol. Clin. Toxicol. 24,429-439. Stenback, E (1976). Testing of cosmetics, ingredients of sunscreen ointments, insect repellents, and detergents on skin of mice and rabbits: Lifespan studies. Unpublished study conducted at the Eppley Institute for Research in Cancer and Allied Diseases under the sponsorship of Morflex Chemical Co., Inc. Sterner, W. (1977). Effect of "MGK Diethyltoluamide" on the embryonic development of rats after oral application. Unpublished study conducted at International Bio-Research, Inc. under the sponsorship of McLaughlin Gonnley King Co. Swentzel, K. C. (1978). Investigation of N,N-diethyl-m-toluamide (M-Del) for dominant lethal effects in the mouse study. Unpublished study conducted at U.S. Anny Environmental Hygiene Agency. Taylor, W. G. (1986). Metabolism of N,N-diethyl-meta-toluarnide by rat liver microsomes. Drug Metab. Disp. 14(5),532-539. Taylor, W. G., Danielson, T. J., Spooner, R w., and Golsteyn, L. R. (1994). Pharmacokinetic assessment of the dennal absorption of N,N-diethyl-mtoluamide (DEET). Drug Metab. Disp. 22(1), 106-112. Tenenbein, M. (1987). Severe toxic reactions and death following the ingestion of diethyltoluamide-containing insect repellents. J. Am. Med. Assoc. 258(11),1509-1511. U.S. EPA (1980). Pesticide Registration Standard-DEET. Registration Division and Special Pesticide Review Division. U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1984). Pesticide Assessment Guidelines, Subdivision F, Hazard Evaluation: Human and Domestic Animals (Revised). Hazard Evaluation Division, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1998). Reregistration Eligibility Document on DEET. Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. Veltri, J. C., Osimitz, T. G., Bradford, D. c., and Page, B. C. (1994). Retrospective analysis of calls to poison control centers resulting from exposure to the insect repellent N,N-diethyl-m-toluamide (DEET) from 1985-1989. J. Toxicol. Clin. Toxicol. 32,1-16. Verschoyle, R. D., Brown, A. w., Nolan, c., Ray, D. E., and Lister, T. (1992). A comparison of the acute toxicity, neuropathology, and electrophysiology of N,N -diethyl-m-toluamide and N,N -dimethyl-2,2-diphenylacetamide in rats. Fund. Appl. Toxicol. 18, 79-88. Weil, C. S. (1973). N,N-diethyl-m-toluamide range finding toxicity studies. Unpublished studies conducted at Carnegie-Mellon University under the sponsorship of Union Carbide Corporation. Whitley, R. J. (1994). Viral encephalitis. Diagnosis and treatment. c.N.S. Drugs 2,355-366. Windheuser, J. J., Haslam, J. L., Cadwell, L., and Shaffer, R D. (1982). The use of N,N -diethyl-m-toluamide to enhance dennal and transdennal delivery of drugs. J. Pharmaceut. Sci. 71(11), 1211-1213.
References Wright, D. M., Hardin, B. D., Goad, P. w., and Chrislip, D. W. (1992). Reproductive and developmental toxicity of N,N-diethyl-m-toluamide in rats. Fund. Appl. Toxieo!. 19,33-42. Wu, A., Pearson, M. L., Shekoski, D. L., Soto, R. 1., and Stewart, R. D. (1979). High resolution gas chromatography/mass spectrometric characterization of urinary metabolites of N,N-diethyl-m-toluamide (DEET) in man. 1. High Resolut. Chromatogr. Commun. 2,558-562. Yeung, 1. M., and Taylor, W. G. (1988). Metabolism of N,N-diethyl-mtoluamide (DEET) by liver microsomes from male and female rats: Simulta-
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neous quantitation of DEET and its metabolites by high performance liquid chromatography. Drug Metab. Disp. 16(4),600--604. Zadikoff, C. M. (1979). Toxic encephalopathy associated with use of insect repellent. 1. Pediatr. 95, 140-142. Zeiger, E., Anderson, B., Haworth, S., Lawlor, T., and Mortelmans, K. (1992). Salmonella mutagenicity tests: V. Results from the testing of 311 chemicals. Environ. Mol. Mutagen. 19(5uppl. 21), 2-141.
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CHAPTER
64 The Safety Assessment of Piperonyl Butoxide Thomas G. Osimitz s.c. lohnson and Sons, Inc. Ruaidhri Breathnach University College Dublin
64.1 CHEMISTRY AND FORMULATIONS
able, although their use is minor relative to that of PBO and pyrethrins/pyrethroids.
Piperonyl butoxide (PBO), 2-(2-butoxyethoxy)ethyl 6-propylpiperonyl ether (IUPAC), is an insecticide synergist produced from the condensation of the sodium salt of 2-(2-butoxyethoxy) ethanol and the chloromethyl derivative of hydrogenated safrole (Fig. 64.1). The methylenedioxyphenylmoiety constitutes over half of the PBO molecule by weight and is derived from sassafras oil. Sassafras oil is an essential oil distilled from several species of trees found in Brazil, China, and Viet Nam. Current world production of PBO averages 92% pure. In the early days, PBO contained small but detectable amounts of safrole and dihydrosafrole (DHS). However, refinements in distillation have resulted in safrole and DHS levels usually below the 40 ppm detection limit by high resolution gas chromatography (Di Blashi, 1998). The development of PBO grew out of a need in the late 1930s and early 1940s to extend the usefulness of the naturally derived insecticide pyrethrum, which was considered a strategic insecticide against mosquitoes and other disease-carrying insects. The chemicals that were developed had little intrinsic pesticidal activity of their own; however, they did increase the effectiveness of a given dose of pyrethrins and were thus called synergists. PBO was one of a series of molecules synthesized (Wachs, 1947). PBO is usually formulated with natural pyrethrins or synthetic pyrethroids in ratios (PBO: pyrethrins) ranging from 5:1 to 20: 1. Formulations of PBO and carbamates are also avail-
Figure 64.1
Chemical structure of piperonyl butoxide (PBO).
Handbook of Pesticide Toxicology Volume 2. Agents
64.2 USES As a synergist, PBO inhibits the mixed function oxidase (MFO) system of insects, thereby reducing the oxidative breakdown of other pesticides like pyrethrum and the synthetic pyrethroids (Casida, 1970). The precise mechanism of inhibition is unknown, but speculation is that a carbene derivative forms and binds to the heme moiety of the cytochrome P-450 enzyme, thereby rendering it inactive (Dahl and Brenzinski, 1985; Delaforge et aI., 1985; Franklin, 1976; Hodgson et aI., 1973; Murray and Reidy, 1989; Philpot and Hodgson, 1971, 1972a, 1972b). The result is that higher levels of the insecticide remain in the insect and are thereby available to exercise their lethal effect on the insect. PBO enhances the pesticidal activity of a given level of active ingredient, thus promoting reduced use of the pesticide. Appearing in several thousand U.S. Environmental Protection Agency (U.S. EPA)-registered products, PBO is one of the most commonly registered pesticides in terms of the number of formulas in which it is present. It is approved for preharvest application to a wide variety of crops including fruits and vegetables. The application rates are low; the highest single rate is 0.386 lbs PBO/acre. It is also used extensively in combination with pyrethrins and some synthetic pyrethroids to control insect pests in and around the home and in food handling establishments. A wide variety of water-based PBO-containing products such as crack and crevice sprays, total release foggers, and flying insect sprays are made for use by consumers in the home. Annual use of PBO in the United States is approximately 1.3-1.5 million pounds (PBOTF, 1997). About 47% of this total
1461
Copyright © 200 1 by Academic Press. All rights of reproduction in any form reserved.
1462
CHAPTER 64
PBO Safety
is used for indoor residential purposes, 17% for indoor food uses in warehouses and food handling establishments, and only 13% goes for agricultural crop applications. Piperonyl butoxide has also been allowed as a food additive in Japan since 1955, its maximum approved level being 0.024 g/kg (24 ppm) in raw cereals.
64.3 HAZARD IDENTIFICATION 64.3.1 ACUTE TOXICITY Numerous acute studies have been conducted over the years with PBO in a variety of species and by various exposure routes. This body of data, including the most recent studies, indicates that PBO is generally of low acute toxicity to animals. It is mildly irritating to the eye and skin. It is not a dermal sensitizer. Table 64.1 summarizes this acute toxicity data as well as European Economic Commission (EEC) labeling classifications and U.S. EPA toxicity categories. 64.3.2 SUBCHRONIC TOXICITY PBO has been tested in dogs, mice, rats, rabbits, and African green monkeys for subchronic toxicity. A summary of the subchronic toxicity studies discussed below is presented in Table 64.2. 64.3.2.1 Dogs Lorber (1972) reported unexpected alterations in blood counts of intact and splenectomized dogs after use of a fogger (containing PBO among other chemicals). The fogger bathed the animals inadvertently in a "dense pesticide mist." The dogs were part of a research project investigating the relationship of spleen, bone marrow, and blood cells. Further work was undertaken in which 17 intact, 12 splenectomized, and 7 partially
splenectomized dogs were purposefully exposed to deodorized kerosene containing only PBO (1.5%). Exposure periods consisted of four intervals of 5 minutes duration, with an 8-minute interval between each exposure. The splenectomized dogs showed a reduction in serum platelet count and an occasional increase in reticulocytes. The authors concluded that the demonstrated greater resistance of intact dogs to the hematotoxic potential of the tested chemicals may have been in part due to the larger spleens in these animals which could perhaps sequester the chemical(s) more effectively. Moreover, they felt that the normal hepatic blood flow in these animals could also enhance the removal of the chemical(s) from the systemic circulation. Goldenthal (1993a) conducted a range-finding study as a prelude to a I-year chronic study. PBO was administered in the diet to dogs (4 animals/sex/dose level) for 8 weeks. The dosage levels were 500, 1000, 2000, and 3000 ppm (approximately equivalent to 12.5, 25, 50, and 75 mg/kg body weight/day, respectively). All dogs survived to study termination. All animals in the 3000 ppm dose group had decreased appetites and reduced defecation during the first week. No other abnormal clinical signs were present. Three out of four dogs lost weight in the highest dose group. Even at 1000 ppm PBO in the diet, weight gains were lower than the control group. Food intake was similar between treated and control groups, except for a slight reduction in some animals at the 3000 ppm dosage level. There were no treatment-related effects on hematological parameters at any dose level, but slight increases in alkaline phosphatase values and slight decreases in cholesterol were noted at doses :::2000 ppm. No treatment-related microscopic changes were noted at necropsy in any group, but a compoundrelated increase in absolute and relative liver and gall bladder weights was recorded in males. Upon histopathologic examination, hypertrophy of hepatocytes was noted in males of all dose levels and in females at dosages 2000 ppm and above. This finding was consistent with the increases in liver weights and serum alkaline phosphatase levels described above. No other
Table 64.1 Summary of Acute Toxicity Data and Classifications EEC labeling
U.S. EPA toxicity
Route
Species
Result
classification
category
Reference
Oral LDSO
Rat
>4 g/kg (male)
Unclassified
Category IV
Gabriel (199Ia)
> 7 g/kg (female) Dermal LDSO
Rabbit
>2g/kg
Unclassified
Category IV
Gabriel (199Ib)
Inhalation LCSO
Rat
>5.9 mglL air
Harmful
Category III
Hoffman (1991)
Eye irritation
Rabbit
Minimally
Labeling not
Category III
Romanelli (1991a)
irritating
indicated Category IV
Romanelli (1991b)
Category IV
Romanelli (199Ic)
Skin irritation
Rabbit
Skin sensitization
Guinea
(Buehler)
pig
Minimally
Labeling not
irritating
indicated
Negative
Labeling not indicated
64.3 Hazard Identification
treatment-related microscopic changes were evident. There was a decrease in the absolute and relative weights of the testes and epididymis in the groups treated with 2000 and 3000 ppm. The dose level of 500 ppm was set as a NOAEL for this study because the changes recorded in the liver were considered adaptive in nature rather than adverse and were not accompanied by any systemic signs of toxicity.
64.3.2.2 Mice Fujitani et at. (1993) reported the results of dosing CD-l mice (10 animals/sex/dose level) with 1000, 3000, or 9000 ppm
1463
(approximately equivalent to 150, 450, and 1350 mg/kg body weight/day, respectively) PBO in the diet for 20 days. Body weights were depressed in the high-dose animals (about 15%) and in the mid-dose females (about 8%). Kidney and spleen weights were also reduced in the high-dose group. A treatmentrelated elevation in liver weights was noted with a 79% increase in the high-dose males. Females in the high-dose group showed higher levels of y-glutamyl transpeptidase (GGT) activity. The high-dose males and females featured higher levels of cholesterol, phospholipids, and total serum proteins. Hepatocyte hypertrophy, single cell necrosis, and inflammation, most
Table 64.2 Summary of Results of Subchronic Toxicity Studies with PBO Species
Route
Dose
Duration
NOAEL
Comments
Reference
Dog
Inhalation
15,000ppm
Four 5 min intervals
Not applicable
Reduction in serum platelet count, increase in reticulocytes (splenectomized animals)
Lorber (1972)
Diet
500-3000 ppm (~12.5-75 mg/kg body weight/day)
8wks
500ppm (~ 12.5 mglkg body weight/day)
Decreased body weight, increased liver weight, hepatocyte hypertrophy
Goldenthal (1993a)
Diet
1000-9000 ppm (~ I 50-1350 mg/kg body weight/day)
20 days
1000 ppm (~ 150 mg/kg body weight/day)
Increased liver weight, hepatocyte hypertrophy, necrosis, inflammatory cell infiltration
Fujitani et al. (1993)
Oral
10-1000 mg/kg body weight/day
90 days
30 mg/kg body weight/day
Increased liver weight, liver necrosis, centrilobular hypertrophy
Chun and Wagner (1993)
Oral
1500--6000 ppm (236-880 mg/kg body weight/day)
7 wks
Not established
Alterations in motor activity
Tanaka (1993)
Gavage
2.5-5 ml/kg body weight/day
31 days
Not established
Anorexia, loss of weight, death
Sarles and Vandergrift (1952)
Oral
1857 mglkg body weight/day
90 days
Not applicable
40% mortality, increased liver weight
Bond et al. (1973)
Diet
62.5-2000 mg/kg body weight/day
28 days
125 mg/kg body weight/day
Increased liver weight, microscopic changes in liver
Modeweg-Hausen et al. (1984)
Oral
6000-24,000 ppm (~300-1200 mg/kg body weight/day)
13 wks
Not established
Decreased body weight, increased liver and kidney weights, hepatocyte hypertrophy
Fujitani et al. (1992)
Gavage
250-4000 mg/kg body weight/day 15-512 mg/m 3
10 days
Ataxia, twitching, dyspnoea, gastric ulceration
Chun and Neeper-Bradley (1992)
90 days
250 mg/kg body weight/day 155 mg/m 3
Alterations in clinical chemistry parameters, increased liver and kidney weights
Newton (1992)
Mouse
Rat
Inhalation
Rabbit
Monkey
Oral
I or 4 ml/kg body weight/wk (5% emulsion)
3 wks
Not applicable
No signs of toxicity
Sarles et al. (1949)
Dermal
100-1000 mg/kg body weight/day
3 wks
1000 mg/kg body weight/day (systemic toxicity) 100 mg/kg body weight/day (local effects)
Slight erythemaledema, fissuring/inflammation of skin
Goldenthal (1992)
Oral
0.03 or 0.1 ml/kg body weight/day
4wks
Not applicable
Minor changes in liver
Sarles and Vandergrift (1952)
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CHAPTER 64 PBO Safety
prominently in the centrilobular region, were also seen in the livers of the mid- and high-dose groups. The NOAEL for this study, based on liver toxicity, was 1000 ppm. As a prelude to a 2-year bioassay, Chun and Wagner (1993) report the conduct of a 90-day oral toxicity study in CD-l mice (15 animals/sex/dose level) using dose levels of 10, 30, 100, 300, and 1000 mg PBO/kg body weight/day in the diet. A significant decrease in body weight was noted in the high-dose males (-34% versus controls). The target organ for toxicity was the liver as indicated by increased liver weights, hepatocyte necrosis, and centrilobular hypertrophy (NOAEL = 30 mg/kg body weight/day). Tanaka (1993) reported the results of a study in male CD-l mice designed to probe neurobehavioral endpoints. Male mice were fed PBO daily from 5-12 weeks of age at concentrations of 1500, 3000, or 6000 ppm (equivalent to 236, 448, and 880 mg/kg body weight/day, respectively as averaged over the 5-11 week period). Twenty animals were used per dose level. No significant adverse effects were observed in water T-maze performance. Significant changes were noted, however, in exploratory behavior at 8 and 11 weeks of age. The incidence of defecation was increased at all dose levels at 8 weeks. Numerous parameters were affected at 11 weeks [i.e., number of movements, movement time, number of horizontal activities, total distance (all dose levels), number of turnings, average distance (all dose levels) and average speed]. The authors concluded that piperonyl butoxide has an influence on the motor activity of exploratory behavior in mice at 11 weeks of age.
64.3.2.3 Rats Sarles and Vandergrift (1952) gavaged six male and six female rats with PBO daily, 6 days/week for 31 days. The first seven doses were at 2.5 ml/kg body weight. Two animals died on the third and fourth days. The remaining animals improved after some initial clinical signs of toxicity. They were dosed with a second seven doses of 3.5 mL/kg body weight each. Little toxicity was noted at this dose level. Hence, the animals thereafter received doses of 5 ml/kg body weight. Clinical signs of toxicity included anorexia and loss of weight. Additional animal deaths occurred from the 17th to 24th days of testing; the next and last death was at 31 days. Bond et al. (1973) administered 1857 mg PBO/kg body weight/day orally to a single group of 20 rats for 90 days. Forty percent of the rats died prior to conclusion of the study. The most significant finding was a dramatic increase in liver weight (i.e., 2.4 times that of untreated controls). The authors also allude to another study they performed, which is unpublished, in which rats were fed 500 mg PBO/kg body weight/day. These animals were reported to have liver and kidney damage. A 4-week range-finding study was conducted by ModewegHausen et al. (1984). Rats (10 animals/sex/dose level) were fed 62.5, 125, 250, 500, 1000, or 2000 mg PBO/kg body weight/day. Hepatic eosinophilic infiltration and the increased vacuolization of hepatocytes were seen with increasing severity among the mid- and high-dose groups. These effects were
viewed as being degenerative changes representing chronic toxicity. Liver weights were elevated at 250 mg/kg body weight/day and above in the males and at 500 mg/kg body weight/day and above in the females. Except for an increase in alkaline phosphatase at the highest dose level, no treatmentrelated changes were reported in hematologic and clinical chemical parameters. Based on the observed liver toxicity, the NOAEL for this study is 125 mg/kg body weight/day. A 13-week subacute oral toxicity study was performed in Fischer F344 rats (10 animals/sex/dose level) at levels of 6000, 12,000, or 24,000 ppm (approximately equivalent to 300, 600, or 1200 mg/kg body weight/day, respectively) PBO in the diet (Fujitani et al., 1992). No mortality occurred. Nasal bleeding and dose-related abdominal distension were reported. A significant decrease in body weight was evident in the high-dose groups (36% decrease in males, 24% decrease in females). Blood hemoglobin levels were reduced in both sexes in the high-dose group and in mid-dose females. Biochemical changes in the high-dose group consisted of increases in albumin, cholesterol, urea, and GGT activity. Liver and kidney weights were increased in a dose-dependent manner. Histopathologic examination revealed hypertrophic hepatocytes (containing a basophilic granular substance) and vacuolation of hepatocytes in periportal areas. Coagulative necrosis and oval cell proliferation were occasionally seen. Atrophy of the epithelial lining of the proximal convoluted tubules in the renal cortex was present in some male rats. A NOAEL for PBO could not be established in this study owing to the presence of liver and kidney effects even at the "low" dose of 6000 ppm. Marked clinical signs of subacute toxicity were seen in the dams of a range-finding study conducted to select doses for a developmental toxicology study (Chun and Neeper-Bradley, 1992). Pregnant female rats (15 animals/dose level) were gavaged on gestational days 6-15 with PBO at levels of 250, 500, 1000, 2000, or 4000 mg/kg body weight/day. Signs of general stress, such as urogenital wetness and periocular encrustation, were evident in many animals during the first three days of the study at dose levels of at least 500 mg/kg body weight/day. At levels of 2000 mg/kg body weight/day, more severe clinical signs such as ataxia, twitching, prostration, dyspnea, gasping, and lacrimation were noted. Ulceration of the lining of the glandular region of the stomach as well as hemorrhage and sloughing of the lining of the nonglandular region were noted at necropsy. In the only subchronic study conducted by the inhalation route, Newton (1992) exposed CD rats (15 animals/sex/dose level) for 6 hr/day,S days/week, for 90 days in whole body exposure chambers. PBO was aerosolized to achieve exposure concentrations of 15, 74, 155, and 512 mg PBO/m3 (MMAD of the aerosol was 1.7 J.!m). Neither body weight gain nor food intake was affected by exposure. In the high-dose group, serum alanine transaminase, aspartate transaminase, and glucose levels were decreased, whereas BUN, total protein, and albumin levels were increased. However, not all of these effects were statistically significant, and there was no clear dose-response re-
64.3 Hazard Identification
lationship. Both absolute and relative liver and kidney weights were elevated in the high-dose group. Minimal to slight irritation of the larynx was observed upon necropsy in all treatment groups. Inflammation, congestion, edema, and debris in the lumen were noted as well. Squamous metaplasia of the laryngeal epithelium was noted in all groups but was most marked in both sexes at the highest dose level. These changes are considered to represent an adaptive response to mild irritation and do not represent systemic toxicity. Moreover, there is no evidence that such metaplasia in the absence of atypia is preneoplastic (Brown, 1990; Monticello et aI., 1990).
64.3.2.4 Rabbits Sarles et al. (1949) performed a subacute oral toxicity experiment in rabbits. A 5% PBO emulsion was fed once weekly in the diet to three rabbits over a 3-week period. The dosage used varied between 1.0 and 4.0 ml/kg body weight/week. There was neither mortality nor clinical signs of toxicity. The rabbit that received the highest dosage was sacrificed 1 week after the last treatment, but no lesions were detected at postmortem examination. A 21-day subchronic dermal toxicity study was conducted in rabbits (5 animals/sex/dose level) in which 100, 300, or 1000 mg PBO/kg body weight was applied topically once a day, 5 days a week, for 3 consecutive weeks (Goldenthal, 1992). Treatment-related effects were limited to minor skin changes at the application site. Dermal irritation was present in all treatment groups (although to a lesser extent and incidence at the 100 mg/kg body weight/day dose level). Dermal lesions consisted of very slight erythema and edema. This irritation usually appeared by day 5 and persisted for the remainder of the study. Desquamation and fissuring of the skin appeared in the 300 and 1000 mg/kg body weight/day groups. Moderate acanthosis, hyperkeratosis, and chronic inflammation of the epidermis were present. The severity of these lesions increased with increasing dosage. Body weights were comparable with those in the control group, and food intake was only slightly lower in treated animals. No treatment-related changes were seen in hematology and clinical chemistry, and no signs of systemic toxicity were present at any dosage level. The NOAEL is 100 mg/kg body weight/day for local effects whereas the NOAEL for systemic toxicity is 1000 mg/kg body weight/day.
64.3.2.5 Other Species A 4-week oral toxicity study was performed by Sarles and Vandergrift (1952) in which two African Green monkeys were fed PBO by capsule, 6 days a week (for 4 weeks), at a dosage level of 0.03 or 0.1 ml/kg body weight/day (one monkey at each dosage level). No gross pathological lesions were evident in the treated monkeys' livers. Upon histopathologic examination of the liver, the monkey on the higher dose level showed evidence of minimal dystrophy and dysplasia, occasional acidophilic and hyaline-necrosis cells, as well as hydropic swelling.
1465
64.3.3 REPRODUCTIVE AND DEVELOPMENTAL TOXICITY A summary of the results of the reproductive and developmental toxicity studies discussed below is presented in Tables 64.3 and 64.4.
64.3.3.1 Reproductive Toxicity Mice A three-generation, one litter per generation, reproductive toxicity study was performed in CD-1 mice by Tanaka et al. (1992). Ten animals of either sex were incorporated at each dosage level and PBO dosage rates were set at 1000, 2000, 4000, and 8000 ppm (purity not specified) in the diet. These dose levels are equivalent to 268, 506, 936, and 1583 mg PBO/kg body weight/day as averaged over Fo and Fl generations from preconception through lactation. Food intake was reduced in the Fo generation at the 8000 ppm dose, except during the mating period, and was also reduced during the lactation period in the Fl generation, also at 8000 ppm. The 4000 ppm treatment groups of both the Fo and the F 1 generations also had a reduction in food intake during the lactation period. Mean F I litter weight was significantly decreased (38%) at 8000 ppm and reduced by 18% at the 4000 ppm treatment level. However, litter size remained unchanged at all levels. Pups born in the 8000 ppm treated group of the Fl generation had a lower survival index at postpartum day 21 (63% versus 91 % for males of control group; 79% versus 89% for females of control group). Pup weights in the Fl generation were decreased for all dosage groups, but there was no dose-related response at lower or mid-dose levels. N0 treatment-related effects were noted in neurobehavioral tests conducted in the F 1 animals. The mean F2 litter size was significantly decreased at the 4000 and 8000 ppm treatment levels. Mean F2 litter weights were decreased in all treated groups. Pups in the 8000 ppm treated group of the F2 generation had a lower survival index than controls at day 21 postpartum (59% in males and 79% in females versus 100% in male and female control groups). Pup weights in the F2 generation were decreased at dosage levels of 2000 ppm and above on postnatal days 4, 7, 14, and 21. Pup weights were reduced on postnatal days 4 and 7 at the 1000 ppm dosage level. No clear dose-response relationship was evident in the neurobehavioral tests conducted with F2 animals. Because pup weights were reduced at all doses tested, a NOAEL could not be set for this study. Tanaka (1992) further probed the developmental neurotoxicity of PBO in CD-l mice in a subsequent study. Ten mice of either sex per group received diets containing 1500, 3000, or 6000 ppm (approximately equivalent to 400, 700, and 1250 mg/kg body weight/day, respectively, based on food consumption values in Tanaka et aI., 1992) PBO during a 4-week period prior to mating (Fo), during gestation and through the time that the F 1 generation was 8 weeks old. The open field test demonstrated a dose-dependent decrease in ambulating and rearing in Fo male mice. However, because of the excessive dosage levels incorporated in this study, pup body weights were
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PBO Safety
Table 64.3 Summary of Results from Reproductive Studies with PBO Species
Route
Dose
Study type
NOAEL
Comments
Reference
Mice
Diet
1000-8000 ppm (268-1583 mg/kg body weight/day)
Threegeneration
Not established
Pup weights reduced at all dose levels.
Tanaka et al. (1992)
Diet
1500-6000 ppm mg/kg body weight/day)
Twogeneration
Not established
Excessive dose levels resulted in decreased pup weights in all treated animals.
Tanaka (1992)
100-25,000 ppm mg/kg body weight/day)
Threegeneration
1000 ppm mg/kg body weight/day)
Very high maternal toxicity at two highest dose levels resulting in marked reductions in the incidence of pregnancies, numbers of litters per dam, general health of the offspring, and average weanling weights of pups.
Sarles and Vandergrift (1952)
300-5000 ppm (~24-400 mg/kg body weight/day)
Twogeneration
Parental toxicity/ pup development: 1000ppm
Body weights of pups born to dams treated at the highest dose level were reduced in the early postpartum period.
Robinson et al. (1986)
(~400-1250
Rat
Diet
(~8-2000
Diet
(~8
(~80mg/kg
body weight/day) Reproductive toxicity: 5000 ppm (~400 mglkg body weight/day)
Dog
Diet
30-500 mg/kg body weight/day
2-yr chronic
Not applicable
Increases in ovarian weight observed in some females at highest dose level.
Butler et al. (1998)
Diet
500-3000 ppm (~12.5-75 mg/kg body weight/day)
Rangefinding
Not applicable
Increased absolute and relative weights of testis and epididymis noted. No microscopic abnormalities observed in the testis.
Goldenthal (1993a)
reduced at birth in all treated animals. By postpartum day 21, the mean pup body weight in the mid-dose group was 7% lower than controls. The mean body weight of high-dose pups was decreased 41 %. The survival index for pups at postpartum day 21 was 79.2% (controls), 92.9% (low-dose group), 80.0% (mid-dose group), and 51.7% (high-dose group). There were no significant differences in the behavioral test during the lactation period, except for a reduction in olfactory orientation in mid- and high-dose group animals. Other than sporadic nondose-dependent changes, the open field test and multiple water T-maze tests were not significantly altered by PBO.
Rats In a reproductive toxicity study reported by Sarles and Vandergrift (1952), groups of 12 male and 12 female rats per dose level were fed diets containing 100, 1000, 10,000, or 25,000 ppm (approximately equivalent to 8, 80, 800, or 2000 mg/kg body weight/day, respectively) of PBO (technical grade, 80% purity), for three generations. None of the female rats at the highest dose level were fertile and there were marked reductions in the incidence of pregnancies, numbers of litters per dam, general health of the offspring, and average weanling
weights of pups born to dams treated at 10,000 ppm. These findings are clearly a result of the high maternal toxicity, especially at 10,000 and 25,000 ppm. No adverse effect on reproduction was observed in three generations of progeny in the 100 and 1000 ppm groups (NOAEL = 1000 ppm). A two-generation reproduction study was performed in rats with PBO by Robinson et al. (1986). Groups of 26 male and 26 female Sprague-Dawley rats were utilized and adults of the Fo and F 1 generations were treated at dose levels of 300, 1000, or 5000 ppm (approximately equivalent to 24, 80, or 400 mg/kg body weight/day, respectively) in the diet. Animals were treated for 83 to 85 days prior to placement for mating, and treatment continued throughout mating, pregnancy, and lactation. The only consistent finding throughout the study period was a lower body weight gain at the highest dosage level. This tendency was partially reversed during the lactation period, when females at this dose level showed higher weight gains when compared with control rats. For both F 1 and F2 generation pups, the viability, survival, and lactation indices were unaffected by treatment. There were no treatment-related abnormal findings for the pups, and weanlings did not reveal any treatment-related
64.3 Hazard Identification
1467
Table 64.4 Summary of Results from Developmental Studies with PBO Species
Route
Dose
NOAEL
Comments
Reference
Mice
Gavage
1065-1800 mglkg body weight/day
Not established
Total resorption rates significantly increased in mid- and high-dose groups. Significant decrease in body weights of male and female fetuses, appearing to be dose-dependent.
Tanaka et al. (1994)
Rat
Gavage
300 or 1000 mg/kg body weight/day
Maternal toxicity: Not established
Maternal body weights were reduced at both dose levels tested.
Kennedy et al. (1977)
Developmental toxicity: 1000 mg/kg body weight/day
Gavage
62.5-500 mg/kg body weight/day
Maternal and developmental toxicity: 500 mg/kg body weight/day
No signs of either maternal or embryofetotoxicity.
Khera et al. (1979)
Gavage
200-1000 mg/kg body weight/day
Maternal toxicity: 200 mg/kg body weight/day
Gestational body weights and body weight gains were reduced in the 500 and 1000 mg/kg body weight/day groups.
Chun and Neeper-Bradley (1991)
Maternal toxicity was evident at 100 and 200 mg/kg body weight/day manifested by decreased defecation and a dosedependent weight loss during the treatment period.
Leng et al. (1986)
Developmental toxicity: 1000 mg/kg body weight/day
Rabbit
Gavage
50-200 mg/kg body weight/day
Maternal toxicity: 50 mg/kg body weight/day
Developmental toxicity: 200 mg/kg body weight/day
adverse effects. Body weights of pups born to dams treated at the highest dose level were reduced in the early postpartum period. The NOAEL for parental toxicity and pup development was thus 1000 ppm PBO in the diet. The NOAEL for reproductive toxicity was set at 5000 ppm PBO in the diet. Other Studies In an 8-week dietary range-finding toxicity study in dogs, PBO was fed daily in the diet to beagles at dose rates of 500, 1000, 2000, or 3000 ppm (approximately equivalent to 12.5, 25, 50, or 75 mg/kg body weight/day, respectively) (Goldenthal, 1993a). There was an increase in the absolute and relative weights of the testes and epididymides. No microscopic abnormalities were noted in the testes. Spermatozoa were being produced. Other details of the study are presented in Section 64.3.2.1. In a 2-year chronic oral toxicity study in Sprague-Dawley rats, animals received dietary administration of 15, 30, 100, or 500 mg PBO/kg body weight/day (89% purity) (Butler et aI., 1998). Only changes in the reproductive system are discussed here. Full details of the study are presented in Section 64.3.4.3. Increases in ovarian weights were observed among some females receiving 500 mg/kg body weight/day, although no histopathologic changes were noted. Atrophy of the testes was seen histologic ally in all male groups and when bilateral atrophy was considered alone, there was an increased incidence in the intermediate and high-dose groups with a corresponding reduction in the incidence of unilateral atrophy. However,
the finding is unlikely to be related to treatment because the dose-response relationship was unclear and the atrophy was not accompanied by changes in the seminiferous tubules or sperm production. Moreover, there were no statistically significant increases or decreases in testes weight when expressed as either absolute weight or relative-to-brain weight. 64.3.3.2 Developmental Toxicity Mice Tanaka et al. (1994) reported a study conducted in CD-l mice in which PBO was administered by gavage on day 9 of gestation to groups of 20 animals at doses of 1065, 1385, or 1800 mg/kg body weight (>95% purity; PBO dissolved in olive oil). No abnormal behavior or mortality patterns were observed in dams. Three abortions occurred in the mid- and high-dose groups. Four litters were resorbed in the two higher dosage groups, but maternal body weights were comparable between all groups. Total re sorption rates were significantly increased in the mid-dose (26%) and the high-dose groups (32%) when compared with the control value (6%). The number of viable fetuses per dam was comparable between all dosage groups. There was a significant decrease in body weights of male and female fetuses derived from treated dams, which did appear to be dose-dependent. Certain external malformations such as exencephaly, craniochisis, open eyelids, omphalocele, kinky tail, and talipes varus were observed in all groups (including controls) and oligodactyly was recorded in the forelimbs of some
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CHAPTER 64 PBO Safety
fetuses derived from treated dams. The incidence of this latter defect was 6% in those fetuses derived from the highest dosage group. The authors concluded that a single high dose of PBO (1065 mg/kg body weight or above), when given orally to pregnant mice on day 9 of gestation, could cause embryofetal toxicity with associated oligodactyly of the forelimbs. The high dose levels of this study make it difficult to interpret the significance of this finding. Rats Kennedy et al. (1977) performed a teratogenicity study with PBO in pregnant rats. Twenty female animals per dose level were gavaged with PBO in corn oil at 300 or 1000 mg PBO/kg body weight/day. Other than a decline in body weight gain in both treated groups (especially in the later stages of gestation), no other treatment-related signs of toxicity occurred. The reproductive parameters of the dams were not significantly affected by treatment. One female from each treatment group resorbed most or all of her litter. The fetuses derived from each treatment group exhibited no internal or external skeletal malformations that could be related to treatment. Because maternal body weights were reduced at both doses tested, a NOAEL was not established for maternal toxicity. Because no developmental effects were seen, the NOAEL for embryo-fetal toxicity for this study was 1000 mg/kg body weight/day. Pregnant female Wistar rats (17-20 per dosage group) were dosed with PBO levels of 62.5, 125, 250, or 500 mg/kg body weight/day from day 6 to day 15 of gestation in a study performed by Khera et al. (1979). The types and incidences of anomalies in fetuses derived from treated dams were comparable with those of the control group and it was concluded that doses as high as 500 mg/kg body weight/day produced no signs of either maternal or embryo-fetal toxicity. A developmental toxicity study with PBO was performed in Sprague-Dawley rats by Chun and Neeper-Bradley (1991). Timed pregnant rats were administered PBO (90.78% purity) by gavage on gestation days 6 to 15. The dosage levels were 200, 500, and 1000 mg/kg body weight/day and 25 animals were included in each group. The pregnancy rate was equivalent among groups and ranged from 88% to 96%. No females aborted, delivered early, or were removed from the study. Gestational body weights and body weight gains were reduced in the 500 and 1000 mg/kg body weight/day groups, as was food intake for the first 7 days, indicating that a sufficiently high dose was achieved. Treatment had no effect on gestational parameters including resorption, pre- and postimplantation losses, percentage of live fetuses, and sex ratios nor did it affect the fetal body weights or the incidence of fetal malformations. However, two common skeletal variations (i.e., nonossification of centrum of vertebrae 5 or 6) had a higher incidence in the two highest dosage groups. These findings were not considered treatment-related, as adjacent vertebrae did not have delayed ossification. The NOAEL for maternal toxicity in the rat was 200 mg/kg body weight/day and the NOAEL for developmental toxicity was at least 1000 mg/kg body weight/day.
Rabbits New Zealand White female rabbits were gavaged with PBO (purity 100%) in corn oil at levels of 50, 100, or 200 mg/kg body weight between day 7 and day 19 of pregnancy (Leng et aI., 1986). Caesarian sections were performed on day 29 of gestation. Maternal toxicity was evident at 100 and 200 mg/kg body weight/day manifested by decreased defecation and a dose-dependent weight loss during the treatment period (these weight losses were recovered post-treatment). Common developmental defects, including an increase in the number of full ribs and the presence of more than 27 presacral vertebrae, were recorded in all dose groups. However, no doseresponse relationship was apparent. The number of litters in the treated groups with these observations was not increased when compared with control values. The NOAEL for maternal toxicity was 50 mg/kg body weight/day, whereas the NOAEL for developmental toxicity was 200 mg/kg body weight/day. 64.3.4 CHRONIC TOXICITY/ONCOGENICITY
Numerous long-term toxicity and oncogenicity studies have been undertaken on piperonyl butoxide over the past 50 years in various species. As evidenced in these studies, the primary target organ is the liver. The results of the studies discussed below are summarized in Table 64.5. 64.3.4.1 Dogs
PBO was administered to dogs in capsule form for a I-year chronic dietary toxicity study (Sarles and Vandergrift, 1952). Groups of four dogs each were treated at dose levels corresponding to 3, 32, 160, or 320 mg/kg body weight/day. The dosage was adjusted in accordance with any alteration in body weight to maintain the same dose in mg/kg body weight, except for one individual animal per dosage group, which received a constant absolute dose throughout the trial. All dogs belonging to the two highest dosage groups lost weight; however, meaningful comparisons between the lower dose groups and control animals were not possible owing to large variations in body weight gains and the small number of animals involved. All dogs at the highest dosage level died. However, no toxic reaction was seen at 3 mg/kg body weight/day. Red blood cell (RBC) and white blood cell (WBC) counts were unchanged at all dose levels. There was a dose-dependent increase in liver, kidney, and adrenal weights. Microscopic changes were quite similar to those in long-term toxicity studies performed in rats, with the liver again being the major target organ for toxicity. Hydropic swelling was evident in hepatocytes in the mid-dose group, with hepatic dystrophy and dysplasia becoming more obvious at the two highest dosage levels. The NOAEL for this study was 32 mg/kg body weight/day. A more recent I-year chronic dietary toxicity study was conducted with PBO in the beagle dog (Goldenthal, 1993b). Groups of four males and four females were fed PBO for 1 year at doses of 100, 600, or 2000 ppm (approximately equivalent to 2.5, 15, or 50 mg/kg body weight/day, respectively) in the diet. All animals survived to study termination. A reduction in
64.3 Hazard Identification
1469
Table 64.5 Summary of Results of Chronic Toxicity/Oncogenicity Studies with PBO Species
Route
Dose
Duration
NOAEL
Comments
Reference
Dog
Oral
3-320 mg/kg body weight/day
1 yr
32 mg/kg body weight/day
Increased liver and kidney weights, hepatic dystrophy and dysplasia
Sarles and Vandergrift (1952)
Diet
100-2000 ppm mg/kg body weight/day)
1 yr
600ppm (~15 mgikg body weight/day)
Increased liver and gall bladder weights, hypertrophy of hepatocytes
Goldenthal (1993b)
Diet
300 or 1112 ppm (~45 or 167 mg/kg body weight/day)
69wks
Not applicable
No significant increase in tumor incidence
Innes et al. (1969)
Diet
45 or 133 mgikg body weight/day
18 months
45 mg/kg body weight/day
No signs of toxicity
Bond et al. (1973)
1036-2804 ppm mgikg body weight/day)
112 wks
Not established
Decreased body weight in both sexes, hepatic nodular hyperplasia in males
U.S. National Cancer Institute (1979)
Diet
6000--12,000 ppm (960--1920 mg/kg body weight/day)
1 yr
Not established
Hepatic adenomas and hepatocarcinomas
Takahashi et al. (1994b)
Diet
30--300 mg/kg body weight/day
78 wks
30 mg/kg body weight/day
Increased liver weight, benign hepatic adenomas
Butler et al. (1998)
Diet
100--25,000 ppm mg/kg body weight/day)
2 yr
100ppm mg/kg body weight/day)
No significant increase in tumor incidence; severe liver damage, increased incidence of liver "hyperdysplastic" nodules
Sarles and Vandergrift (1952)
Diet
~90 mg/kg body weight/day
2 yr
Not applicable
Decreased body weight
Hunter et al. (1977)
Diet
5000--10,000 ppm (~250--500 mg/kg body weight/day)
107 wks
Not established
Dose-related increase in hepatocytomegaly, dose-related increase in lymphomas in female rats, but incidence in controls was also high
Cardy et al. (1979)
Diet
5000--10,000 ppm (~250--500 mg/kg body weight/day)
2 yr
Not established
No significant increase in tumor incidence; dose-related increase in ileocaecal ulcers
Maekawa et al. (1985)
Diet
6000-24,000 ppm (526-2187 mg/kg body weight/day)
95-96 wks
Not established
Hepatic adenomas and hepatocarcinoma at mid- and highdose levels, caecal hemorrhaging, severe general and hepatic toxicity
Takahashi et al. (l994a)
Diet
15-500 mg/kg body weight/day
2 yr
30 mg/kg body weight/day
Increased liver and kidney weights, centrilobular hepatocyte hypertrophy, focal hyperplasia
Butler et al. (1998)
Diet
2.0mLlday (~1000 ppm)
1 yr
Not applicable
Slight hepatic dystrophy and dysplasia
Sarles and Vandergrift (1952)
(~2.5-50
Mouse
Diet
(~148-298
Rat
(~5
(~5-1250
Goat
body weight gain and food intake was evident in the 2000 ppm group. Physical examinations were otherwise normal throughout the test period. Biochemical analysis showed increases in serum alkaline phosphatase levels at 6 and 12 months in the highest dosage group. Female beagles showed a decrease in serum cholesterol at the 2000 ppm dosage level. Increased liver and gall bladder weights, with mild hypertrophy of hepatocytes, were also recorded at this highest dosage level. A small increase in thyroid gland and parathyroid gland weights was also noted. However, no microscopic abnormalities were detected in the thyroid gland. No treatment-related histopathologic changes
were seen on the study. Based on the changes seen in the liver, the NOAEL for this study was 600 ppm.
64.3.4.2 Mice Innes and co-workers (1969) studied the effect of PBO on tumorigenicity in mice by administering the maximal tolerated dose. Animals (I8/sex/strain) from two hybrid stocks (C57BLl6 x C3H1Anf or C57BLl6 x AKR) were gavaged with 100 mg undiluted PBOlkg body weight or 464 mg PBOlkg body weight in solvent vehicle from 7 to 28 days of age. Thereafter, they re-
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CHAPTER 64 PBO Safety
ceived 300 ppm undiluted PBO or 1112 ppm PBO in solvent vehicle (approximately equivalent to 45 and 167 mg/kg body weight/day, respectively) in the diet for 69 weeks. These researchers found no significant increase in tumor incidence as a result of PBO treatment. Bond et a!. (1973) reported no adverse effects following dosing of mice with 45 or 133 mg/kg body weight/day PBO in the diet for 18 months. Few details are available for this early study, however. The D.S. National Cancer Institute (1979) conducted a mouse carcinogenicity study in which male and female B6C3F1 animals (50/sex/dose level) were initially dosed with 2500 or 5000 ppm of PBO in the diet. Toxicity appearing in both these groups resulted in a reduction in the doses to 500 and 2000 ppm, respectively after 30 weeks of dosing. The timeweighted average doses in the diets were approximately 1036 and 2804 ppm (approximately equivalent to 148 and 298 mg/kg body weight/day, respectively). Dose-dependent decreases in body weight and body weight gains were observed in both dose groups and both sexes. Nodular hyperplasia of the liver was slightly elevated in males. Although tumors were observed in the liver and lacrimal gland, the incidence was not statistically significant. Thus, the authors concluded that PBO was not oncogenic. The Tokyo Metropolitan Research Laboratory reported the results of a I-year chronic toxicity study conducted in CD1 mice using dietary doses of 6000 and 12,000 ppm PBO (equivalent to 960 and 1920 mg/kg body weight/day, respectively) (Takahashi et a!., 1994b). Animals were allocated to three groups consisting of 52, 53, and 100 animals for dose levels of 0,6000, and 12,000 ppm, respectively. Significant depressions in body weight and body weight gain were noted for low and high doses of PBO. Only 81 % of the high-dose animals survived the 12-month study period compared to 98% and 94% of the animals in the low-dose and control groups, respectively. Hepatic adenomas and hepatocarcinomas were observed in both treatment groups as were hemangiosarcomas and hemangioendothelial sarcomas. However, the doses used in this study are clearly in excess of internationally accepted maximum tolerated dose (MTD) criteria and thus they are of questionable relevance in determining the hazard and risk for humans of PBO exposure. Butler et a!. (1998) report a study in which groups of 60 male and 60 female CD-1 mice were administered PBO in the diet at doses of 0 (two separate control groups), 30, 100, or 300 mglkg body weight/day for at least 78 weeks. No treatment-related clinical signs of toxicity or changes in food consumption or clinical chemistry were observed. The mean absolute body weight and mean body weight gains were generally slightly decreased throughout the study at the high dose in both males and females, indicating that the MTD was reached. A dose-related increase in the mean absolute and relative liver weights was seen in the mid- and high-dose groups of both sexes. The mean absolute and relative liver weights of the low-dose group of male mice were also slightly increased. Both males and females clearly showed an increased incidence of be-
nign hepatic nodules diagnosed as adenomas. The further characterization of the adenomas showed that the increased burden of lesions was due to the increased incidence of eosinophilic adenomas, similar to the lesions induced by a range of enzyme inducers in the mouse (Butler, 1996). There was no increase of either basophilic adenomas or hepatocarcinomas. The NOAEL in this study was 30 mg/kg body weight/day.
64.3.4.3 Rats In an early study, Sarles and Vandergrift (1952) fed Wistar rats with diets containing from 100 to 25,000 ppm (approximately equivalent to 5 to 1250 mg/kg body weight/day) PBO for 2 years. Twelve males and 12 females were used at each dose level. The entire high-dose group died by week 68 and showed severe liver damage upon necropsy. An increased incidence of "hyperdysplastic" hepatic nodules, characterized by the authors as the appearance of larger cells and increased polyploidy, was seen in the treated groups. Dystrophy and dysplasias were also observed in the livers from animals fed 1000 ppm or greater PBO. The authors concluded that there was no evidence of carcinogenicity; the 100 ppm dose level was considered "nontoxic." Because PBO is most often used as a synergist with pyrethrins, a 2-year dietary study was conducted in SpragueDawley rats using a mixture of pyrethrins (53.1 % purity) and piperonyl butoxide (95% purity) (Hunter et a!., 1977). Fortyfive males and 45 females were fed diets containing 400 ppm pyrethrins plus 2000 ppm piperonyl butoxide. The average daily doses received by the animals over the study period were 16 + 79 mg/kg body weight/day (pyrethrins + PBO) for males and 20 + 101 mg/kg body weight/day (pyrethrins + PBO) for females. Body weights were depressed in the females during the first 78 weeks of treatment and among males during the first 26 weeks of treatment. No other treatment-related effects were noted and no treatment-related change in tumor incidence was seen. The D.S. National Cancer Institute conducted a two-year cancer bioassay in Fisher 344 rats (Cardy et a!., 1979). Fifty male and 50 female animals per dose level were allocated to low- and high-dose groups which received PBO (88.4% purity) in the diet at 5000 or 10,000 ppm (approximately equivalent to 250 or 500 mg/kg body weight/day, respectively) for 107 weeks. A dose-dependent decrease in the mean body weights of treated groups was noted. Other than increased hepatocytomegaly, no dose-related increases in the incidence of tumors or other microscopic findings were observed in the liver. The hepatocytomegaly consisted of foci of enlarged hepatocytes, often associated with large, vesicular nuclei and numerous cytoplasmic vacuoles, giving the cytoplasm a "ground glass" appearance. Distortion of lobular architecture in these foci was minimal, and trabeculae were continuous with adjacent normal hepatocytes. These lesions appear similar to those described by Squire and Levitt (1975) as "eosinophilic foci," "ground glass foci," or "clear cell foci." Although a dose-dependent increase in lymphomas was noted in females, the incidence of lymphomas, leukemias, and
64.3 Hazard Identification reticuloses observed was not significantly different from the historical rates from the laboratory. Thus, this study showed that, under the conditions of the bioassay, PBO was not carcinogenic in Fischer 344 rats. The carcinogenicity of piperonyl butoxide was also studied in F344IDuCrj rats by Maekawa et al. (1985). Animals (50 sex/dose level) were fed a dietary level of 5000 or 10,000 ppm (approximately equivalent to 250 and 500 mg/kg body weight/day, respectively) for 2 years but no significant dose-related increase in the incidence of any tumor was found. A dose-related incidence of ileocaecal ulcers, however, was found in animals of both sexes. The Tokyo Metropolitan Research Laboratory conducted a 2-year chronic toxicity study in the rat at dose levels up 24,000 ppm (1000 times the maximum level approved in raw cereals in Japan) (Takahashi et aI., 1994a). Fischer F344IDuCrj rats (30-33 per group) received a diet containing PBO at 6,000, 12,000 or 24,000 ppm (equivalent to 526,1052, or 2187 mg/kg body weight/day, respectively) for 95-96 weeks. Beginning at about 40 weeks, 10 rats in the 12,000 ppm male group died due to caecal hemorrhages. By the end of the study, gastrointestinal hemorrhage occurred at all dose levels. Organ weights (with the exception of the liver) were reduced in all animals in the high-dose group. "Probable essential thrombocytopenia" was present in all treated male groups. Body weight gains relative to controls were reduced in all treated groups of both sexes and reached approximately 50% in the high-dose group. A dosedependent increase in hepatocellular hyperplasia (seen as liver nodules) was reported. Although the nomenclature is different, these lesions are much like those described by Sarles and Vandergrift (1952) at toxic doses of PBO. Takahashi also reported hepatocellular adenomas and carcinomas in the mid- and highdose groups. It is important to note that the study was not intended to be a carcinogenicity study and the procedures for collecting and examining tissues were not performed according to current USEPA/OECD standards. Thus, not all tissues were taken or prepared for histological examination. Because of the highdose levels used and resulting toxicity, it is difficult to interpret the carcinogenicity findings and their relevance for hazard assessment. Moreover, several investigators have reported that hepatotoxicity, and the resulting regenerative hyperplasia, can contribute to the formation of liver tumors by nongenotoxic mechanisms (Kociba et aI., 1978; McClean et aI., 1990; Mutai et aI., 1990; Tatematsu et aI., 1990; Van Miller et aI., 1977). This and other mechanistic aspects of this oncogenicity response are discussed in Section 64.3.5. The most recent report is from another dietary study in which the Sprague-Dawley rat was utilized (Butler et aI., 1998). Animals were divided into groups of 60 animals of each sex and were administered 15, 30, 100, or 500 mg PBO/kg body weight/day. Three control groups were included. Because a 4-week range-finding study did not provide clear evidence of a NOAEL with respect to minor alterations in liver cell morphology, additional animals were used during the early stages of this study. Thus, after completion of 4 weeks of treatment, 10 males
1471
and 10 females from each low-dose group and a special control group consisting of 10 males and 10 females were sacrificed for both gross pathological and histopathologic examinations of the liver. Since the results of the histopathologic examination showed no abnormal findings in the livers of these rats, the low dose level of 30 mg/kg body weight/day was continued on the 2-year study. The 15 mg/kg body weight/day group was discontinued. No adverse treatment-related effects on survival and no treatment-related clinical signs were seen. A reduced growth rate and a minimal reduction in food intake were noted for males and females receiving 500 mg/kg body weight/day. These females were shown to have increased serum cholesterol levels and slightly higher total serum protein levels. The blood urea nitrogen levels were also slightly higher in this group on one occasion. Dose-related increases in liver and kidney weights were noted in both sexes at 100 and 500 mg/kg body weight/day. Histologically, the most common effect in the liver was centrilobular hepatocyte hypertrophy and the presence of eosinophilic and mixed (basophilic and eosinophilic) cell foci. The severity of focal hyperplasia in the liver was also greater in the intermediate- and high-dose groups. The hyperplastic foci contained either basophilic, normal, or enlarged eosinophilic cells that were variable in size. Normal lobular architecture was retained. Portal triads and central veins were present in the lesions. Neither the incidence of adenomas nor incidence of carcinomas was increased. Pituitary adenomas were common in both males and females but showed no treatment-related effect upon incidence of the adenomas. Thyroid changes, including increased pigment in colloid, and follicular hyperplasia (particularly at 500 mg/kg body weight/day) were seen in both sexes at the end of the study as well as in the high-dose group males dying or sacrificed during the study. In the kidney, glomerulonephritis was more common in the male than the female. No significant increase in the incidence of glomerulonephritis was observed, but the severity of the lesions was increased slightly in the intermediate- and high-dose groups. The non-neoplastic changes observed in this rat study are consistent with induction of the hepatic mixed function oxidase system. The liver observations such as increased liver weight, centrilobular hypertrophy, eosinophilic foci, and eosinophilic focal hyperplasia are probably due to enzyme induction. Likewise, the thyroid follicular hyperplasia is a likely secondary response of the thyroid gland to prolonged TSH stimulation resulting from decreased circulating levels of T3 and T4. Reduced T3 and T4 levels are due to increased conjugation and excretion of the thyroid hormones resulting from liver enzyme induction. This pattern has been observed with other liver enzyme inducers such as phenobarbital (McClain, 1989). Given these considerations, the authors concluded that there was no evidence of carcinogenic activity and that the NOAEL based on liver changes was 30 mg/kg body weight/day.
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CHAPTER 64 PBO Safety
64.3.4.4 Other Species Sarles and Vandergrift (1952) also reported a chronic oral toxicity experiment where a mature female goat was fed a daily dose of 2.0 mL PBO by capsule, 6 days a week, for 1 year. This dose equated to approximately 1000 ppm PBO in the diet. The dose started 4 days after the goat gave birth to a female kid, with both dam and offspring being observed for 1 year to ascertain any signs of direct or indirect (i.e., PBO in dam's milk) toxic effects. The general health of both dam and kid was unaffected by treatment. The kid was nursed by the treated dam for approximately 6 months and continued to grow and thrive as expected. RBC and WBC counts were unremarkable. At postmortem examination, the dam's liver revealed slight dystrophy and dysplasia, with central hydrophic swelling and slight fatty accumulation. No abnormalities were detected in the organs of the kid goat.
64.3.5 MECHANISTIC CONSIDERATIONS FOR ONCOGENICITY The chronic toxicity/oncogenicity studies discussed in Section 64.3.4 indicate that liver is a target for both oncogenic and nononcogenic changes in both the mouse and the rat. In addition, one study showed an apparent hyperplastic response in the thyroid (Butler et aI., 1998). Many nongenotoxic compounds have been shown to induce tumors in rodents (Ames et aI., 1993; Butler, 1996; Cohen and Ellwein, 1990; Grasso and Hinton, 1991; Grasso et aI., 1991; Loury et aI., 1987; Wilson et aI., 1992). Although the mechanism(s) of liver tumor production by nongenotoxic inducers of hepatic xenobiotic metabolism, such as sodium phenobarbital (NaPB), remains to be fully elucidated (Grasso and Hinton, 1991; Grasso et aI., 1991), it is clear that the mitogenic and promotional effects of such agents are important. Tumor formation may involve the promotion of particular populations of hepatocytes through differential effects on cell replication, growth factors, intercellular communication, etc. (Anderson et aI., 1995; Grasso and Hinton, 1991; Grasso et aI., 1991; Jirtle, 1994; Law, 1991; Lubet et aI., 1989; Neveu et aI., 1994; Whysner et aI., 1996). It is important to note that threshold doses have been reported for both liver enlargement/enzyme induction and, at higher dose levels, for tumor formation (Grasso and Hinton, 1991; Grasso et aI., 1991). There is evidence that such lesions may regress upon withdrawal of the inducing compound (Evans et aI., 1992; Ito et aI., 1976; Malarkey et aI., 1995). Other studies have shown that the lesions found in MFO-induced mice are fundamentally different from both spontaneous lesions and those induced by genotoxic carcinogens in that they do not express oncogenes (Fox et aI., 1990; Rumsby et aI., 1991) and fail to grow in semisolid agar (Pedrick et aI., 1994). Thus, although these lesions are usually diagnosed as adenomas, the evidence suggests that the lesions are compound-dependent, rather than autonomous, and do not progress to hepatic carcinoma.
Besides the additive hyperplasia that may occur from mitogenic and promotional agents, tumors may also result from regenerative hyperplasia produced in response to cell necrosis (as with chloroform and furan). Such a mechanism would also be expected to demonstrate a threshold for effects, including tumorigenesis. Substantial evidence suggests that both additive and regenerative hyperplasia may result from sufficiently high doses ofPBO. Several studies have shown that PBO is an inducer of hepatic xenobiotic metabolism in the mouse and rat (Fennell et aI., 1980; Goldstein et aI., 1973; Lake et aI., 1973; Phillips et aI., 1997; Wag staff and Short, 1971). Phillips et al. (1997) characterized the induction of enzyme activities in both the mouse and rat and compared it to the classic inducer, phenobarbital. Four groups of 16 male F-344 rats were fed PBO in the diet at 100, 550, 1050, or 1850 mg/kg body weight/day. Animals were treated for either 7 or 42 days and were sacrificed, and liver studies were performed. Even the low-dose group (100 mg/kg body weight/day PBO) showed increased relative liver weights and microsomal protein content (42 days treatment), increased cytochrome P-450 levels (7 days treatment), increased GGT levels (42 days treatment), and increases in certain MFO enzyme activities. PBO appeared to be a mixed-type enzyme inducer in the rat in that it induced hepatic cytochrome P450 isoenzymes in the CYPIA, CYP2B, and CYP3A subfamilies. Recent work by Watanabe et al. (1998) in the rat is in agreement with these findings of Phillips. These workers also noted weak induction of the CYP4A isozyme. The induction pattern was similar to that they observed for NaPB, with the exception that NaPB did not induce CYP1A. A NOEL of 0.05% PBO in the diet (approximately 50 mg/kg body weight) over 4 weeks was reported for this enzyme induction. PBO treatment of the mouse by Phillips and co-workers also resulted in a dose-related induction of cytochrome P450 content and ethylmorphine demethylase activity (CYP3A). Taken with other studies that have shown induction of CYPIA and CYP2B isoenzymes in mouse liver (Adams et aI., 1993; Fennell et aI., 1980), PBO appears to be able to induce CYPIA, CYP2B, and CYP3A isoenzymes in CD-l mouse liver as well. Aside from the enzyme induction discussed above, certain nongenotoxic rodent liver carcinogens produce either a transient or a sustained stimulation of cell replication (Goldsworthy et aI., 1991). Both PBO and NaPB produced a stimulation of cell replication after 7 but not 42 days of treatment in the mouse (Phillips et aI., 1997). Like NaPB, PBO also increased relative liver weight in CD-l mice and produced liver hypertrophy, although a difference in the lobular distribution of this effect was noted. Generally, the effects of PBO on relative liver weight, liver morphology, replicative DNA synthesis, and xenobiotic metabolism occurred at the 100 and 300 mg/kg body weight/day dose levels, the same doses where eosinophilic nodules were observed in a 2-year study by Butler et al. (1998). In addition, the eosinophilic nodules produced by PBO in mouse liver (Butler, 1996) appear similar to those formed by NaPB (Evans et aI., 1992).
64.3 Hazard Identification
Thus, the data show that PBO is an inducer of the MFO enzymes and hepatocyte proliferation in the mouse and such induction results in enlarged livers, centrilobular hypertrophy and hyperplasia, and an increase of benign eosinophilic adenomas by a mechanism similar to that of NaPB (Butler, 1996; Evans et aI., 1986; lones and Butler, 1975; Phillips et aI., 1997). With respect to the rat, Watanabe et al. (1998), demonstrated the production of centrilobular hypertrophy in the rat liver following 4 weeks of dosing with 2% PBO in the diet. The degree of response was similar to that observed after 4 weeks dosing with 0.1 % NaPB. Further evidence of the similarity between the action of PBO and NaPB comes from a report by Okamiya et al. (1998) showing an increase in proliferating cell nuclear antigen in rats fed 0.2% PBO in the diet for 4 weeks. They also demonstrated a decrease in the gap junction protein connexin 32 (Cx32) at the highest dose tested (2.0%) after dosing for 1,2, but not 4 weeks. Phillips et al. (1997) reported that high doses of PBO (1850 mg/kg body weight/day) given to rats caused a significant reduction of body weight gain and of food consumption throughout the 42 days of dosing. Morphological examination of liver showed individual cell necrosis in rats dosed with 1050 and 1850 mg PBO/kg body weight/day. While the severity of the individual cell necrosis was similar in rats given 1050 and 1850 mg PBO/kg body weight/day, the incidence was greater (seven of eight animals examined) in rats given 1850 mg/kg body weight/day. Like NaPB, the increase in relative liver weight in PBO-treated animals was associated with hypertrophy, although a difference in the lobular distribution of this effect was noted. Replicative DNA synthesis was stimulated by 550 and 1050 mg PBO/kg body weight/day and 0.05% NaPB after 7 days of administration, most likely due to transient mitogenesis typical of enzyme inducers. In contrast, the stimulation of cell replication observed after 42 days treatment by 1050 mg/kg body weight/day is more likely to be associated with the onset of a regenerative hyperplasia. The mitogenic and hypertrophic effects of PBO were observed at doses (e.g., 550 mg/kg body weight/day) lower than those required to produce individual cell necrosis, where a high incidence of necrosis was only observed in rats given 1850 mg/kg body weight/day for 42 days. Chronic treatment (i.e., >42 days) with PBO at high-dose levels such as that in an oncogenesis study could result in a sustained stimulation of replicative DNA synthesis and an increased likelihood of oncogenesis. It is important to note that Takahashi et al. (1994b) reported no increase in liver tumors following 2 years of dosing at the lowest study dose of 547 mg/kg body weight/day, a dose calculated by the authors to be about 18,000 times the allowable daily intake (ADI) for humans. In contrast, higher doses of PBO were both toxic to the rat liver and produced tumors. Thus, the data suggest that eosinophilic nodules in mouse liver may result from a mechanism similar to that of NaPB and other enzyme inducers, whereas tumor formation in rats at greater than MTD doses is most likely related both to significant induction of hepatic metabolism in conjunction with a regenerative hyperplasia resulting from PBO-induced hepatotoxicity. Both of these mechanisms are threshold phenomenon
1473
and suggest that at doses likely to be encountered by humans, PBO poses essentially no oncogenic risk. 64.3.6 GENOTOXICITY Piperonyl butoxide has shown no evidence of mutagenic activity in a number of bacterial assays involving Salmonella typhimurium, Bacillus subtilis, and Escherichia coli both in the presence or absence of rat liver microsomes (S-9) (AshwoodSmith et aI., 1972; Butler et al., 1996; Ishidate et aI., 1984; Kawachi et aI., 1980; Moriye et aI., 1983; White et aI., 1977). Most of the studies in systems using mammalian cells in culture show no evidence of mutation or a chromosome damaging effect. Galloway et al. (1987) investigated a wide range of compounds including piperonyl butoxide in Chinese hamster ovary cells and failed to find chromosome aberrations and sister chromatid exchange in the presence or absence of rat liver S-9. PBO had no effect on Chinese hamster ovary cells in the report of Butler et al. (1996), and produced a small increase in sister chromatid exchanges only in the absence of S-9 in the recent study by Tayama (1996). Tayama also concluded that the metabolites of PBO are unlikely to be genotoxic. In addition, piperonyl butoxide did not produce chromosomal aberrations in Chinese hamster lung cells (Ishidate et aI., 1984, 1988; Kawachi et aI., 1980) or induce mutations in the CHOIHGPT assay (Butler et aI., 1996). However, in L5178Y mouse lymphoma cells, piperonyl butoxide showed evidence of mutagenic activity only in the absence of additional metabolic activation. In this study, the mutagenic activity was observed only where cytotoxicity was evident. The relative total cell growth at the lowest mutagenic concentration of 30 )l.g/mL was around 60% (McGregor et aI., 1988). Piperonyl butoxide also induced cell transformation in Syrian hamster embryo cells (Amacher and Zelljadt, 1983). In this study, three dose levels (0.5, 1.0, and 3.0 )l.g/ml), were used and only 2 transformed colonies out of 2761 were observed. No indication is given of the dose level that caused the 2 transformed colonies. Suzuki and Suzuki (1995) investigated the mutagenicity of piperonyl butoxide in human RSA cells, a cell line of double transformed human embryonal fibroblasts considered to be hypermutable, by determining ouabain resistance. The results show an unusual dose response in that despite having little effect upon survival at dose levels above 0.2 )l.g/ml piperonyl butoxide, the incidence of mutation declined. The authors also report mutation of K-ras codon 12 with an apparent similar dose response. While K-ras mutation has been reported in human tumors at various sites (Almoguera et al., 1988; Bos et aI., 1987) no such association has been observed in rodent liver tumors the apparent target site of piperony1butoxide (Maronpot et aI., 1995). Butler et al. (1996) reported that PBO did not induce unscheduled DNA synthesis in rat hepatocytes. Moreover, Beamand et al. (1996) showed a similar negative response to PBO in cultured human hepatocytes. In vivo studies have also failed to demonstrate convincing genotoxic effects of piperonyl butoxide. A dominant lethal as-
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PBO Safety
say in ICRlHa Swiss mice using both single and multiple doses of piperonyl butoxide given either by intraperitoneal injection or by gavage resulted in toxicity and death of the male mice. Although there was some evidence of reduced reproductive efficiency and an increase in early fetal death, the results were not consistent and the authors concluded the study was equivocal (Epstein et al., 1972). Other studies have been reported only briefly as abstracts and have stated that no chromosomal aberrations or sister chromatid exchanges were produced in either rat or mouse bone marrow (Ivett and Tice, 1983; Kawachi et al.,1980). 64.3.7 HUMAN STUDIES Wintersteiger and Juan (1991) investigated the absorption of combination pyrethrin and PBO sprays across the skin of six healthy subjects in Austria. The spray was applied over a wide area of the back with a total dose of approximately 3.3 mg pyrethrum extract and 13.2 mg PBO being applied. No untoward clinical signs were noted. Cutaneous absorption of PBO was shown to be extremely low, with plasma samples containing no more than lOng PBO/mL. Wester et al. (1994) investigated the percutaneous absorption of both PBO and pyrethrin compounds across the skin of the ventral forearm in six volunteers. Based on the recovery of radioactivity in the urine, it was calculated that 2.1 % ± 0.6% of the dose of PBO was absorbed through the skin. Not surprisingly, higher levels of absorption of PBO were achieved when the compound was applied to the skin of the scalp (8.3%). There was no evidence of any local or systemic toxicity of PBO when used as a topical agent in humans. Selim (1995) investigated the absorption, excretion, and mass balance of 14C PBO from two different formulations following dermal application to healthy volunteers. The first preparation applied was a 4% (w/w) solution of PBO in an aqueous formulation. This product was applied to four healthy human volunteers. The second preparation tested was a 3% (w/w) solution of PBO in isopropyl alcohol. In the former case, the mean amount of PBO applied was 3.8 mg per volunteer (approximately 39.9 J..LCi of radioactivity per volunteer), while the average exposure was 3.0 mg PBO per volunteer (approximately 40 J..LCi of radioactivity per volunteer) in the case of the isopropyl alcohol solution. Results from this study show that there was a similar dermal absorption pattern for both formulations. The principal route of excretion of absorbed radioactivity was via urine. Fecal samples from volunteers contained negligible levels of radioactivity. Dermally applied PBO was rapidly excreted from the volunteers. The majority of the applied radioactivity remained at the application site, with less than 3% of the applied dose being absorbed during the 8-hour test period. Radioactivity did not accumulate in the skin. The only study to examine the ability ofPBO to inhibit xenobiotic metabolism in humans was that by Conney et al. (1972). Using the rate of antipyrine metabolism as a gauge of P-450 activity, two healthy men (weighing 87.4 kg and 82.6 kg, respectively) were given capsules containing increasing amounts
(5, 10, 20, and 50 mg) of PBO at consecutive intervals of approximately 1 week. No signs of toxicity were seen. Clinical chemical analysis of blood and urine samples taken at 4, 8, and 24 hours after ingestion of the capsules did not reveal any adverse effects. In a subsequent study, eight men were given PBO as a single dose of 0.71 mg/kg body weight. A control group received a placebo. Two hours later, both the treated and control groups received a 250 mg oral dose of antipyrine. Antipyrine was analyzed in blood samples taken at intervals over the next 31 hours. PBO had no effect on the rate of clearance of antipyrine. Although no systematic epidemiology studies have been conducted on PBO-exposed individuals, no evidence suggests that PBO has resulted in any significant adverse effects to human health. Occupational health data collected at a PBO manufacturing site in Italy, consisting of routine health checks, have been made on potentially exposed workers from 1974 to 1994. Sixty workers were examined, 11 of which were employed in the manufacture of PBO for periods of 15-26 years. Workers received x-rays and spirometric evaluations every three years and annual clinical chemistry analysis. No adverse clinical signs or symptoms related to PBO were found (Endura, 1996). Similarly, at a manufacturing site in Scotland, where PBO was made from 1962 to 1990, "no cases of toxic symptoms or adverse effects attributable to PBO manufacture" were noted in production workers at the plant. Moreover, no adverse effects were reported in operations involved in the handling and use of PBO at a site in England (JMPR, 1993; Pitman Moore, 1990; Wellcome, 1991). The only clinical report referring to PBO exposure is the case of two sisters who gave birth, within 2 weeks of each other, to children who each had coarction of the aorta (Hall et al., 1975). Both mothers had been on a camping trip at 2 months gestation where they used "large amounts" of insect repellents and insecticides containing, among other chemicals, PBO, pyrethrins, DEET, and DDVP. No cause-and-effect relationship was established. There have been no reported suicide attempts by humans with PBO. Based on its acute toxicity to experimental animals, a probable oral lethal dose for humans is estimated to be 5-15 g/kg body weight (i.e., approximately 300 to 900 g PBO for a 60 kg human).
64.4 PHARMACODYNAMICS Yamamoto (1973) suggests that the primary function of synergists such as PBO when formulated with pyrethrins or pyrethroids is to provide an alternative substrate for the MFO enzyme system, which would normally metabolize such insecticides. Inhibition of MFO-mediated oxidation of the transmethyl groups and the alcohol moiety on the pyrethrin molecule appear to be the most important functions of PBO. Inhibition of ester hydrolysis may also contribute to the effectiveness of PBO as a synergist.
64.5 Risk Characterization
As a known alternative substrate for the liver microsomal enzyme system, PBO will inhibit the metabolism of many xenobiotics including drugs and pesticides. Brown (1970) reported that the detoxification of certain drugs such as pentobarbital, zoxazolamine, antipyrine, and benzopyrene were inhibited by PBO, presumably due to the inhibition of their microsomal oxidation. Conney et al. (1972) investigated the inhibition of antipyrine metabolism in rats and mice. Both species were treated intraperitoneally (i.p.) with a single dose of PBO, followed by a further i.p. injection of antipyrine (200 mg/kg body weight) 1 hour later. A marked species difference was noted in the response; the NOEL for inhibition of antipyrine metabolism in the mouse was 0.5-1.0 mg PBO/kg body weight, whereas the NOEL for the rat was 100 mg PBO/kg body weight. The effects of PBO on the metabolism of benzopyrene in Sprague-Dawley rats were studied by Falk et al. (1965). PBO was administered by the oral, intraperitoneal (i.p.), or intravenous (i.v.) routes at various times before the i.v. injection of labeled benzopyrene. The level of radioactivity was then measured in bile at frequent intervals up to 4 hours. This author demonstrated marked inhibition of benzopyrene metabolism when PBO was administered i. v. at 262 mg/kg body weight, some 5 minutes to 16 hours before the administration of bezopyrene. However, this effect is much reduced at 121 mg/kg body weight. Virtually no effect is seen at 25 hours postdosing. This implies that single large doses of PBO are quickly metabolized by rats. Administration of PBO by the oral and i.p. routes resulted in a greatly reduced effect when compared with the i. v. route. A second similar study performed by Conney et al. (1972), where the effects of i.p. administration of PBO on the metabolism of benzopyrene were investigated, showed less sensitive results when rats of lighter weight were used (approximately 180 g versus 400 g in Falk's study). It was postulated by Brown (1970) that the extra fat in the animals in Falk's study could possibly act as a reservoir of PBO and lead to a longer duration of action. It would appear from these studies in rats that 250 mg of PBO per kg body weight is the minimum oral dose required to give any significant effect on benzopyrene metabolism. Whereas a single dose of PBO will generally inhibit the metabolism of pentobarbital, repeated PBO doses will generally induce the metabolism of phenobarbital and other xenobiotics. Brown (1970) reports an experiment in rats where a single i.p. dose of PBO (333-1000 mg/kg body weight) increased the sleeping time of the animals following administration of pentobarbital. However, the i.p. administration of eight injections of 50 mg PBO/kg body weight, each at 12-hour intervals, followed by the injection of pentobarbital some 18 hours later, caused a reduction in sleeping time in rats. The administration of 50 mg PBO per kg body weight i.p. to rats (Anders, 1968) and mice (Graham et ai., 1970) prior to treatment with hexobarbital approximately doubled the sleeping time of both species. CD-l mice given a single i.p. dose of 600 mg PBO per kg body weight were found to have suffered less hepatotoxicity when treated with acetaminophen (600 mg/kg body weight, p.o.) at either 2 hours prior to or 1 hour following PBO ad-
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ministration. This reduced hepatotoxicity was measured via GSH and sorbitol dehydrogenase levels, as well as subsequent histopathology of the liver. Since the hepatic MFO system metabolizes acetaminophen to a toxic metabolite, the decreased toxicity seen in this experiment is likely due to inhibition of such oxidase enzymes by PBO (Brady et ai., 1988). Many other studies have been undertaken relevant to the pharmacodynamics of PBO and are reported elsewhere (Skrinijaric-Spoljar et ai., 1971; Conney et ai., 1972; Goldstein et ai., 1973). More recent work is discussed in Section 64.3.5 of this chapter.
64.5 RISK CHARACTERIZATION The U.S. EPA evaluated the weight of evidence relating to the potential oncogenicity of PBO and classified it as a Group C-Possible Human Carcinogen (U.S. EPA, 1995a). This was based on the increases in hepatocellular tumors in both male and female mice (adenomas, carcinomas, and combined adenomas and carcinomas in the males and adenomas only in the females). However, because of the generally low concern for mutagenicity, and the minor significance of other tumors observed in the rat studies, rather than recommend the QT linearized multi stage model for risk characterization based on oncogenicity, the U.S. EPA endorsed the use of a Reference Dose (RID) and Margin of Exposure (MOE) approaches using non-oncogenic endpoints, such as body weight changes. The Joint FAOIWHO Meeting on Pesticide Residues (JMPR) evaluated the toxicology of PBO in 1965, 1966, 1972, 1992, and, most recently in 1995. They recognized that, at doses up to internationally accepted standards for a Maximum Tolerated Dose, PBO is not oncogenic in the mouse or rat. Thus, based on the NOAEL of 600 ppm (16 mg/kg body weight/day) in the most sensitive toxicology study (1 year feeding study in dogs), they established an ADI for humans of 0.2 mg/kg body weight.
64.6 EXPOSURE AND RISK ASSESSMENT The first formal discussion of human exposure to PBO took place at a joint FAOIWHO Codex Alimentarius in 1987. FAOIWHO estimated that the average daily human diet (1.4 kg) might contain as much as 1 ppm of PBO, corresponding to a daily dose of 1.4 mg. This is likely an overestimation as cooking may destroy up to 90% of PBO present. FAOIWHO estimated the exposure from aerosol consumer products (from ingestion and inhalation) to be approximately 0.63 mg/day. Thus, the total intake from dietary and nondietary sources was estimated to be approximately 2.03 mg/day. This corresponds to a daily dose of 0.029 mg/kg body weight for a 70 kg adult and 0.14 mg/kg body weight for a 15 kg child, well below the AD! for humans of 0.2 mg/kg body weight for PBO set by the JMPR (1995). In a refinement of this exposure estimate, Crampton (1994) calculated that the daily exposure of adults to PBO residues in food
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PBO Safety
was 0.0037 mg/kg body weight/day, less than 1I50th the AD! set by JMPR. From the 1990s and through the present, the PBO Task Force (PBTF) and a subsequent entity, the Non-Dietary Exposure Task Force (NDETF), both consortia of PBO producers and marketers, have been developing exposure data for PBO. Their goal has been to estimate the aggregate human exposure to PBO from both dietary and nondietary routes. The PBTF conducted a series of crop residue studies on various crops representing 12 crop groups. In addition, studies were conducted where residue levels and transfer factors were obtained following application of PBO to livestock, and estimation of potential dietary intake of PBO was done using DietRisk™, a software tool that incorporates the USEPA's Dietary Risk Evaluation System (Burin, 2000). Determination of chronic dietary exposure requires the use of the anticipated PBO levels in a given food commodity (obtained from field studies) and the quantity of the commodity ingested. In addition, because PBO is not used widely on any given commodity, the anticipated value is adjusted by an estimate of the percentage of the crop that it is treated. Given the above information, the estimated dietary intake of PBO was 2.34 ug/kg body weight/day, about 1.2% of the JMPR AD!, and 7.8% of the U.S. EPA RID. Because many conservative assumptions were built into this assessment (e.g., no account for losses during washing and cooking of foods), it is likely that the actual human exposure is much lower. The NDETF is also currently conducting studies to define the various dermal and hand-to-mouth "transferability" parameters to support the development of a predictive stochastic model estimating potential distributions of postapplication exposure and absorbed doses following different residential exposure scenarios (e.g., indoor foggers, carpet and room aerosol sprays). This work is ongoing and will be published when available. As a precursor to the exposure monitoring program, the NDETF performed conservative screening-level assessments, based on existing information and default assumptions, to deterministically estimate potential exposures following the use of a PBO-containing total release indoor fogger (Dragula et aI., 1996). Such products are typically used to combat flea and roach infestations. Given the intermittent nature of this exposure scenario and most nondietary exposures to PBO-based consumer products, subchronic toxicity endpoints were used as the basis for calculation of margins of exposure (MOEs). Thus, MOEs were derived using the most sensitive subchronic study, the 90-day rat inhalation study. The NOAEL for systemic toxicity in this study was 155 mg/m 3 that converts to 15 mg/kg body weight using a conversion factor relating lung ventilation to body weight (0.365 ml/min/g; U.S. EPA, 1988). For purposes of the screening-level assessment for indoor foggers containing PBO, potential adult applicator exposures were estimated using relevant surrogate data from the EPA's Pesticide Handlers Exposure Database (U.S. EPA, 1995a, b). Potential postapplication exposures were estimated using the conservative assumption that consumers reenter treated areas immediately postapplication (when surface residues dry) and remain
Table 64.6 PBO Route-Specific and Total Absorbed Doses (Day of Application), and Corresponding MOE, associated with Exposures to Adults Involved in Application of an Indoor Fogger Product and Post-Application Activities in a Treated Residence Daily dose (mg/kg body weight/day) Exposure period
Inhalation
Dermal
Application
1.8 x 10- 5
Postapplication
7.2 x 10-4
3.5 x 10- 5 3.5 x 10- 5
7.4 x 10-4
1.0 x 10- 6 7.1 x 10- 5
+ Route-specific total Total daily absorbed dose
8.1 x 10-4
(mg/kg body weight/day) MOE
19,000
(based on total daily absorbed dose)
in these areas throughout the day. Inhalation exposures via indoor air, dermal exposures via contact with surface residue levels, and, in the case of toddlers, incidental ingestion exposures resulting from hand-to-mouth activities were estimated on the day of application. Inhalation exposures were conservatively based on surrogate data from an indoor chamber study and assumptions regarding exposure duration (i.e., 8 hours in the treated room). The assessment also assumed no decline in surface residues during this time. In the case of the postapplication dermal (for adults and children) and incidental ingestion (children) routes, conservative estimates were based on surrogate postapplication dermal (including hand measurements used in children's hand-to-mouth exposure estimation) exposure monitoring studies involving high contact activities (i.e., Jazzercizing®) (Ross et aI., 1990). The resulting absorbed doses for adults, including application and postapplication exposures, are shown in Table 64.6. The associated MOE (using the systemic NOAEL of 15 mg/kg body weight/day) for the estimated total absorbed dose across both routes is also shown in Table 64.6. The screening-level MOE exceed 100, indicating this particular adult, nondietary exposure scenario presents an insignificant health risk. The postapplication daily absorbed dose of PBO from inhalation, dermal contact, and incidental ingestion for a 10.2 kg toddler (child) are estimated to be 1.5 x 10- 3 , 1.3 X 10-7 , and 3.9 x 10-6 mg/kg body weight/day, respectively. For these calculations, dermal absorption was estimated to be 0.58% for aqueous/water-based formulations and 2.4% for isopropanol/solvent-based formulations based on Selim (1995). The total absorbed dose (sum across the three routes) is estimated to be approximately 1.5 x 10- 3 mg/kg body weight/day. The MOE based on this simplistic point estimate of total absorbed dose is approximately 10,000 (using the systemic NOAEL of 15 mglkg body weight/day), indicating no significant health risk for this particular age group.
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Sarles, M. P., and Vandergrift, W. B. (1952). Chronic toxicity and related studies on animals with the insecticide and pyrethrum synergist piperonyl butoxide. Am. 1. Trap. Med. 1, 862-883. Selim, S. (1995). "Absorption, Excretion, and Mass Balance of 14C Piperonyl Butoxide from Two Different Formulations after Dermal Application to Healthy Volunteers." Unpublished Rep. P0594006 from Biological Test Center, Irvine, CA, undertaken for the PBO Task Force, Washington, DC. Skrinijaric-Spoljar, M., Matthew, H. B., Engel, J. L., and Casida, J. E. (1971). Response of hepatic microsomal mixed-function oxidases to various types of insecticide chemical synergists administered to mice. Bioehem. Pharmacol. 20, 1607-1618. Squire, R A., and Levitt, M. H. (1975). Report of a workshop on classification of specific hepatocellular lesions in rats. Cancer Res. 35, 3214-3223. Suzuki, H., and Suzuki, N. (1995). Piperonyl butoxide mutagenicity in human RSa cells. Mutation Res. 344, 27-30. Takahashi, 0., Oishi, T., Fujitani, T, Tanaka, T, and Yoneyama, M. (1994a). Chronic toxicity studies of piperonyl butoxide in F344 rats: Induction of hepatocellular carcinoma. Fundam. App!. Toxieol. 22,293-303. Takahashi, 0., Oishi, T, Fujitani, T, Tanaka, T, and Yoneyama, M. (1994b). PBO induces hepatocellular carcinoma in CDl mice. Arch. Toxieol. 68, 467-469. Tanaka, T (1992). Effects of piperonyl butoxide on Fl generation mice. Toxieo!. Lett. 60, 83-90. Tanaka, T. (1993). Behavioral effects of piperonyl butoxide in male mice. Toxieo!' Lett. 69, 155-161. Tanaka, T, Takahashi, 0., and Oishi, S. (1992). Reproductive and neurobehavioral effects in three-generation toxicity study of piperonyl butoxide administered to mice. Food Chem. Toxieo!. 30, 1015-1019. Tanaka, T., Fujitani, T, Takahashi, 0., and Oishi, S. (1994). Developmental toxicity evaluation of piperonyl butoxide in CD-l mice. Toxicol. Lett. 71, 123-129. Tatematsu, M., Ozaki, K, Mutai, M., Shichino, Y., Furihata, c., and Ito, N. (1990). Enhancing effects of various gastric carcinogens on development of pepsinogen-altered pyloric glands in rats. Carcinogenesis 11(11), 19751978. Tayama, S. (1996). Cytogenic effects of PBO and safrole in CHO-Kl cells. Mutation Res. 368, 249-260. U.S. Environmental Protection Agency (U.S. EPA) (1988). "Standard Evaluation Procedure for Inhalation Studies." USEPA, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA) (1995a). "List of Chemicals Evaluated for Carcinogenic Potential." Office of Pesticide Programs, Washington, DC. U.S. Environmental Protection Agency (U.S. EPA) (1995b). "Pesticide Handlers Exposure Database (PHED) Evaluation Guidance," PHED VI. 1. Occupational and Residential Exposure Branch, Office of Pesticide Programs, Washington, DC. U.S. National Cancer Institute (1979). "Bioassay of Piperonyl Butoxide for Possible Carcinogenicity." DHEW PubL 79-1375, Bethesda, MD, U.S. Department of Health, Education and Welfare. Van Miller, J. P., Lalich, J. J., and Alien, J. R. (1977). Increased incidence of neoplasms in rats exposed to low levels of 2,3,7,8-tetrachlorodibenzo-rhodioxin. Chemosphere 6(9), 537-544. Wachs, H. (1947). Synergistic insecticides. Science 105, 397-401. Waf, D. J., and Short, C. R. (1971). Induction of hepatic microsomal hydroxylating enzymes by technical piperonyJ butoxide and some of its analogues. Toxieo!. App!. Pharmaeo!' 19,54-61. Wagstaff, D. J., and Short, C. R (1971). Induction of hepatic microsomal hydroxylating enzymes by technical piperonyl butoxide and some of its analogues. Toxieol. Appl. Pharmaeol. 19,54-61. Watanabe, T., Manabe, S., Ohashi, Y., Okamiya, H., Onodera, H., and Mitsumori, K. (1998). Comparison of the induction profile of hepatic drugmetabolizing enzymes between piperonyl butoxide and phenobarbital in rats. 1. Toxieo!. Pathol. 11, 1-10. Wellcome Environmental Health Unit (1991). Users ofPBO. Wester, R. c., Bucks, D. A. w., and Maibach, H. I. (1994). Human in vivo percutaneous absorption of pyretbrin and piperonyl butoxide. Food Chem. Toxieo!. 32,51-53.
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PBO Safety
White, T. J., Goodman, D., Shulgin, A. T., Castagnoli, N., Jr., Lee, R, and Petrakis, N. L. (1977). Mutagenic activity of some centrally active aromatic amines in Salmonella typhimurium. Mutat. Res. 56, 199-202. Whysner, J., Ross, P. M., and Williams, G. M. (1996). Phenobarbital mechanistic data and risk assessment: Enzyme induction, enhanced cell proliferation, and tumor promotion. Pharmacal. Ther. 71, 153-191. Wilson, D. M., Goldsworthy, T. L., Popp, J. A., and Butterworth, B. E. (1992). Evaluation of genotoxicity, pathological lesions, and cell proliferation in
livers of rats and mice treated with furan. Environ. Mol. Mutagen. 19, 209222. Wintersteiger, R, and Juan, H. (1991). "Resorption Study of Tyrason after Dermal Application (Study Performed on Healthy Subjects)." J.S.w.Experimental Research, Studie Analytik. 01191, p. 1. Yamamoto,L (1973). Mode of action of synergists in enhancing the insecticidal activity of pyrethrum and pyrethroids. In "Pyrethrum-Natural Insecticide," pp. 195-210. Academic Press, LondonlNew York.
CHAPTER
65 Pentachlorophenol Gay Goodman Human Health Risk Resources, Inc.
65.1 IDENTITY, PROPERTIES, AND USES
See Figure 65.1 for molecular structure.
The compound consists of needlelike crystals. PCP of 98% purity has been described as cream-colored, while technicalgrade PCP is pale brown in color (NTP, 1989). PCP is almost insoluble in water (8 mg in 100 ml), freely soluble in alcohol or ether, soluble in benzene, and slightly soluble in cold petroleum ether (Budavari, 1996). Sodium pentachlorophenate (NaPCP), the sodium salt of the anion, has empirical formula C6ClSO-Na+ and molecular weight 288.32. NaPCP is freely soluble in water. PCP has a pungent odor when heated. The odor threshold in humans is approximately 1.6 mg/l (WHO, 1987).
65.1.3 SYNONYMS
65.1.5 HISTORY, FORMULATIONS, AND USES
Pentachlorophenol, commonly abbreviated as PCP, is also known as penta, chlorophen, penchlorol, pentachlorofenol, and pentachlorofenolo. Current and former trade names include Dowicide 7, Dowicide EC-7, Dow Pentachlorophenol DP-2 Antimicrobial, Fungifen, Fungol, Glazd Penta, Permacide, Permagard, Permasa, Permatox, Permite, Santophen, Term-i-Trol, Thompson's Wood Fix, Weedone, and Witophen P.
PCP, in common with other chlorophenols, has a broad range of biocidal activity. In particular, PCP has been found to be effective as an algicide, bactericide, fungicide, herbicide, insecticide, and molluscicide (WHO, 1987). Collectively, PCP and its sodium salt previously constituted one of the most heavily used pesticides in the United States. Net production of the three U.S. manufacturers active in 1980 was 30,600 tons (Jones, 1981). Usage patterns tabulated around that time indicate that 95-98% of North American PCP production (of which 94% originated in the United States) was employed directly or indirectly in wood treatment (Economist Intelligence Unit, 1981). Wood preservation uses of PCP and NaPCP included commercial wood treatment, fence-post treatment, paint treatment, and sapstain control in pressboard and other woodderived materials (WHO, 1987).
65.1.1 CHEMICAL NAME Pentachlorophenol is the chemical name.
65.1.2 STRUCTURE
65.1.4 PHYSICAL AND CHEMICAL PROPERTIES Pentachlorophenol (PCP) is a chlorinated derivative of phenol with empirical formula C6CISOH and molecular weight 266.34. OH
65.1.5.1 U.S. Federal Regulation of Pesticidal Use Cl
Cl
o
Cl
Cl
Figure 65.1
Molecular structure of pentachlorophenol.
Handbook of Pesticide Toxicology Volume 2. Agents
Pesticidal use of PCP is regulated at the federal level under the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA). PCP was designated a Restricted Use pesticide on July 13, 1984 by the Office of Pesticide Programs (U.S. EPA, 1984a). Most uses as an herbicide, antimicrobial agent (e.g., in cooling towers), defoliant, disinfectant, and molluscicide (e.g., in marine paint) were discontinued at that time. It is permitted to use PCP as a biocidal agent on wood, but not on wood to be used for log homes or the interiors of buildings. It is also permitted to use PCP as a biocide in oil field flood waters and in pulp and
1481
Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
1482
CHAPTER 65
Pentachlorophenol
paper mill solutions. The label is required to state that PCP can only be sold to and used by certified pesticide applicators who must wear specific items of protective clothing and take specific handling precautions. The label also must state that application to logs for use in the construction of log homes is explicitly prohibited (U.S. EPA, 1986). 65.1.5.2 U.S. Federal Regulation of Toxic Contaminants
Levels of the most toxic contaminants of PCP and PCP salts are regulated by U.S. federal statute. Both hexachlorodibenzo-pdioxin (HxCDD) and hexachlorobenzene (HCB) are considered by the U.S. EPA to pose potential carcinogenic, teratogenic, and fetotoxic risks. As of February 2, 1989, the maximum allowable level of HxCDD is 4 ppm in any batch and the maximum allowable monthly average is 2 ppm (U.S. EPA, 1987). Registrants are instructed that the method used to lower HxCDD is not to increase concentrations of HCB and polychlorodibenzofurans (pCDFs) "above the levels in products marketed at the time of publication of this Notice" (U.S. EPA, 1984a); for HCB the maximum level was later specified as 75 ppm (U.S. EPA, 1987). The other contaminant specifically regulated in the earlier ruling was the potent rodent carcinogen 2,3,7,8tetrachlorodibenzo-p-dioxin (TCDD), which was required to be below detection by an acceptable method (U.S. EPA, 1984a, b). The later ruling established that the detection limit for TCDD shall be no higher than 1 ppb (U.S. EPA, 1987). 65.1.5.3 U.S. Federal Drinking Water Standards and Guidelines
Pursuant to the requirements of the Safe Drinking Water Act, the U.S. EPA established the level ofPCP permissible in drinking water: the maximum contaminant level (MCL) was set at 0.001 mg/l (U.S. EPA, 1991a). Based upon potential carcinogenicity, the maximum contaminant level goal (MCLG) was set at zero mg/l (U.S. EPA, 1991b).
65.2 PHARMACOKINETICS 65.2.1 ABSORPTION
PCP is readily absorbed by oral, inhalation, and dermal routes of exposure. 65.2.1.1 Oral
Absorption of PCP or NaPCP by the oral route has been studied in the mouse (Reigner et aI., 1992a), rat (Braun et aI., 1977; Reigner et aI., 1991; Yuan et aI., 1994), rhesus monkey (Braun and Sauerhoff, 1976), and human (Braun et aI., 1979). The time for peak plasma levels to be attained following an oral dose of PCP has been found to be comparable in rodents (1.5-6 hr) and humans (4 hr) but longer in monkeys (12-24 hr). The oral absorption efficiency was found to be greater than 90% in rats,
monkeys, and humans for doses in the range 0.1-15 mg/kg. Approximately 4-20% (dependent on dose and species) of a single ingested dose of PCP is excreted in the feces, while essentially all of the remainder appears in the urine. The slow time course of excretion along with the completeness of absorption into the bloodstream are evidence that enterohepatic recirculation plays an important role in the disposition of PCP in all species examined. 65.2.1.2 Inhalation
Absorption of an aerosol of NaPCP was studied in rats by Hoben et al. (1976a); approximately 80% of the inhaled dose (6 mg/kg over 20 minutes) was excreted unmetabolized in urine during the 72-hr period following exposure. However, actual absorption must have been greater than 80% because urinary excretion of the major metabolite in rodents was not quantified (see Section 65.2.3). An estimate of the amount of inhaled PCP absorbed by humans under conditions in which dermal exposure was expected to be insignificant was obtained in two volunteers with no previous occupational exposure to PCP (Casarett et aI., 1969). The two subjects were exposed to mean ambient PCP concentrations of 230 and 432 mg/m 3 PCP, respectively, for 45 minutes in a wood-treatment plant during brush application of PCP. Calculated exposure doses were 91 and 147 !-lg. Based upon chemical analysis ofPCP in the 7-day urine, respiratory tract absorption was calculated as 88% for the volunteer exposed to the lower concentration. A less complete accounting of absorption (76%) was available for the other volunteer, as urinary PCP levels were still well above baseline at the end of the 5-day collection period. 65.2.1.3 Dermal
Wester et al. (1993) measured dermal absorption of PCP in four female rhesus monkeys. 14C-Iabeled PCP (98.6% pure) was prepared in either acetone or premoistened soil and applied topically to an area of shaved abdominal skin. The PCP doses applied were 0.8 !-lg/cm2 for the acetone sample and 0.7 !-lg/cm2 for the soil sample. Urine was collected for 14 days. The percentage of the applied dose absorbed percutaneously was determined to be 29 ± 6% of the acetone sample and 24 ± 6% of the soil sample. 65.2.2 DISTRIBUTION AND ELIMINATION
The kinetics of removal from plasma are summarized in Table 65.1, while Table 65.2 summarizes the kinetics of excretion via the urine and feces. There is good agreement in the published literature on plasma half-lives of PCP in various experimental animals. Following oral or intravenous (i.v.) administration, mean half-lives of 5-6 hr in mice, 2-11 hr in rats, and 72-84 hr in monkeys have been calculated based on a first-order model representing the major portion of plasma PCP. Urinary excretion rates are similar to the corresponding plasma distribution rates, with estimated mean half-lives of 13 hr in rats
65.2 Pharmacokinetics
1483
Table 65.1 Kinetics of PCP Removal from Plasma in Mice, Rats, Monkeys, and Humans Following Single-Dose Exposurea Dose
Collection
Analytical
Mean half-life
Study
Route
Species
Sex
No.
(mg/kg)
period
method
in plasma (hr)
Reigner et aI., 1992a
oral
mouse
M
6
15
36 hr
Chem.
5.8 b
6 2d
15
36 hr
10
Chem. 14C
6.9, (24)",e
10
Reigner et aI., 1992a
i.v.
mouse
M
Braun et aI., 1977
oral
rat
M
Braun et aI., 1977
oral
rat
F
2d
Reigner et aI., 1991
oral
rat
M
5
Reigner et aI., 1991
i.v.
rat
M
5
Reigner et aI., 1991
i.v.
rat
M
Reigner et aI., 1991
i.v.
rat
M
Yuan et aI., 1994
oral
rat
M
Yuan et aI., 1994
oral
rat
6 days
5.2b
14C
11, (30)",e
2.5
48 hr
Chem.
7.5 b
2.5
6 days 48 hr
Chem.
(0.7),7.1'
20
96 hr
4.1, (36)
20
96 hr
Chem. 14C
40hr
Chem.
8.6b 6.3 b <3 1 ,5.6e
M
3d 3d
38
60hr
Chem.
5
9.5
4.5, (45)
Yuan et aI., 1994
i.v.
rat
M
3d
20hr
Chem.
Yuan et aI., 1994
i.v.
rat
F
5
20 hr
Chem.
<41 ,9.S'
Meerman et aI., 1983
i.v.
rat
M
3d 2d
10.7
36 hr
2.2,7.2e
Braun and Sauerhoff, 1976
n.g.
monkey
M
3
10
7 days
Chem. 14C/Chem.
Braun and Sauerhoff, 1976
n.g.
monkey
F
3
10
7 days
14C/Chem.
84b
Braun et aI., 1979
oral
human
M
4
6 days
Chem.
30b
0.1
nb
aAbbreviations: 14C, radioactive counts; Chem., chemical analysis; i.v., intravenous; n.g., nasogastric. bMonophasic model. cValue estimated using linear extrapolation from data in Fig. 2 of the citation. dNumber of animals sacrificed or sampled per time point. eBiphasic model. Half-lives accounting for only a minor portion of PCP are in parenthesis. 1Upper limit of initial-phase half-life estimated from inspection of data in Fig. 1 of the citation.
Table 65.2 Kinetics of PCP Excretion in Rats, Monkeys, and Humans Following Single-Dose Exposurea
Study
Route
Species
Sex
No.
Mean
Mean elimination half-life (hr)
Dose
Collection
Analytical
elimination half-lifeb
(mglkg)
period
method
(hr) 17.4, (40.2)"
urinary
Braun et aI., 1977
oral
rat
M
3
10
9 days
14C
Braun et aI., 1977
oral
rat
F
3
10
9 days
14C
13.4, (32.5)"
Braun et al., 1977
oral
rat
M
3
100
8 days
14C
12.8, (121)C
Braun et aI., 1977
oral
rat
F
3
100
8 days
14C
27.2d
Braun et aI., 1977
oral
rat
n.s.
3
100
8 days
14C
13, (31)"·e
Braun and Sauerhoff, 1976
n.g.
monkey
M
3
10
7 days
14C/Chem.
41 d
10
7 days
14C/Chem.
92d
6 days
Chem. 13C
432d 480d
Braun and Sauerhoff, 1976
n.g.
monkey
F
3
Braun et aI., 1979
oral
human
M
4
Uhl et aI., 1986
oral
human
M
Uhl et aI., 1986
oral
human
M
I
0.31
Uhl et aI., 1986
oral
human
M
3
0.055-0.15
Young and Haley, 1978
oral
human
M
0.1 0.016
::::24008
53 days 70 days
Chem.
6-14 days
Chem.
7 days
Chem.
33 d
116d
:s144d ,J 128d
aAbbreviations: 13C, isotopic substitution; 14C, radioactive counts; Chem., chemical analysis; i.v., intravenous; n.g., nasogastric; n.s., not specified. bCombined urinary and fecal excretion. cBiphasic model. Half-lives accounting for only a minor portion of PCP are in parenthesis. dMonophasic model. eValues estimated using linear extrapolation from data in Fig. 4 of the citation. 1Estimated from Fig. I of the citation. 8 Accidental poisoning case study.
1484
CHAPTER 65
Pentachlorophenol
and 41-92 hr in monkeys. In rats, excretion by combined urinary and feca1 routes has also been measured; the major portion of dose is excreted with an estimated mean half-life of 13-27 hr. The single human study to examine the rate of PCP elimination from plasma reported a mean half-life of 30 hr, nearly identical to the mean half-life for urinary excretion found in the same study (33 hr). Other human studies of urinary excretion reported much longer half-lives (128--480 hr). The disparity in reported values is unexplained.
of Juhl et al. (1985), who found evidence of PCP metabolism to TCHQ in a microsomal extract (S-9 fraction) of human liver (from a 61-year-old woman) and compared the time course of metabolite formation in the human liver S-9 to that in rat liver S-9. The rate of formation of TCHQ in the rat and human S-9 fractions during the first half-hour of incubation with the same (unspecified) PCP concentration can be estimated from Fig. 1 of Juhl et al. (1985): approximately 0.93 pmol/mg protein/min in human liver S-9 and 1.7 pmol/mg protein/min in rat liver S-9. Without information as to the concentration of PCP in the incubation mixtures, it is impossible to judge the relevance of this 65.2.3 METABOLISM in vitro result to human exposures in vivo. The inducible cytochrome P450 family of enzymes (the An across-study comparison of urinary excretion profiles in- mixed-function oxygenase system) plays an important role in dicates that the disposition of PCP in rodents is qualitatively the Phase I biotransformation of a broad spectrum of endogeand quantitatively dissimilar to its disposition in monkeys and nous compounds and xenobiotics (Sipes and Gandolfi, 1991). humans. In mice and rats, a substantial, dose-dependent frac- Rats and humans differ substantially with respect to cytochrome tion (16--48%) of ingested PCP is excreted in the urine as P450 sUbtypes (Paine, 1995). The human P450 3A isozymes tetrachloro-1,4-hydroquinone (TCHQ). By contrast, every ex- function as mediators of the hydroxylation of steroids and varperimental study entailing administration of PCP to humans or ious drugs. Human P450 3A3/4 and 3A7 variants are found monkeys has failed to find any evidence for the metabolism of in both adult and fetal liver (Hakkola et al., 1994). The gene PCP to TCHQ. In two male and two female rhesus monkeys for the human P450 3A4 variant has been expressed in a given [14C]PCP by nasogastric intubation at a single dose of strain of the yeast Saccharomyces cerevisiae, in which it was 10 mg/kg, all recovered urinary radioactivity occurred as PCP able to catalyze the oxidative transformation of an extremely (Braun and Sauerhoff, 1976). Using gas chromatography/mass diverse group of molecules (Brian et aI., 1990). Mehmood spectrometry (GCIMS), Braun et al. (1979) found no trace of et al. (1996) used this yeast system to express human P450 TCHQ in the urine of four human male volunteers exposed 3A4 and found that in both intact yeast cultures and microorally to non-radiolabeled PCP at a single dose of 0.1 mg/kg. somal extracts, PCP is transformed to TCHQ in a process Using a more sensitive method, GCIMS detection of 13C_ dependent upon the presence of the cytochrome, oxygen, and labelled compounds, Uhl et al. (1986) failed to find detectable NADPH. The affinity constant (Ka) for PCP was reported to levels of [13CJ-labeled TCHQ, 2,3,4,5-tetrachlorophenol, or be 85 !1M, although data used in the derivation of this value 2,3,4,6-tetrachlorophenol in the urine of two human volun- were not given. In microsomal extracts, incubation for 30 min teers exposed orally to [13C]PCP at a single dose of 0.98 or in the presence of 113 ll-M PCP produced 8 pmol TCHQ/nmol 2.4 mg/kg. P450/min. Nevertheless, the absence of PCP metabolism in humans Taken together, these data suggest that humans, like roremains controversial. Using GCIMS, Ahlborg et al. (1974) dents, possess the enzymatic capability to metabolize pep to identified (without quantification) both TCHQ and PCP in the TCHQ, but in both the human and the rhesus monkey (the only urine of two occupationally exposed male pesticide applicators. nonhuman primate species tested) this enzymatic pathway is Because complete exposure profiles for the applicators were unimportant, at least when PCP is given in a single dose as not obtained, one cannot rule out the possibility of concurrent large as 2.4 mg/kg (human) or 10 mg/kg (monkey). Reigner exposure to the pesticide lindane (y-hexachlorocyclohexane). et al. (1992b) suggested that the failure of some investigators to It is known that lindane undergoes hydroxylation followed by find TCHQ in the urine of humans exposed to PCP may reflect aromatization (Gopalaswamy and Aiyar, 1986) to form 2,3,4,6- the instability of this compound in urine. However, it is diffiand 2,3,5,6-tetrachlorophenol as major metabolites in the rat cult to imagine why the breakdown product(s) would not then (Engst et aI., 1976), while TCHQ is the primary metabolite be detected, especially in the [13C]PCP experiment of Uhl et al. of 2,3,5,6-tetrachlorophenol administered to rats (Ahlborg and (1986). Larsson, 1978). This pathway does not entail formation of PCP. The question of whether humans metabolize PCP to TCHQ Furthermore, tetrachlorophenol is a contaminant of the com- to any significant extent is more than academic: TCHQ has been mercial product (i.e., technical-grade PCP). The percentage demonstrated to be genotoxic in numerous short-term tests (intetrachlorophenol in technical-grade PCP varies over time and cluding those for mutagenicity), whereas the data base on the from product to product, but as examples, the two PCP formula- genotoxicity of PCP consists of essentially negative or equivtions tested in a mouse carcinogenicity bioassay by the National ocal results (see Section 65.6). If, at a given PCP dose level, Toxicology Program contained tetrachlorophenol at levels of TCHQ is formed to a significant extent in mice but not humans, then very likely the carcinogenic potency of PCP in mice is not 3.8% and 9.4%, respectively (NTP, 1989). Another finding in apparent conflict with the in vivo obser- relevant to humans. Given the importance of this question, it vations of no PCP metabolism to TCHQ is the in vitro result is surprising how little attention has been focused on resolving
65.2 Pharmacokinetics
1485
Table 65.3 Relative Amounts of Conjugated and Nonconjugated PCP and TCHQ in Urine Following Single-Dose Exposure to PCpa Percent of total urinary recoveryb
Collection
Dose
period
(mg/kg)
Study
Species
Sex
Route
Jakobson and Yllner, 1971
mouse
F
i.p.
24hr
7.4--8.2
Jakobson and Yllner, 1971
mouse
F
i.p.
24hr
15-19
J akobson and Yllner, 1971
mouse
F
i.p.
24hr
Ahlborg et ai., 1974
mouse
n.S.
i.p.
24hr
Reigner et ai., 1992a
mouse
M
gav.
48 hr 24hr
PCP
PCP-c
TCHQ
TCHQ-c 44%C 33 c •d
54%C
54%
12%
36-37
50%
20%
10-25
41%
13%
24%
22%
15
8%
51%
5%
47%
10
60%
100
75%
9% 100%C
48%C,d
Ahlborg et al., 1978
rat
n.s.
i.p.
Braun et aI., 1977
rat
MIP
gav.
8 days
Braun and Sauerhoff, 1976
monkey
MIF
n.g.
7-15 days
10
Braun et ai., 1979
human
M
oral
7 days
0.1
86%
14%
n.d.
n.d.
Uhl et aI., 1986
human
M
oral
0.31
71%
29%
n.d.
n.d.
24 hr"
9-16%
22%
7% 16%C
n.d. c
gav., gavage; i.p., intraperitoneal; n.d., none detected; n.g., nasogastric; n.s., not specified; PCP-c, PCP conjugate; TCHQ-c, TCHQ conjugate. bMean values except as indicated. cThe conjugated and nonconjugated forms were not differentiated. dValue for the single animal tested. eSingle sample collected 24 hr after dosing. a Abbreviations:
the discrepancy between the in vivo and in vitro findings. An attempt should be made to replicate the in vitro results of Juhl et al. (1985), at the same time improving upon the study design. It would be useful to learn how the rate of TCHQ formation in intact rodent and human liver cells as well as microsomes depends on PCP, particularly at concentrations expected to occur in blood following administration of PCP at doses tested in the pharmacokinetic studies. It would also be helpful to know whether [13C]TCHQ or its metabolites are excreted by humans or nonhuman primates during multiday exposure to [13C]PCP. The discovery that horseradish peroxidase can catalyze the hydroperoxide-dependent oxidation of PCP to TCHQ in vitro has led to speculation that mammalian peroxidases might catalyze this transformation in vivo as well (Samokyszyn et aI., 1995). Whether this possible pathway is relevant to interspecies differences in PCP metabolism remains to be discovered. The potential influence of exposure route on the metabolism of PCP has not been studied methodically. However, there is no evidence to suggest route dependence of either rates or pathways. Glucuronide and sulfide conjugates of PCP and TCHQ contribute to the total urinary excretion, as revealed by complete hydrolysis (Ahlborg et aI., 1978; Edgerton and Moseman, 1979). Experimentally observed urinary excretion profiles of free and conjugated PCP and TCHQ following single-dose administration of PCP are given in Table 65.3. 65.2.4 PLASMA PROTEIN BINDING
In plasma isolated from Sprague-Dawley (SD) rats, PCP was found to have high- and low-affinity binding constants of approximately 106 and 104 M-I, respectively, suggesting that binding is strong enough to influence distribution and
metabolism (Braun et aI., 1977). In an in vivo study in the same rat strain, 81 % of plasma PCP was found to be bound to protein (G6mez-Catahin et aI., 1991) Hoben et al. (1976b) reported that the ratio of bound PCP to albumin (mol/mol) was 1.3 for human plasma and 0.86 for plasma from an unspecified rat strain; the investigators suggested that this difference, which reflects differential binding to nonalbumin sites, may contribute to the longer retention time of PCP in human plasma. 65.2.5 PROTEIN ADDUCTS OF REACTIVE METABOLITES
Waidyanatha et al. (1996) found that PCP-derived adducts of hemoglobin and albumin occurred in a dose-dependent manner in SD rats given a single gavage dose of PCP at 5 to 40 mg/kg. The investigators demonstrated that the adducts resulted from reaction of protein binding sites with the quinones tetrachloro-l,4benzoquinone (TC-l,4-BQ) and tetrachloro-l,2-benzoquinone (TC-l,2-BQ) and the semiquinones tetrachloro-l ,4-benzosemiquinone (TC-l,4-BSQ) and tetrachloro-l,2-benzosemiquinone (TC-l,2-BSQ), all of which are formed by metabolism of PCP via TCHQ or tetrachlorocatechol (TCC): PCP
~
TCHQ
PCP
~
TCC
~
~
TC-l,4-BSQ
TC-l,2-BSQ
~
~
TC-l,4-BQ
TC-l,2-BQ
In lifetime carcinogenicity bioassays, PCP produces liver tumors in mice but not rats (see Section 65.5). Lin et al. (1999) measured the in vivo formation of PCP-derived benzoquinone and benzosemiquinone adducts of liver cytosolic and nuclear proteins in B6C3Fr mice and SD rats and attempted to interpret the results in light of the known interspecies difference in liver tumorigenicity. Rats and mice were given a single gavage dose of PCP at 5 to 40 mg/kg and sacrificed 24 hr later. The extent
1486
CHAPTER 65
Pentachlorophenol
and type of adduct formation apparently differed between mice and rats. The primary differences appear to have been: (a) rats formed adducts of TC-1,2-BSQ and TC-1,4-BSQ in both cytoso1ic and nuclear protein whereas mice formed the cytosolic adducts (at an order of magnitude lower rate) but not the nuclear adducts, and (b) mice formed mono-S-substituted adducts of TC-1,2-BQ in both cytosolic and nuclear protein whereas rats formed neither. The authors argued that adduct formation in mice was linear in dose whereas that in rats was sublinear, but in fact, one or more minor (benzosemiquinone) adducts in both species demonstrated sublinearity while the major (benzoquinone) adducts were about equally linear in dose in both species. Based on the results described, one might tentatively hypothesize that PCP is able to produce liver tumors in B6C3Fl mice but not SD rats (Section 65.5) because the mice form mono-S-substituted TC-1 ,2-BQ protein adducts while the rats do not. However, in the absence of plausible data showing that the ability to form this particular adduct above all others determines PCP tumorigenicity in mice, it is tempting to argue that the results of Lin et al. (1999) provide evidence against the hypothesis that protein adduct formation is related to the liver tumorigenicity of PCP.
44-min exposures to an aerosol of NaPCP containing PCP at approximately 1 ppm resulted in 33 to 83% mortality. Based on the assumption that the inhalation volume was 80 ml/min, the corresponding range of PCP doses was calculated as 10 to 14.5 mg/kg and the LD50 as 12 mg/kg (Hoben et aI., 1976c). Oral LD50 values in rats for an unspecified grade of PCP were reported to be 146 mg/kg for males and 175 mg/kg for females (Gaines, 1969). Renner et al. (1986) tested purified PCP in mice and found oral and intraperitoneal (i.p.) LDso values of approximately 130 and 60 mg/kg, respectively, in both sexes. Similar oral LDso values for purified PCP in mice (177 mg/kg for males and 117 mg/kg for females) were reported by Borzel1eca et al. (1985). In an earlier study in female mice, Ahlborg and Larsson (1978) tested an unspecified grade of PCP delivered in 40% ethanol and found oral and i.p. LDso values of 74 and 32 mg/kg, respectively. See Gasiewicz (1991) for a summary of dermal, subcutaneous, and earlier oral and i.p. studies and Goodman et al. (1998) for descriptions of several unpublished studies.
65.4 SUB CHRONIC TOXICITY
65.3 ACUTE TOXICITY
The results of subchronic toxicity studies of PCP in mice and rats are summarized in Table 65.4.
The acute toxicity of PCP appears to be greatest when the route of exposure is inhalation. In a study in male rats, 28- to
Table 65.4 Subchronic Toxicity of PCP in Dietary Exposure Studies in Mice and Ratsa Strain/species/
LOEL
PCP
Study
sex
Duration
grade
Effects at LOELb
NTP,1989c
B6C3F 1 Mouse,
30 days
TGC
30 days
EC-7
30 days
AG
t serum cholesterol (F); t liver porphyrins (M); t liver lesions d (MIF); t platelets (MIF) t serum cholesterol (F); t liver porphyrins (M); 1. reticulocytes (F); t y -globulin (F) t liver porphyrins (M); t liver lesions d (F)
TGC, DP-2
MIF
NTP,1989 c
B6C3F 1 Mouse, MIF
NTP,1989c
B6C3F 1 Mouse,
NOEL
(mg/kg-day) 102
2.8
24
4.3
27
4.1
Liver lesionsd (MIF); urinary bladder changes (MIF);
43
None
immune suppression (MIF) Liver lesionsd (MIF); urinary bladder changes (MIF)
54
None
Liver lesionsd ;
53
None
MIF
NTP,1989c
B6C3Fl Mouse,
6 months
MIF
NTP,1989c
B6C3F] Mouse,
6 months
Renner et al., 1987
Sprague-Dawley
EC-7, AG
MIF
28 days
AG
Rat, F NTP,1999
F344 Rat,
t
relative liver and kidney weights;
mildanemia 28 days
AG
t
absolute and relative liver weights (F)
20
None
27 weeks
AG
Liver cell hypertrophy, necrosis (minimal),
71
None
MIF
Kurtz and Hejtmancik, 1993 e
F344 Rat, MIF
and enzyme changes (MIF);
1. body weight gain (MIF)
aAbbreviations: AG, analytical-grade PCP; LOEL, lowest-observed-effect level; NOEL, no-observed-effect level; TGC, technical-grade composite. bS ex affected at the LOEL is in parenthesis. CWith additional data provided by the NTP. dLiver lesions included cytomegaly, karyomegaly, nuclear atypia, hepatocellular degeneration, and necrosis. 'Interim sacrifice of the 2-year bioassay described in NTP (1999).
65.4 Subchronic Toxicity
65.4.1 MOUSE The NTP (1989) conducted 30-day and 6-month dietary exposure studies of PCP in B6C3Fl mice. Three PCP formulations were tested in the 30-day study: analytical-grade (AG-PCP), a partially purified product (EC-7), and a technical-grade composite (TGC-PCP). Two formulations were tested in the 6-month study: AG-PCP and a technical-grade product (DP-2). All formulations produced liver lesions in both sexes (NTP, 1989). Hematologic changes (altered platelet or reticulocyte counts) developed upon exposure to TGC-PCP and EC-7 but not AG-PCP, suggesting that one or more contaminants were intrinsic to these responses. In the 30-day study, no-ob servedeffect levels (NOELs) in male and female mice were 2.8, 4.3, and 4.1 mg/kg-day for TGC-PCP, EC-7, and AG-PCP, respectively. In the 6-month study, lowest-observed-effect levels (LOELs) in male and female mice were 43 mg/kg-d for TGCPCP and DP-2 and 54 mg/kg-d for EC-7 and AG-PCP, the lowest doses tested. The major contaminants of the test substances EC-7, DP-2, and TGC-PCP were other chlorophen01s. For example, tetrachlorophenol comprised 9.4% of EC-7 and 3.8% of TGC-PCP. Hexachlorobenzene was also a contaminant of these three formulations, present in EC-7 and TGC-PCP at 65 and 50 ppm, respectively (NTP, 1989). The extent of contamination of the test substances with polychlorinated dibenzodioxin (pC DD) and pCDF compounds is shown in Table 65.5.
65.4.2 RAT In an NTP study in rats, dietary exposure to purified PCP for 27 weeks at 71 mg/kg-day, the only dose tested, produced liver leTable 65.5 Chlorinated Dibenzodioxin and Dibenzofuran Contaminants of PCP Formulations Tested by the NTP in Toxicity Studies in Mice a PCP formulation Contaminant
EC-7 (91% PCP)h
TGC (90% PCP)b
(91.6% PCPjC
AG (98.6% PCP)b,c
TCDD
<0.04ppm
n.d.
n.d.
n.d.
HxCDD
0.19ppm
HpCDD
0.53 ppm
296 ppm
0.69 ppm
1386 ppm
OCDD PCDF
n.d.
HxCDF
0.13 ppm
HpCDF
0.15 ppm
OCDF
n.d.
10.1 ppm
1.4ppm 9.9ppm
DP-2
0.59
0.28
n.d.
173 n.d. 13.0
88ppm
172
n.d.
43ppm
320
n.d.
"Information drawn from Table 3 ofNTP (1989). Abbreviations: AG, analytical grade; HpCDD, heptachlorodibenzo-p-dioxin; HpCDF, heptachlorodibenzofuran; HxCDD, hexachlorodibenzo-p-dioxin; HxCDF, hexachlorodibenzofuran; OCDD, octachlorodibenzo-p-dioxin; OCDF, octachlorodibenzofuran; PCDF, pentachlorodibenzofuran; TGC, technical-grade composite; n.d., nondetectable (detection limit not given). bUsed in the 30-day study. TCDD was nondetectable at the detection limit of 40ppb. CUsed in the 6-month study.
1487
sions in both sexes (Kurtz and Hejtmancik, 1993). Minimal to mild degeneration of hepatocytes was observed in both male and female rats given purified PCP for 28 days at doses of 40 mg/kg-day and above (NTP, 1999). Mild anemia and elevated kidney weights were also found in female rats in response to 28-day dietary exposure to purified PCP at the only dose tested, 53 mg/kg-day (Renner et aI., 1987). A subchronic exposure study of PCP, limited to histoiogic examination of the sciatic nerve, liver, and kidney, was conducted in Chile by Villena et al. (1992). The investigators reported that young male Wistar rats were exposed to analytical-grade PCP in drinking water at 0.3 mM (80 mg/l) for 60 days, 1.0 mM (270 mg/l) for 60 or 90 days, or 3.0 mM (800 mg/l) for 120 days. Assuming default values for drinking-water consumption and body weight of 0.032 l/day and 0.217 kg, respectively (U.S. EPA, 1988), the PCP doses can be estimated as 12,39, and 120 mg/kg-day. Presumably, animals in each exposure group were sacrificed at the end of their exposure period. The four exposure groups were compared to a single control group; it is not clear at what stage the control animals were sacrificed. In rats exposed to 1.0 mM (39 mg/kg-day) for 90 days or 3.0 mM (120 mg/kg-day) for 120 days, the investigators reported myelin degeneration of the sciatic nerve, hepatocellular degeneration, hepatic vascular congestion, glomerular congestion of the renal cortex, and turbid tumefaction of the renal proximal tubules. However, the reliability of this study is questionable in light of several flaws and insufficient provision of experimental details. One major problem with this study is that the solubility of PCP in water is only 20 mg/l at 30°C (IPCS, 1987), a factor of 4--40 less than the target doses. The investigators apparently did not measure the actual concentrations attained. It is conceivable that PCP precipitated out of solution and sank to the bottom of the water bottles, resulting in higher doses than intended. Measurement of food intake, body weight, and organ weight constitutes minimal insurance against exceeding the maximum tolerated dose; none of these parameters was reported. An additional concern is that standard laboratory practices do not appear to have been followed concerning quantification of the observed histologic effects. Still another problem is that control animals were not sacrificed at the same times as exposure groups, i.e., at 60, 90, and 120 days. Thus, age-related effects, unintentional crosscontamination with other toxic agents, or other laboratory error cannot be excluded. Because of these concerns, and principally because the doses must be considered unspecified, the study is not included in Table 65.4.
65.4.3 COW A 5-month ingestion experiment in yearling Holstein cows was performed by McConnell et al. (1980). The data are useful for discriminating the effects of PCP from those of its technicalgrade contaminants. Three heifers per group were given a diet containing one of the following four mixtures of AG and TG
1488
CHAPTER 65
Pentachlorophenol
PCP: 100% AG, 90% AG/lO% TG, 65% AG/35% TG, and 100% TG. A control group of three heifers received an untreated diet. The dose in all treatment groups was originally set at 20 mg/kg-day but after 42 days was reduced to 15 mg/kg-day because of concerns about decreased body-weight gain compared to the controls. The overall time-weighted average dose was 16.3 mg/kg-day in all treatment groups. The distribution of chlorinated nonphenolic contaminants in the TG-PCP was similar to that in the TGC-PCP used in the NTP mouse study (Table 65.5). The purity of the AG-PCP was 99%. Statistically significant decreases in serum thyroxine (2940%) and serum triiodothyronine (35-44%) occurred in all four PCP-treatment groups. These changes echo those seen by van Raaij et al. (1991 b) and lekat et al. (1994) in rats exposed to PCP acutely and subchronically, respectively (see Section 65.7.1). The proliferative response of peripheral blood lymphocytes (PBLs) to in vitro stimulation by T-cell mitogens was increased with increasing percentage TG-PCP and exposure duration. By contrast, in PBLs from cows treated with 100% AG-PCP there was a nonsignificant decrease in the proliferative response. This result is consistent with the report by Kerkvliet et al. (1982a) that in vivo T-cell activation in mice is unaffected by treatment with analytical-grade PCP. Dose-dependent, statistically significant perturbations in serum immunoglobulins were observed in the TG-PCP groups. In the 100% AG-PCP group the changes in immunoglobulin levels were in the same direction as in the TG-PCP groups, but smaller and not statistically significant. In all treatment groups there was a marked, highly significant depression of thymus weight, both absolute and relative to body weight. Macroscopic changes were observed in the spleens of some cows in the 35% and 100% TG-PCP groups; microscopic
examination revealed hyper- or hypoplastic foci in these animals. Relative liver weights were significantly elevated in all treatment groups, but absolute liver weights were significantly increased only in TG-PCP groups. Increased serum y-glutamyl transpeptidase was seen only in the TG-PCP groups. This is indicative of cholestasis, a condition generally associated with lesions of the liver or bile secretory system; histopathologic examination revealed damage to the liver, bile duct, and gall bladder in animals receiving 100% TG-PCP. Although increased amounts of smooth endoplasmic reticulum were found in the 100% AG-PCP group, no other microscopic signs of liver toxicity occurred in these animals. In the absence of other signs of toxicity, an increase in SER might reflect an adaptive response to the xenobiotic load.
65.5 CARCINOGENICITY AND CHRONIC TOXICITY Two rodent bioassays conducted by the NTP provide the most useful information on the carcinogenicity of PCP in these species. The NTP rat bioassay tested analytical-grade (99% pure) PCP, whereas the NTP mouse bioassay tested technicalgrade and partially purified PCP formulations. Given the fact that several of the impurities are known or suspected carcinogens, the absence of interstudy consistency in the materials tested makes it difficult to compare the underlying carcinogenicity of PCP in mice and rats. Tables 65.6 (mice) and 65.7 (rats) give the incidence rates of exposure-related neoplastic and non-neoplastic lesions observed in these two studies.
Table 65.6 Summary of the Carcinogenicity and Chronic Toxicity of PCP Fonnulations Tested in Dietary Exposure Studies in Mice a B6C3Fl Mice (NTP, 1989; McConnell eta!., 1991) Doses in feed
TOC
Males EC-7
TOC
Females EC-7
0, 100,200 ppm (0,18,35 mg/kg-day)
0, 100,200,600 ppm (0,18,37,118 mglkg-day)
0, 100, 200 ppm (0,17,35 mg/kg-day)
0, 100, 200, 600 ppm (0,17,34,114 mg/kg-day)
3%,22%,22%
3%,40%,27%,2%
0%,8%,4%
6%,2%, 11 %,35%
0%,22%,51%
3%,8%,44%,90%
0%,4%,2%
0%, 2%, 4%, 78%
Nonneoplastic lesions Adrenal gland medulla: Hyperplasia Neoplastic lesions Adrenal gland medulla: Benign pheochromocytoma Liver: Adenoma
16%,43%,69%
14%,27%,35%,65%
9%,16%,16%
3%,6%, 12%,63%
Carcinoma
6%,21%,25%
3%, 15%, 15%, 18%
0%,2%,2%
0%, 2%, 0%, 4%
Adenoma or Carcinoma
22%,55%,77%
17%,40%,44%,69%
9%,18%,18%
3%, 8%, 12%,65%
0%,4%,2%
0%, 8%, 4%, 6%
0%,6%,12%
0%,2%,6%, 16%
Liver or spleen: Hemangiosarcoma
aIncidence rates in bold were significantly elevated (p < 0.05) relative to the control group. Abbreviation: TOe, technical-grade composite.
65.5 Carcinogenicity and Chronic Toxicity
1489
Table 65.7 Summary of the Carcinogenicity and Chronic Toxicity of 99% Pure PCP in Dietary Exposure Studies in Ratsa Males (NTP, 1999; Chhabra et aI., 1999) Doses in feed
Females
I-Year Onl1- Year Off
F3441N Rats
2-Year Exposure
I-Year Onl1-Year Off
Stop-Exposure
2-Year Exposure
Stop-Exposure
0, 200, 400, 600 ppm
1000 ppm
0, 100, 200 ppm
0,100,200,600 ppm
(0, 10,20,30 mg/kg-day)
(60 mglkg-day)
(0, 17, 35 mg/kg-day)
(0, 17,34, 114 mg/kg-day)
No exposure-related increase at any site
Increased hepatocyte centrilobular hypertrophy and cytoplasmic vacuolization at 7-month sacrifice but not at terminal sacrifice
No exposure-related increase at any site
Increased hepatocyte centrilobular hypertrophy at 7 -month sacrifice but not at terminal sacrifice
Malignant mesothelioma
2%, 0%,4%, 0%
18%
0%,0%,0%
0%, 0%, 0%, 0%
Nasal squamous cell carcinoma
2%, 6%, 2%, 0%
10%
0%,0%,0%
0%, 0%, 0%, 0%
Nonneoplastic lesions
Neoplastic lesions
aIncidence rates in bold were significantly elevated (p < 0.05) relative to the control group.
65.5.1 TWO-YEAR EXPOSURE IN MICE
65.5.1.3 Neoplastic Lesions of the Adrenal Medulla
The NTP conducted a 2-year feed study of PCP in B6C3F1 mice (McConnell et aI., 1991; NTP, 1989). The technical-grade composite (TGC) and partially purified (EC-7) PCP formulations tested are described in Section 65.4.1. Fifty animals per sex per group received TGC-PCP at 100 or 200 ppm diet (18 or 35 mg/kg-day) or EC-7 at 100,200, or 600 ppm diet (18,36, or 116 mg/kg-day). Separate control groups for the two formulations each consisted of 35 animals of each sex.
In male mice, TGC-PCP and EC-7 produced similar, doserelated increases in benign pheochromocytoma, with significant elevation in animals receiving 100 ppm TGC-PCP or 200 ppm EC-7. In female mice, significant elevation of pheochromocytoma occurred only in the 600 ppm EC-7 group.
65.5.1.1 Neoplastic Lesions of the Liver
In male mice, both EC-7 and TGC-PCP produced dose-related increases in benign and malignant liver tumors (hepatocellular adenoma and carcinoma). There was a significantly (p < 0.05) elevated incidence of adenoma in males exposed to TGC-PCP at either dose or to EC-7 at 200 ppm or higher, while a significantly elevated incidence of carcinoma occurred in males treated with 200 ppm TGC-PCP or 600 ppm EC-7. In female mice, a significant elevation of adenoma occurred only in the 600 ppm EC-7 group. 65.5.1.2 Non-neoplastic Lesions of the Liver
A large proportion of male and female mice treated with either EC-7 or TGC-PCP developed dose-related, non-neoplastic liver and bile-duct lesions (including cytomegaly, multifocal proliferation of hematopoietic cells, diffuse chronic inflammation, and acute diffuse necrosis). In low-dose TGC-PCP males and females the incidence rates for acute diffuse necrosis were 87% and 90%, respectively, while in low-dose EC-7 males and females the incidence rates were 98% and 42%, respectively.
65.5.1.4 Non-neoplastic Lesions of the Adrenal Medulla
In male mice, statistically significant increases in hyperplasia of the adrenal medulla were produced at both doses of TGC-PCP and at the two lower doses (but not the high dose) of EC-7. Incidence rates were 0%, 40%, 27%, and 2% in the 0, 100,200, and 600 ppm EC-7 groups, respectively. The V-shaped doseresponse observed in EC-7 males reflects the existence of a morphologic continuum leading to benign pheochromocytoma, which increased sharply between 200 and 600 ppm in parallel with the decline in hyperplasia. In female mice, hyperplasia of the adrenal medulla was significantly elevated only in the group receiving 600 ppm EC-7. 65.5.1.5 Neoplastic Lesions of the Vascular System
The incidence of a malignant tumor of the blood vessels (hemangiosarcoma) was significantly elevated in female mice treated with 200 ppm TGC-PCP or 600 ppm EC-7. The investigators indicated that these tumors occurred mostly in the spleen while a few were found in the liver. 65.5.1.6 Non-neoplastic Lesions of the Vascular System
Diffuse hematopoietic cell proliferation in the red pulp of the spleen (extramedullary hematopoiesis) was increased in male
1490
CHAPTER 65
Pentachlorophenol
and female mice at both doses of TGC-PCP. Incidence rates in the control, 100 ppm, and 200 ppm groups were 17%, 65%, and 39% in males and 6%, 31 %, and 23% in females, respectively. The U-shaped dose-response observed in both males and females may be partially explained by the existence of a morphologic continuum leading to splenic hemangiosarcoma, but there were too few males with this tumor at the 200 ppm dose (6%) to account for the magnitude of the downturn. With respect to the argument that the liver toxicity attributable to pCDD or pCDF contaminants was responsible for the observed hepatocarcinogenicity, the investigators noted that while non-neoplastic lesions of identical morphology and approximately equal severity occurred in females treated with EC-7 at 200 and 600 ppm, liver neoplasms were significantly elevated only in the 600 ppm group. The investigators performed separate analyses for benign adenoma, malignant carcinoma, and the presence of adenoma or carcinoma. It is the combined adenoma/carcinoma category which provides the strongest statistical association with PCP exposure. The use of such a category is not unusual and rests upon histologic evidence that hepatocellular adenoma and carcinoma are part of a physiological continuum (Bannasch et aI., 1986). It is noteworthy that the incidence of adenoma/carcinoma in males on the 200 ppm EC-7 regime (44%) was lower than the rate in males receiving TGC-PCP at 200 ppm (77%). Similarly, the incidence rates for hemangiosarcoma in females in the 100 and 200 ppm EC7 dose groups were lower (2% and 6%) than in females given TGC-PCP at the same levels (6% and 12%). IfPCP alone were responsible for the neoplastic response in liver and blood vessels, one expects that the incidence rates at a given dose level of TGC-PCP and EC-7 would have been roughly the same because both formulations contain approximately 90% PCP. The fact that TGC-PCP was more potent suggests that the contaminants in TGC-PCP acted as co-carcinogens. Concerning the argument (noted above) that the liver tumorigenicity of the PCP formulations may have been secondary to toxicity, the Environmental Health Committee of U.S. EPA's Science Advisory Board was asked to examine the results of the study and consider this question, among other issues. The Committee recommended that "the observed dose-dependent increase in the incidence of hepatocellular carcinomas and adenomas be considered a valid indicator" of the carcinogenic potential of PCP. The Committee also pointed out that the mouse strain tested in the study, B6C3F I, is extremely sensitive to hepatocarcinogenesis and therefore recommended that exposure-related liver tumors be considered less relevant to human risk than the production of hemangiosarcomas (U.S. EPA SAB,1991). The incidence rates for benign pheochromocytoma in male mice were approximately the same in the 200 ppm TGC-PCP and EC-7 groups (51 % and 44%, respectively), while the rate in the 100 ppm TGC-PCP group was notably higher than in the 100 ppm EC-7 group (22% vs. 8%). These results suggest that the technical-grade impurities were key to the response at the 100 ppm dose but less important at the 200 ppm dose, when PCP played a partial role in the formation of pheochromocy-
tomas. In female mice, the only elevated pheochromocytoma incidence occurred in groups treated with 600 ppm EC-7 (78%), with a sharp increase over rates in the 100 and 200 ppm groups (4% in both). This was similar to the dose-response relationship observed for hepatocellular adenoma/carcinoma in the female EC-7 groups. The NTP concluded that there was clear evidence of carcinogenic activity in male B6C3FI mice exposed to TGC-PCP or EC-7, clear evidence ofcarcinogenic activity in female B6C3F I mice exposed to EC-7, and some evidence of carcinogenic activity in female B6C3FI mice exposed to TGC-PCP (NTP, 1989). 65.5.2 INITIATIONIPROMOTION IN MICE
Umemura et al. (1999) conducted an initiation/promotion study of 99% pure PCP in male B6C3Fl mice to investigate the mechanism by which PCP produces liver cancer. In the portion of the study devoted to evaluating the ability of PCP to act as a promoter, groups of 20 mice received PCP at 0, 300, or 600 ppm diet for 25 weeks following preinitiation with diethylnitrosamine (DEN). The incidence of animals bearing liver adenoma or carcinoma was significantly elevated in both the 300 ppm (10/15, p < 0.05) and 600 ppm (13/18, p < 0.01) PCP groups relative to DEN treatment alone (4/15). The liver tumor incidence was 0/19 in mice receiving PCP for 25 weeks without prior DEN treatment. In the portion of the study devoted to evaluating the ability of PCP to act as an initiator, groups of 20 mice were given PCP at 600 or 1200 ppm diet for 13 weeks followed by phenobarbital in drinking water at 500 ppm for 29 weeks. No liver tumors were observed. The study by Umemura et al. (1999) demonstrated that 99% pure PCP at 300 ppm diet has the ability to promote liver carcinogenesis in preinitiated male B6C3FI mice. The study also revealed that purified PCP at 1200 ppm diet for 13 weeks does not appear to initiate liver tumors. 65.5.3 TWO-YEAR EXPOSURE AND STOP-EXPOSURE IN RATS
The NTP conducted a 2-year, dietary-exposure study of 99% pure pep in F3441N rats (Chhabra et al., 1999; NTP, 1999). Rats on the standard regime were given PCP at 200, 400, or 600 ppmdiet (10, 20, or 30 mg/kg-day) for two years, while rats on the stop-exposure regime were given a diet containing PCP at 1000 ppm the first year and no PCP the second year (timeaveraged dose = 60 mg/kg-day). A total of 50 rats per sex per dose were on the standard regime while 60 rats per sex were on the stop-exposure regime. Ten rats of each sex in the stopexposure group were necropsied after 7 months of exposure. 65.5.3.1 Neoplastic Lesions
There was no evidence of exposure-related neoplasia in male or female rats at any dose on the standard regime or in stop-
65.6 Genotoxicity of PCP and TCHQ
exposure female rats. By the end of the two-year study, stopexposure male rats had an increased incidence of malignant mesothelioma relative to the controls (18% vs. 2%). The incidence of nasal squamous cell carcinoma was likewise elevated relative to the controls in stop-exposure male rats (10% vs. 2%), but this difference was not statistically significant (p > 0.05). However, the incidence of nasal cell carcinoma in this group was significantly elevated relative to that in NTP historical controls (0.4% ± 1.0%).
65.5.3.2 Non-neoplastic Lesions At the 7 -month interim evaluation of the stop-exposure groups, an increased incidence of centrilobular hepatocyte hypertrophy was observed in both males and females, while hepatocyte cytoplasmic vacuolization was elevated in males only. At the 2-year sacrifice there was no hepatocyte hypertrophy and no elevation of hepatocyte cytoplasmic vacuolization, nor was there any exposure-related increase in non-neoplastic lesions at any site. The NTP concluded that there was no evidence of carcinogenic activity in male or female F3441N rats fed diets containing 99% pure PCP at 200, 400, or 600 ppm for 2 years, some evidence of carcinogenicity in male F3441N rats fed a diet containing 99% pure PCP at 1000 ppm for 1 year followed by a control diet for 1 year, and no evidence of carcinogenic activity in female F3441N rats fed a diet containing 99% pure PCP at 1000 ppm for 1 year followed by a control diet for 1 year (NTP, 1999).
65.6 GENOTOXICITY OF PCP AND TCHQ PCP is almost a textbook example of a nonmutagenic, nonDNA-reactive compound. Because PCP has been tested numerous times for genotoxicity, it is expected that false positives will have occurred at a rate consistent with the overall falsepositive rate of the tests applied. All positive genotoxicity tests performed in the presence of microsomal extracts are likely to reflect the genotoxicity of TCHQ or other DNA-reactive metabolites rather than the parent compound. Van Ommen et al. (1986) demonstrated that the rate of PCP binding to calf thymus DNA in the presence of rat liver microsomes is correlated with the formation of TCHQ and the 1,2-isomer of TCHQ. Keeping in mind that yeast are rich in cytochrome P450, it is reasonable to wonder whether observations that PCP is mutagenic or recombinogenic in these lower eukaryotes might be explained by conversion to TCHQ or other metabolites. It would clarify matters if metabolite concentrations were routinely measured whenever genotoxicity testing of PCP is performed.
65.6.1 SHORT-TERM TESTS OF PCP 65.6.1.1 Isolated DNA PCP at concentrations up to 100 mM (27 mg/ml) did not bind covalently to calf-thymus DNA and produced no single-strand DNA breaks in bacteriophage DNA (Witte et aI., 1985).
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65.6.1.2 Bacteria A comprehensive review ofthe literature by Seiler (1991) and a more concise review by an anonymous author (NTP, 1989) both indicated that PCP was almost uniformly negative for induction of gene mutation in bacteria. The one exception was a study by Nishimura et al. (1982), performed in the presence of activated rat liver microsomes (S9). More recently, Gopalaswamy and Nair (1992) reported that PCP was mutagenic in Salmonella typhimurium in the presence of activated S9. Note, however, that Haworth et al. (1983) found that technical-grade PCP was negative for mutagenicity in four S. typhimurium strains in both the presence and absence of activated S9.
65.6.1.3 Lower Eukaryotes PCP generally has been found to be positive for genotoxicity in yeast and mold, although most studies were inadequate or inadequately reported (reviewed by Seiler, 1991). Fahrig et al. (1978) found evidence for mutation and intragenic (but not intergenic) recombination in yeast incubated with PCP at 400 ).Lg/ml. A separate group of investigators reported both positive (Simmon et al., 1979) and negative (Simmon and Kauhanen, 1978) results for mitotic recombination in yeast.
65.6.1.4 Mammalian Cells In Vitro Ehrlich (1990) found no evidence of DNA single-strand breaks or alkali-labile sites in Chinese hamster ovary (CHO) cells cultured in the presence of PCP at concentrations up to 10 ).Lg/ml. Wang and Lin (1995) found no evidence of increased DNA breakage in mouse embryonic fibroblasts cultured with PCP at concentrations up to 240 ).LM (64 ).Lg/ml). Jansson and Jansson (1986) found no induction of point mutations in Chinese hamster lung fibroblasts (V79 cells) with PCP concentrations in the range 6-50 ).Lg/ml. Galloway et al. (1987) looked for cytogenetic changes in CHO cells treated with the same technicalgrade PCP formulation used in the NTP mouse chronic carcinogenicity study (NTP, 1989). In the absence of S9 the assay was clearly negative for chromosomal aberrations (CAbs); in the presence of S9 the results were positive in one trial and equivocal in another. In the presence of S9 the sister chromatid exchange (SCE) response was clearly negative. The SCE response was said to be "weakly positive" in the absence of S9, but this conclusion does not appear to be supported by the data. In the absence of exogenous metabolic activation, Ishidate (1988) found no induction of CAbs in cultured Chinese hamster lung fibroblasts at PCP concentrations up to 60 ).Lg/ml; when the toxic threshold was raised by changing to an intermittent treatment regimen, CAbs were induced at a PCP concentration of 300 ).Lg/ml. Ziemsen et al. (1987) found no elevation of SCEs or CAbs in lymphocytes from healthy donors following incubation with technical-grade NaPCP at concentrations up to 90 ).Lg/ml. It should be noted that the in vitro SCE assay appears to be of little value to cancer risk assessment owing to its low specificity (Ashby, 1993; Tennant et aI., 1987).
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65.6.1.5 Mice Exposed In Vivo A mammalian spot test (designed to reveal mutation, deletion, and recombination events) of purified PCP in hybrid mice yielded equivocal results. Single transplacental injection of 152, 86, or 40 females, respectively, with maternal PCP doses of 0, 50, or 100 mg/kg produced color spots said to be of "definite" genetic relevance in 1/967 (0.10%),2/316 (0.63%), and 2/157 (1.3%) offspring (Fahrig et aI., 1978). The response was dose-related, but the small numbers easily could have arisen by chance. A negative interpretation is supported by the finding, in the same test system, that treatment with purified PCP at 50 mg/kg had no effect on the incidence of genetically relevant color spots produced by ethylnitrosourea (Fahrig and Steinkamp-Zucht, 1996). Male mice injected i.p. with technical- or reagent-grade PCP at doses up to 50 mg/kgday for 5 days did not exhibit dose-related increases in the frequency of abnormal sperm morphology 35 days after the initial injection (Osterloh et aI., 1983). 65.6.2 OCCUPATIONAL EXPOSURE TO PCP An investigation of CAbs and SCEs was performed on the peripheral blood lymphocytes of 22 male workers engaged in the production of PCP or NaPCP (Bauchinger et aI., 1982). At the time of the measurements the mean length of employment was 11.4 years. All 22 of the PCPlNaPCP-exposed men were smokers while the sex- and age-matched control group included 9 smokers and 13 nonsmokers. The percentage of cells containing CAbs was significantly elevated in the exposed workers vs. the controls (p < 0.004), primarily due to elevation of dicentrics (p = 0.020) and acentrics (p = 0.037). Comparison of the exposed workers to the smoking controls also revealed elevation of dicentrics (p = 0.021) and acentrics (p = 0.043). A slight elevation of SCEs in the exposed workers vs. the controls was no longer significant when the comparison was limited to smoking controls. Several factors complicate interpretation of the results of Bauchinger et al. (1982). For one, the investigators did not indicate whether chromosome analyses were conducted in a blind fashion with respect to exposure status. In addition, the degree to which smoking may have confounded the CAb results is not clear because the authors did not report individual CAb data or summary statistics for the 9 controls who smoked. Without such data it is not possible to compare the exposed workers and smoking controls with respect to the percentage of cells containing CAbs or to discern the influence of individual controls on the overall comparison. It is also worrisome that there was no mention of a relationship between exposure level and CAb incidence even though the workers were divided into low- and high-exposure groups. The failure to note the existence of such a relationship can only be interpreted as signifying that no dose dependence was found. Given the apparent absence of a doseresponse and the relatively low CAb frequency in the exposed workers (i.e., within the range considered normal), the overall response must be considered equivocal.
Ziemsen et al. (1987) examined CAb and SCE frequencies in the lymphocytes of 20 workers occupationally exposed to PCp. Mean blood PCP levels in the low- and high-exposure groups were 58 and 330 !1-g/ml, respectively. There was no control group. The average number of CAbs per cell was slightly greater in the high-exposure group (0.040 vs. 0.026), but the difference was not statistically significant. The frequency of SCEs was similar in the two groups. Consideration of the smokers separately yielded negative results as well. 65.6.3 GENOTOXICITY OF TCHQ Unlike PCP, TCHQ binds isolated DNA covalently and produces single-strand breaks (Witte et aI., 1985). TCHQ is clearly mutagenic and damaging to DNA in cultured mammalian cells in the absence of metabolic activation. TCHQ at concentrations as low as 20-25 !1-M (5-6 !1-g/ml) produced single-strand breaks and formed DNA adducts in CHO and V79 cells (Dahlhaus et aI., 1995, 1996; Ehrlich, 1990) and increased the frequency of mutations at a distinct genetic locus in V79 cells (Jansson and Jansson, 1991). TCHQ concentrations as low as 10 !1-M (2.5 !1-g/ml) induced micronuclei in V79 cells (Jansson and Jansson, 1992). One or more reactive oxygen species (including hydrogen peroxide) produced through reaction of an autooxidation product, the semiquinone radical, are implicated in the DNA-damaging actions of TCHQ (Carstens et aI., 1990; Naito et aI., 1994).
65.7 REPRODUCTIVE AND ENDOCRINE TOXICITY Several acute, subchronic, and reproductive studies in rats, mink, and sheep have examined the effects of PCP on reproductive and endocrine function. The results of these studies are summarized in Table 65.8. The most consistent outcome was a decrease in serum thyroxine (T4), a potentially adverse effect. A LOEL between 0.6 and 1 mg/kg-day for the effect of PCP on serum T4 has been observed in rats, mink, and sheep. No NOEL for this endpoint has been identified. The effects ofPCP on serum T4 are similar to those produced by much lower concentrations of TCDD. Sewall et al. (1995) found that serum total thyroxine (TT4) levels were significantly depressed following 30-week administration of TCDD to rats at doses as low as 11 ng/kg-day, while a TCDD dose of 125 ng/kg-day produced a degree of depression comparable to that observed at a PCP dose of 3 mg/kg-day. TCDD at 125 ng/kgday also produced a 2.5-fold increase in serum thyrotropin (TSH). By contrast, PCP-induced depression of T4 has been found to be associated with decreased serum TSH and thyroid mass. The latter relationship between T4 and TSH is found in a limited set of physiological conditions, including caloric restriction (Scanlon and Toft, 1996). The disparity in the direction of the TSH change produced by TCDD and PCP suggests that these substances act via dissimilar mechanisms to depress T4 levels.
65.7 Reproductive and Endocrine Toxicity
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Table 65.8 Summary of the Reproductive and Endocrine Toxicity of PCpa PCPlNaPCP Route
grade
Doses
van Raaij et aI., 1991b, 1993 (acute exposure study in males)
i.p.
AG-PCP
0,0.6, 1.1, 1.8, 28 mglkg
van Raaij et aI., 1994 (acute exposure study in males)
i.p.
AG-PCP
0,7,14, 28 mg/kg
Jekat et aI., 1994 (28-day exposure study in females)
gav.
Study
Effects at the LOEL
LOEL
NOEL
+serum TT4 and FT4
1.8 mg/kg
1.1 mg/kg
+uptake of T 4 into cerebrospinal
7 mg/kg
n.d.
TG-NaPCP: serum TT4. TT3. TT4:TT3, and TSH AG-PCP: serum TT4 and TSH
TG-NaPCP. AG-PCP: 3 mg/kg-day
TG-NaPCP, AG-PCP: n.d.
+maternal and neonatal body weight; +pup survival
30 mg/kg-day
3 mg/kg-day
+whelping rate; t severity of
I mg/kg-day
n.d.
+serum T4 in F2 and F3 males and F3 females; +thyroid mass
1 mg/kg-day
n.d.
+serum T4; t serum insulin;
0.6 mg/kg-day
n.d.
1 mg/kg-day
n.d.
1 mg/kg-day
n.d.
Thyroid-hormone related effects in rats
fluid and brain
TG-NaPCP; AG-PCP
vehicle, 3 mg/kg-day TG-PCP,3 and 30 mg/kg-day AG-PCP
diet
TG-PCP
0,3, 30 mg/kg-day
Beard et aI., 1997 (one-generation study; maternal exposure only)
diet
AG-PCP
0, I mg/kg-day (target)
Beard and Rawlings, 1998 (three-generation study: in F 1, maternal exposure only)
diet
+
+
Reproductive effects in rats Schwetz et aI., 1978 (one-generation study) Reproductive effects in mink
AG-PCP
1 mg/kg-day (target)
maternal uterine cysts
in F3 females; t severity of testicular hyperplasia in F3 males; t incidence of mild, multifocal, cystic hyperplasia of the prostate in F2 and F3
Endocrine-related effects in ewes Rawlings et aI., 1998 (43-day study, twice weekly dosing)
i.g.
AG-PCP
vehicle, 2 mg/kg (time-averaged: 0.6 mglkg-day)
t
severity of oviductal intraepithelial cysts
Reproductive effects in sheep Beard et aI., 1999a (one-generation study: maternal exposure only)
diet
AG-PCP
0, 1 mglkg-day (target)
+maternal serum TT4; t maternal thyroid follicle size and lymphocytic infiltration into the uterine endometrium; weaning weight of female lambs
+
Beard et aI., 1999b (two-generation study: in Fj, maternal exposure only; in F2, exposure of rams only)
diet
AG-PCP
0, 1 mg/kg-day (target)
+
In F2: serum TT4; t scrotal size; t severity of semiferous tubule atrophy; sperm density
+
aAbbreviations: AG, analytical or purified grade; FT3, free triiodothyronine; FT4, free thyroxine; gav., gavage; i.g., intragastric; i.p., intraperitoneal; n.d., none determined; TG, technical grade; TSH, thyrotropin; TT4, total thyroxine.
The effect of PCP on T 4 levels suggests a role for this
et aI., 1986). In ex vivo experiments, PCP, but not TCHQ, was
molecule, the principal metabolite of hexachlorobenzene
a competitive inhibitor of T4 binding to rat sera, indicating that
(HCB), in the altered thyroid function accompanying HCB-
PCP depression of T4 levels does not involve metabolism to
induced porphyria (Kleiman de Pisarev et aI., 1990; Rozman
TCHQ (van Raaij et aI., 1991a),
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CHAPTER 65 Pentachlorophenol
65.7.1 THYROID HORMONE EFFECTS IN RATS Two studies in male, 12-week-old Wistar rats examined the effects of a single i.p. dose of purified PCP on serum levels of thyroid hormones. Maximum depression of TT4 was observed at 4-6 hr postexposure with recovery within 48 hr. Serum TT4 was depressed at a dose of 1.8 mg/kg (p < 0.001) but not at 0.6 mg/kg (van Raaij et aI., 1991b) or 1.1 mg/kg (van Raaij et aI., 1993). Depression of serum free thyroxine (FT4) demonstrated a similar dose-response (van Raaij et aI., 1993). The results of these studies indicate an acute, i.p. NOEL in rats of 1.1 mg/kg for decreased serum TT4 and FT4. The same group of investigators found that single i.p. doses of purified PCP at 7, 14, or 28 mg/kg decreased the uptake of T4 into the cerebrospinal fluid of male Wistar rats by 34, 60, or 75%, respectively, while uptake of T 4 into total brain tissue and specific brain structures was also decreased (van Raaij et aI., 1994). Jekat et al. (1994) studied the effects of subchronic exposure to PCP on serum TSH, TT4, and total triiodothyronine (TT3) in female Wistar rats. Eight animals per group were treated twice daily by gavage for 28 days with purified PCP at 3 or 30 mg/kg-day or with technical-grade NaPCP (TG-NaPCP) at 3 mg PCP/kg-day. The control group received the gavage vehicle. Mean serum TT4 was approximately 50% of the control value in both the TG-NaPCP and low-dose PCP groups and approximately 30% of the control value in the high-dose PCP group (p < 0.0025). Lesser reductions were seen for serum TT3 with TG-NaPCP (p < 0.05) and with the high dose of purified PCP (p < 0.01). The serum TT4: TT3 ratio was substantially reduced in all treatment groups (p < 0.0025). Serum TSH was approximately 70% of the control level in all treatment groups (p < 0.05). 65.7.2 REPRODUCTIVE EFFECTS IN RATS In a single-generation, dietary exposure study of reproductive toxicity, technical-grade PCP at 0, 3, or 30 mg/kg-day was fed to groups of 10 male and 20 female SD rats for 62 days before mating, throughout gestation, and for 21 days postpartum (Schwetz et aI., 1978). The maternal NOEL was 3 mg/kg-day based on reduced body weight at 30 mg/kg-day. The NOEL for the offspring was 3 mg/kg-day based on reproductive effects at 30 mg/kg-day (reduced pup survival across litters and reduced neonatal body weight). The study design was limited: only one generation, only two treatment groups, and too few males under test. The reproductive effects observed were likely to have been secondary to maternal toxicity. 65.7.3 REPRODUCTIVE AND ENDOCRINE EFFECTS IN MINK The effects of analytical-grade PCP on fertility, hormone levels, and histopathology were examined in a single-generation study
in mink (Beard et aI., 1997). PCP at a target dose of 1 mg/kgday was fed to 10 female mink; a control group of 10 females received no treatment. PCP was "sprayed evenly onto a weekly supply of the pelleted ration" which was mixed with other dietary components before feeding. The investigators apparently did not verify the PCP content of the feed or the uniformity of distribution. The PCP-containing diet was fed to adult females for 3 weeks before mating, throughout gestation, and until 8 weeks postpartum, when weaning occurred and the maternal mink were sacrificed. Where treatment-related effects on fertility were noted the absolute differences were small; the low number of animals tested rendered the results ambiguous. The fraction of mated females that subsequently accepted the second mating was significantly decreased in the PCP group (5/9 vs. 7/8, p < 0.01). The toxicological relevance of such a measure is obscure. A better measure of fertility, the fraction of mated females that subsequently whelped (whelping rate), was also significantly reduced in PCP-treated animals (5/9 vs. 7/8, p < 0.01). Nonsignificant reductions were observed in the proportion of mated females with implantation sites visible at necropsy (7/9 vs. 8/8), the mean number of implantation sites per animal (4.6 vs. 6.8), and the mean litter size of mink that whelped (3.4 vs. 4.5). PCP exposure had no effect on maternal serum levels of the measured hormones (progesterone, estradiol, cortisol, and T4). The maternal tissues subject to histopathologic examination were not specified. The only treatment-related histologic change reported was an increased severity score for uterine cysts (p < 0.05). The same laboratory conducted a three-generation reproductive toxicity study of analytical-grade PCP in mink, continuing to apply their method of spraying PCP onto a weekly supply of pelleted ration to deliver a target dose of 1 mg/kg-day (Beard and Rawlings, 1998). The F2 mink in this study were the offspring from the one-generation study discussed above. Thus, the F2 mink (8 males, 8 females) were born to mothers exposed from 3 weeks prior to mating, while the F3 mink (8 males, 10 females) were born to mothers exposed from the point of conception. Implantation sites were not counted, but otherwise the tests of fertility was the same as those applied in the one-generation study. No evidence of treatment-related effects on fertility was observed. Measurement of serum progesterone, estradiol, cortisol, T 4, and testosterone at sacrifice revealed depression of T4 in all treatment groups, achieving statistical significance in F2 males, F3 males, and F3 females (p < 0.05). The animals were necropsied and all major organs subjected to histopathologic examination. Mass or size changes in the thyroid gland, adrenal glands, and testis were reported; only the effect on the thyroid appears to have been of potential toxicological relevance. Thyroid mass was lower in all PCP treatment groups, achieving statistical significance only in F3 females (p < 0.05). Adrenal mass was higher in F2 females, but this difference probably reflected unusually low values in the control group. Mean adrenal mass values in F2 and F3 control females were 0.105 and 0.150 mg/kg, while mean values in F2
65.7 Reproductive and Endocrine Toxicity and F3 treated females were 0.149 and 0.146 mg/kg. During development, testis length in the F2 was smaller (p < 0.05), but body-weight adjusted testis length was not. Testis mass and length at autopsy were unaltered by treatment in either the F2 or F3. Testicular hyperplasia was more severe in PCP-treated than untreated F2 males (p < 0.05). Fifty percent of PCPtreated males had mild, multifocal, cystic hyperplasia of the prostate; the control incidence was not given. Treated males displaying multifocal cystic hyperplasia had twofold higher mean serum testosterone levels than treated males without the lesion. In treated F2 females and F3 males, the severity of adrenal cortex vacuolization was less severe than in their respective control groups. 65.7.4 REPRODUCTIVE, ENDOCRINE, AND HISTOLOGIC EFFECTS IN SHEEP
Rawlings et al. (1998) conducted a study of the effects of analytical-grade PCP on hormone levels and histopathology in adult ewes. Six ewes were given PCP intragastrically (i.g.) in gelatin capsules twice weekly for 43 days at a dose of 2 mg/kg. A designated control group of six age- and weight-matched ewes received empty capsules. Two other control groups for experiments not entailing PCP were started 22 and 11 days earlier. Ewes were estrus-synchronized during treatment. On treatment day 36, ewes were bled at regular intervals to detect any changes in the pulsatile patterns of luteinizing hormone (LH) and follicle-stimulating hormone (FSH) secretion. In addition, serum levels of progesterone, estradiol, cortisol, T4 (presumably TT4), insulin, LH, and FSH were measured in blood pooled over each hour of collection. At the end of the treatment period, animals were necropsied and histopathologic examination performed on all major tissues. The only treatment-related effects on hormones were significant T4 depression (p < 0.05) and insulin elevation (p < 0.05) on the day of the intensive bleed. The only treatment-related histopathologic change noted was increased severity of oviductal intraepithelial cysts; data supporting this observation were not provided. It was generally not clear whether a given analysis reflected comparison to the designated or combined controls. The same laboratory conducted a single-generation study of the effects of PCP on reproductive and general endocrine function in ewes (Beard et al., 1999a). Two groups of 13 ewes were given a control diet or feed treated with PCP to deliver a target dose of 1 mg/kg-day. Treated feed, prepared weekly, consisted of alfalfa pellets sprayed with PCP. The PCPcontaining diet was fed to adult females for 5 weeks before mating, throughout gestation, and until 2 weeks after weaning, when the ewes were sacrificed. The investigators argued that it was not necessary to adjust the dose based on feed intake because each ewe consumed all alfalfa pellets offered. Ewes were estrus-synchronized prior to mating. Blood samples were taken frequently from the ewes during pregnancy and lactation. On postweaning day 5, six ewes per group were bled at regular intervals during the day and night for determination of pulsatile
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LH and FSH secretion. Two days later the same ewes were bled at regular intervals for 1 hr before and 5 hr after injection of a mixture of gonadotropin-releasing hormone (GnRH), bovine TSH, and adrenocorticotropin (ACTH). The same serum hormones measured by Rawlings et al. (1998) were obtained in hourly pooled blood drawn during the periods of intensive bleeding. Fertility parameters (size and number of ovarian follicles and corpora lutea) and fetal growth rate (head diameter) were obtained in vivo using ultrasonography. At the end of the treatment period, ewes were necropsied and histopathologic examination performed on all major tissues. The weaning weight of female lambs was depressed compared to the female controls, consistent with there being slightly more lambs per ewe (fewer singletons) in the PCP groups. Cortisol levels in both control and PCP-treated ewes were highly variable. PCP-treated ewes exhibited significant depression of T4 throughout pregnancy and lactation and on postweaning day 5 (p < 0.05). PCP had no effect on the response to stimulation with the trophic hormones. Histopathologic examination revealed increased thyroid follicle size (p < 0.01) and increased lymphocytic infiltration into the uterine endometrium (p < 0.05). The same laboratory also conducted a study of reproductive and endocrine function in rams exposed to PCP, using the same method of treating the feed to deliver a target PCP dose of 1 mg/kg-day (Beard et aI., 1999b). Beginning 5 weeks prior to conception, maternal ewes were fed either a control diet or a PCP-containing diet. Ewes received the treated diet through weaning; the 5 male offspring in the PCP group were maintained on the same treated diet until sacrifice at age 28 weeks. Tests of testicular growth (using ultrasound), sperm quality, and sexual behavior were performed at appropriate time points during development. When the rams were 27 weeks of age, levels of the same serum hormones as those measured by Rawlings et al. (1998) plus testosterone were evaluated before and after stimulation with a mixture of GnRH, TSH, and ACTH. Hormones were also monitored in blood samples taken every 2 weeks beginning at age 6 weeks. At sacrifice, animals were necropsied and the pituitary, adrenal, thyroid, pancreas, and reproductive tissues were subjected to histopathologic examination. Rams in the PCP group had increased testicular development; scrotal circumference was approximately 10 to 15% greater than that of the controls throughout the measurement period (p < 0.05). The only effect of PCP on hormones was decreased T4 in rams prior to 18 weeks of age. From age 18 to 26 weeks, T4 levels in the treatment group were indistinguishable from control values. At age 28 weeks a significant decrease was again observed. At necropsy, the only possibly treatmentrelated effect noted was a nonsignificant increase in mean testes weight in the PCP group (586 g) compared to the controls (496 g), a finding in support of the larger scrotal circumference observed in the PCP-treated animals during life. Histologic examination revealed that PCP-treated rams had higher severity scores for seminiferous tubule atrophy (p < 0.05) and reduced sperm density in the body of the epididymides (p < 0.05). Be-
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cause sperm quality in the living rams was unaffected by PCP treatment, the reported histopathologic findings must be considered tentative in the absence of independent replication.
65.8 DEVELOPMENTAL TOXICITY Four teratology studies of oral PCP were identified, three in rats and one in rabbits. The rabbit study and the most recent rat study were performed by the same laboratory. These two studies tested recent (c. 1992) lots of a commercially available, technical-grade PCP. Results of the teratology studies in rats and rabbits are summarized in Table 65.9.
65.8.1 RAT Hoberman (1994a) performed a teratology study of technicalgrade PCP (89% pure) in a strain of CD rats under contract to the Pentachlorophenol Task Force, an industry group. Twentyfive female rats per group were mated and then administered PCP by gavage at dose levels of 0, 10, 30, or 80 mg/kg-day on gestation days 6-15. The litter and/or fetal incidences of various malformations or variations were increased (significantly or otherwise) at 80 mg/kg-day. The incidence of "slight to moderate" dilatation of the kidney pelvis was significantly elevated at the high dose (p :::: 0.01), occurring in 4/22 litters (4 fetuses) at this dose, 1123 litters (1 fetus) at 30 mg/kg-day, and in none of the low-dose or
Table 65.9 Summary of the Developmental Toxicity of PCP in Rats and Rabbits
PCP
Species/strain:
LOEL
study
Route
grade
Effects at LOEL
CD rat: Hoberman (l994a)
gavage
technical
Maternal:
NOEL (mg/kg-day)
80
30
80
30
-l- weight gain. Developmental:
t gross malformations (external, soft tissue);
t skeletal anomalies (delayed or residual ossification at multiple sites, extra vertebra and rib pair) Sprague-Dawley rat:
gavage
Schwetz et al. (1974)
technical and
Maternal:
technical: 30
technical: 15
purified
t weight gain
purified: 34.7
purified: 15
Developmental:
technical: 15
technical: 5.8
t percentage males;
purified: 5.0
purified: n.d.
43
I3
I3
4
15
7.5
None
30 (highest
-l- fetal body weight;
t subcutaneous edema; t skeletal anomalies (lumbar spurs, delayed skull ossification) Sprague-Dawley rat:
diet
purified
Maternal:
t
Welsh et al. (1987)
weight gain;
t "ringed eye"; t
vaginal hemorrhaging
Developmental:
t fetal body weight;
t skeletal anomalies (misshapen centra) New Zealand rabbit: Hoberman (l994b)
gavage
technical
Maternal:
-l- weight gain Developmental: No effects
dose tested)
65.8 Developmental Toxicity
control litters. The only other gross external or tissue lesion in the 30 mg/kg-day group was a ventricular septal defect. Minor ossification changes observed at 30 mg/kg-day are judged to be of no developmental significance. Based on malformations and other effects observed at the high dose, the developmental NOEL for this study was 30 mg/kg-day. Aside from the difference in rat strains, there is no obvious explanation for why the NOEL observed in this study was so much higher than the NOELs found in the two other rat studies discussed below. Schwetz et al. (1974) evaluated the teratogenic potential of sample lots of commercial (TG) and purified PCP in SD rats. The test substances were given by gavage on gestation days 615 at dose levels of 5.8, 15,34.7, or 50 mg/kg-day for TG-PCP and 5.0, 15, 30, or 50 mg/kg-day for purified PCP, with one control group for both PCP formulations. The TG-PCP doses of 5.8 and 34.7 mg/kg-day were chosen to correspond to PCP doses of 5.0 and 30 mg/kg-day, respectively. Both the TG-PCP and purified PCP contained nondetectable levels of TCDD, but the detection limit at that time was quite high (50 ppb). The TG-PCP contained levels of the nonphenolic contaminants HxCDD, HxCDF, HpCDF, OCDF, HpCDD, and OCDD at 4, 30, 80, and 80, 125, and 2500 ppm, respectively. Between days 6 and 21, average maternal weight gains in the 30 mg/kg-day purified PCP and 34.7 mg/kg-day TG-PCP groups were, respectively, 26% and 74% of the control value. The maternal NOEL was 15 mg/kg-day for both PCP grades. The percentages of resorbed fetuses per animal and per litter were significantly increased in groups receiving TG-PCP at doses ~ 15 mg/kg-day or purified PCP at doses ~30 mg/kg-day. Even though the LOEL for fetal re sorption was lower for animals given TG-PCP, at the two highest dose levels (30/34.7 and 50 mg/kg-day) the percentage of resorbed fetuses was greater for purified PCP (98 and 100%) than for TG-PCP (27 and 58%). There were no data on fetal sex, weight, and length for the 50 mg/kg-day purified PCP group because all fetuses were resorbed. The sex ratio of surviving pups was markedly altered at the higher doses, with male to female ratios of 3.8 to 1 in the 50 mg/kg-day TG-PCP group and 4.9 to 1 in the 30 mg/kgday purified PCP group. Fetal body weights were significantly decreased at the 30 mg/kg-day dose of purified PCP and at TG-PCP doses ~34.7 mg/kg-day. Crown-rump length was significantly decreased in the 30 mg/kg-day purified PCP group. At 5.0 mg/kg-day purified PCP, the percentage of litters with delayed skull ossification was significantly elevated (9/15 vs. 6/31 in controls). Based on the dose-related effects on fetal development observed at all doses, this study produced no NOEL for the developmental toxicity of purified PCP. In the 15 mg/kgday TG-PCP group, the percentage of litters with subcutaneous edema or lumbar spurs was significantly elevated. In the absence of other signs of teratogenicity at the lowest dose, the developmental NOEL for the technical-grade PCP tested was 5.8 mg/kg-day based on skeletal and soft tissue anomalies detected at higher dose levels. The results of Schwetz et al. (1974) seem to suggest that impurities present in technical-grade PCP are responsible for diminishing some of the maternal and developmental toxicity of
1497
PCP by unknown mechanisms. The differential toxicity of the TG-PCP and purified PCP formulations cannot be explained by small differences in their relative PCP content; for example, the incidence of delayed ossification for the 15 mg/kg-day purified PCP group was approximately twice that of the TG-PCP group receiving nominally double the dose (30 mg/kg-day). Whatever the explanation for the differential toxicity of the two formulations, this study showed that the developmental toxicity of PCP does not result from impurities present in the TG product. A study performed by the U.S. Food and Drug Administration examined the teratogenic potential of purified PCP in SD rats (Welsh et aI., 1987). Weanling animals of both sexes received PCP at 0, 60, 200, or 600 ppm diet (corresponding to average daily intakes of 0, 4.0, 13, or 43 mg/kg-day for females). After 181 days on this dietary regimen, males and females were permitted to mate with animals from the same dose group. For the teratology phase, females in each group were continued on the same daily dose until Caesarian section on gestation day 20. The test compound was a highly purified product recrystallized from >99% pure PCP. The only pCDD detected was 1.3 ppb OCDD; no pCDF impurities were found. Signs of maternal toxicity observed at the highest dose were reduced body weight gain, "ringed eye" (50% incidence), and vaginal hemorrhaging (25% incidence). The maternal NOEL was 13 mg/kg-day. At a maternal dose of 43 mg/kg-day, essentially complete resorption of fetuses (99.5%) was observed, indicating a profound effect on reproductive competence. In the 13 mg/kg-day group, the percentage of dams with two or more resorbed fetuses was twice as great as in the control group. At this dose, the major fetal effects were a highly significant (p < 0.01) increase in misshapen centra by litter and by fetus and an approximately 10% weight reduction in fetuses of both sexes. Fetal body weight was depressed to a lesser extent (3-4%) in male and female pups in the 4.0 mg/kg-day group, approaching statistical significance for males (p = 0.06). There were no fetal abnormalities at the low dose. The developmental NOEL for highly purified PCP was 4.0 mg/kg-day based on fetal weight reduction and skeletal abnormalities at 13 and 43 mg/kg-day. Unlike the study performed by Schwetz et al. (1974) in the same rat strain, Welsh et al. (1987) found no evidence for delayed skull ossification. Potential explanations for this discrepancy may lie in differences in exposure route and onset (dietary exposure from weaning vs. gavage exposure on gestation days 6-15). It is conceivable that adaptive tolerance played a role in limiting the toxic effects of PCP in the FDA study. 65.8.2 RABBIT
Hoberman (1994b) performed a developmental toxicity study of technical-grade PCP (88-89% pure) in New Zealand white rabbits under contract to the Pentachlorophenol Task Force. Twenty does per group were artificially inseminated, presumed pregnant, and then administered PCP by gavage at dose levels of 0, 7.5, 15, or 30 mg/kg-day on gestation days 6-18. At Caesarian section on gestation day 29, the number of litters was 17
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Pentachlorophenol
in the control group and 17, l3, and 18 in the low-, middle-, and high-dose PCP groups, respectively. Mean maternal body weight gain in the 15 mg/kg-day group was marginally (:'52.5%) but significantly depressed during the dosing period but recovered by gestation day 19. Mean maternal body weight gain in the 30 mg/kg-day group was likewise marginally (:'53.5%) but significantly depressed during the dosing period and remained depressed (:'53.8%) throughout the entire 29 days of gestation. The maternal NOEL was 7.5 mg/kg-day based on transient decreased weight gain in animals receiving 15 mg/kg-day and lasting decreased weight gain in animals receiving 30 mg/kg-day. No deaths, abortions, or preliminary deliveries occurred. No treatment-related developmental toxicity was observed at any dose. The developmental NOEL was 30 mg/kg-day based on the observation of no developmental effects at the highest dose tested.
65.9IMMUNOTOXICITY In several studies, TG-PCP was found to suppress humoral immunity in mice treated subchronically at oral doses as low as 9 or 10 mg/kg-day; no NOEL was identified. There is suggestive evidence for a developmental immunotoxic effect of PCP; partially purified PCP suppressed delayed-type hypersensitivity and humoral immunity in rats exposed prenatally and then subchronically to oral doses as low as an estimated 0.74 mg/kgday, the lowest dose tested. However, confidence in that finding is weak because of the absence of dose dependence. The only clearly dose-related immunotoxic response to AGPCP was the unexpected occurrence of splenic tumors in mice surviving primary challenge with tumor virus and secondary challenge with virus-transformed sarcoma cells; the LOEL was 9 mg/kg-day (the lowest dose tested). This finding suggests that PCP possesses a partial complement of the immunosuppressive activity associated with the technical-grade formulation. Apparently there has not yet been any attempt to replicate the result. Nevertheless, the observation of increased susceptibility to invasive tumorigenesis must be considered as evidence of a potentially important immunotoxic effect of PCP. Studies are presented chronologically within Sections 65.9.165.9.3. 65.9.1 RAT 65.9.1.1 lWo-Generation Dietary Exposure Female SD rats were fed partially purified (97%) PCP at 0, 5, 50, or 500 ppm diet from weaning, throughout mating, gestation, and lactation (Exon and Koller, 1983). Progeny were weaned at 3 weeks and maintained on the maternal exposure regime until commencement of immunotoxicity testing at 13 weeks. Based on recommended values for body weight and food consumption for female and weanling SD rats in subchronic studies (U.S. EPA, 1988), the average PCP dose
levels can be estimated as 0.49, 4.9, and 49 mg/kg-day for the dams and 0.74, 7.4, and 74 mg/kg-day for the weanlings. Delayed-type hypersensitivity (DTH), humoral immunity, and macrophage activation were assayed following in vivo antigen challenge. The investigators found evidence for suppression of DTH and humoral immunity in PCP-treated progeny, but dose dependence was lacking. The DTH response was significantly depressed in all PCP treatment groups (p :'5 0.03); mean values were 82%, 72%, and 74% of controls in the 5, 50, and 500 ppm groups. Humoral immunity (serum antibody response to bovine serum albumin) was likewise reported to be significantly depressed in all treatment groups (p :'5 0.0001); mean values were 44%, 58%, and 45% of controls in the 5, 50, and 500 ppm groups. Without further information it is not possible to distinguish between a saturated response (spanning two orders of magnitude in dose) and a spurious result. Clear evidence of dose-related immunoactivation (increased phagocytosis by macrophages) was observed, with statistical significance achieved at the two highest doses (p :'5 0.01). 65.9.1.2 lWenty-Eight-Day Dietary Exposure Blakley et al. (1998) administered 99% pure PCP via oral gavage to 10 male F344 rats twice weekly for 28 days at a dose of 2.0 mg/kg (time-averaged dose, 0.57 mg/kg-day). A control group of 10 rats received the olive oil vehicle. At the end of the exposure period animals were sacrificed and organ weights recorded. Immune parameters assayed were IgM antibody plaque formation (following immunization with sheep red blood cells) and lymphocyte blastogenesis in spleen cell suspensions, phagocytic function of peritoneal macrophages, and lymphocyte surface marker expression in whole blood. Mean body-weight gain, kidneylbody-weight ratio, and liverlbody-weight ratio were moderately increased in PCPtreated animals. PCP inhibited background lymphocyte blastogenesis but not mitogen-stimulated blastogenesis, an effect that cannot by itself be considered a sign of reduced humoral immunity. PCP exposure did not affect the number of plaque-forming colonies (PFCs) per spleen, despite the fact that the number of PFCs per unit viable cells was reduced to 61 % of the control value (p = 0.006). These results are difficult to interpret. As discussed by the investigators, increased spleen cellularity (in compensation for diminished lymphocyte function) is an unlikely explanation because the spleenlbody-weightratio was not at all increased by PCP treatment. Based on the absence of effects in the other tests applied, PCP treatment did not appear to affect humoral immunity overall. 65.9.2 MOUSE 65.9.2.1 Subchronic Dietary Exposure Kerkvliet et al. (1982a) fed male B6 mice a diet containing 50 or 500 ppm AG-PCP (>99% pure), 50 or 500 ppm TG-PCP (86% pure), or a control diet for 10-12 weeks prior to immunotoxicity testing. Measured concentrations of both PCP grades
65.9 Immunotoxicity
were within 15% of target levels. B6 mice (along with C3 mice) are progenitors of the B6C3Fl mice routinely used by the NTP in toxicity and carcinogenicity studies. Recommended body weight and food consumption values for B6C3Fl mice in subchronic studies are available (U.S. EPA, 1988); based on those recommended values, PCP dose levels in the study by Kerkvliet et al. (l982a) can be estimated as 9.0 and 90 mg/kg-day for AG-PCP and 7.7 and 77 mg/kg-day for TG-PCP. Immunocompetence was measured with tests of host susceptibility to virus infection and tumor growth. Components of the immune response presumed to underlie an organism's net infectibility and tumor susceptibility (i.e., macrophage phagocytic activity and T-cell cytotoxicity) were tested separately following in vivo antigen challenge. Mice given TG-PCP exhibited a dose-related increase in tumor incidence following challenge with syngeneic sarcoma cells, with statistically significant elevation at the high dose. In mice given a primary challenge with Moloney sarcoma virus (MSV), TG-PCP at 500 ppm diet produced progressive (i.e., fatal) primary tumors in 6/11 mice compared to 0/11 in the 50 ppm group and 0/16 in the controls. Fifteen weeks later, survivors were given a secondary challenge with MSVtransformed sarcoma cells (MSB). The incidence of progressive secondary tumors in PCP-treated mice was elevated, but the elevation was not significant and not dose-dependent. Necropsy of animals resistant to both the MSV and MSB challenges revealed splenic tumor formation in 0/13, 3/6, and 0/3 of the mice given 0, 50, and 500 ppm TG-PCP, respectively. An in vitro test of T-cell activity following in vivo sensitization revealed a dose-related decrease in cytotoxicity that could be explained by a diminished number of lysing units per spleen. In contrast to its immunosuppressive actions, TG-PCP improved resistance to a lethal virus, but not significantly. Resistance to EMCV-induced mortality is considered to be macrophage-dependent; a direct test of macrophage activation in vitro (following in vivo sensitization) revealed a dose-related enhancement of phagocytosis. Contrary to the results with TG-PCP, animals exposed to AG-PCP showed no alteration in host susceptibility to syngeneic tumor cells, primary challenge with MSV, or secondary challenge with MSB. T-cell activation and macrophage phagocytic activity were similarly unaffected by AG-PCP. However, necropsy of animals resistant to both primary MSV and secondary MSB challenges revealed splenic tumor formation in 0/13, 2/9, and 4/9 of mice given 0, 50, and 500 ppm AG-PCP, respectively. These findings suggest that the PCP molecule, apart from contaminants present in technical-grade formulations, possesses immunosuppressive activity. Kerkv liet et al. (1982b) further assessed the differential immunotoxicity of TG-PCP (86% PCP) and AG-PCP (>99% PCP) in female mice. Groups of 7-10 Swiss-Webster (SW) or B6 mice were exposed to TG-PCP at 0, 50, 100, 250, or 500 ppm diet for 8 weeks prior to primary immunization. Groups of 15-16 SW or B6 mice were exposed to AG-PCP at 0 or 1000 ppm diet. Based on recommended default values for body weight and food consumption in B6C3Fl mice in subchronic studies (V.S. EPA, 1988), the five PCP dose levels can
1499
be calculated as 9, 18,45,90, and 180 mg/kg-day for AG-PCP and 7.7, 15, 39, 77, and 155 mg/kg-day for TG-PCP. Humoral immune responses to in vivo antigen challenge were assessed with assays for hemolytic antibody isotope release, hemolytic plaque formation, and serum hemagglutination. Several measures of humoral immunity were significantly depressed at all TG-PCP levels tested (50-500 ppm), while no depression ofhumoral immunity was observed at the only AG-PCP level tested (1000ppm). 65.9.2.2 Fourteen-Day Gavage Exposure A study in female B6C3Fj mice conducted by Holsapple et al. (1987) also differentiated the immunosuppressive properties of two grades of PCP: partially purified (EC-7) and TG-PCP. Eight mice per dose group were given TG-PCP at 0, 10, 30, or 100 mg/kg-day or EC-7 at 100 mg/kg-day by corn oil gavage for 14 days. The effect of PCP administration on humoral immunity was measured as the spleen-cell antibody response to antigenic activation in vivo and in vitro. The results of in vivo antigen challenge in this study were in agreement with the results of Kerkvliet et al. (1982a); i.e., TG-PCP but not EC-7 was found to suppress the spleen-cell antibody response. However, the in vitro antigen challenge assay revealed no evidence for suppression of the spleen-cell antibody response by TG-PCP. The negative result in the in vitro assay suggests that TG-PCP produces an immunosuppressive action on spleen cells in vivo though a mechanism external to the spleen. By contrast, the in vivo immunosuppressive action of HCDD (a major contaminant of TG-PCP) can be reproduced in the in vitro spleen cell assay (Holsapple et aI., 1984). 65.9.2.3 Acute Gavage Exposure Various contaminant TG-PCP fractions and purified isomers isolated from TG-PCP were examined for their effect on humoral immunity (splenic plasma-cell antibody response to in vivo antibody challenge) in an acute gavage exposure study in B6 mice of both sexes (Kerkvliet et aI., 1985). Outcomes were compared to those for acute exposure to the TG-PCP and AGPCP formulations tested by Kerkvliet et al. (1 982a, b). The investigators found that a pCDD/pCDF fraction was significantly immunosuppressive. The oral dose at which 50% immunosuppression occurred (lDso) was calculated as 7.1, 85, and 208 Il-g/kg for HxCDD, HpCDD, and HpCDF, respectively. OCDD was not immunosuppressive at doses up to 500 Il-g/kg, the highest dose tested. For comparison, the investigators calculated an !Dso of 83 mg/kg for TG-PCP containing < 5 ppm HxCDD and 88 ppm HpCDD. AG-PCP was not immunosuppressive at doses up to 120 mg/kg, the highest dose tested. Co-administration of HxCDD and HpCDD produced an additive immunosuppressive effect, suggesting that toxic pCDD isomers present in TG-PCP act in concert. Co-administration of 60 mg/kg AG-PCP and 100 Il-g/kg HpCDD resulted in no increase over the degree of immunosuppression produced by HpCDD alone. For the TG-PCP tested, the full extent of immunosuppression observed in the splenic plasma-cell antibody
1500
CHAPTER 65
Pentachlorophenol
response assay could be explained by a combination of the pCDDs present as impurities. 65.9.3 IN VITRO EXPOSURE Holsapple et at. (1987) compared the effect of two grades of PCP on the spleen cell antigen-specific IgM response in spleen cells isolated from untreated female B6C3F1 mice. Spleen cells were cultured with one of three antigens in the presence of TGPCP or EC-7. The PCP concentration range is given in the text of the citation as 0.05-50 mg/culture and in Table 3 of the citation as 0.05-50 f.1g/culture. Both TG-PCP and EC-7 suppressed the immune response to one or more antigens at concentrations above 5 mg (or 5 f.1g) per culture. However, the investigators stated that immunosuppression was associated with markedly decreased viability in all cases, indicating that cytotoxicity interfered with the test. Thus, the results of this study cannot be interpreted as having demonstrated a specific immunotoxic effect of TG-PCP or EC-7. Lang and Mueller-Ruchholtz (1991) investigated the effect of in vitro exposure to TG or partially purified (Dowicide 7) PCP on the immunocompetence of peripheral blood lymphocytes (PBLs) isolated from 7-12 healthy human donors. The TG-PCP contained 2040 ppm pCDDs and pCDFs. The Dowicide 7 was >99% pure and contained 6 ppm pCDDs and pCDFs. Final concentrations of PCP in the lymphocyte culture media were in the range 2.7 to 54 mg/l (approximately 10 to 200 f.1M). Immunocompetence was measured with assays of mitogen responsiveness, interleukin-2 (IL-2) production, and stimulated (T-cell-dependent or independent) synthesis of IgM and IgG. Viability was found to be unaffected at PCP concentrations ::s40 f.1M. In vitro exposure of lymphocytes to TG-PCP resulted in stimulated synthesis of both IgM and IgG at concentrations ~1O or 20 f.1M. Dose-related suppression of IL-2 occurred only at concentrations ~80 f.1M, above the threshold for cytotoxicity. Similar results were obtained in experiments with Dowicide 7. The concordant in vitro results for Dowicide 7 and TG-PCP obtained by Lang and Mueller-Ruchholtz (1991) are in stark contrast with the results of animal studies showing that TGPCP but not AG-PCP suppresses antibody release and T-cell activation in response to in vivo antigen challenge (Exon and Koller, 1983; Kerkvliet et al., 1982a, b, 1985; Holsapple et at., 1987). However, Kerkvliet et at. (1982a) also found that exposure of mice to AG-PCP resulted in increased susceptibility of the spleen to viral tumorigenesis, indicating that isolated tests of humoral and T-cell-mediated immunity may have been inadequate to detect a more complex immunosuppressive response.
65.10 BIOCHEMICAL MECHANISMS TCHQ is clearly DNA-reactive and mutagenic whereas the weight of evidence suggests little or no genotoxicity for PCP. Because TCHQ is the major metabolite of PCP in rodents, it
seems plausible that the DNA reactivity attributed to PCP in rodent experiments (or in mammalian cells incubated with a microsomal extract of rat liver) is actually a consequence of TCHQ formation. The ability to promote oxidative DNA damage is typically regarded as being indicative of a potential to initiate carcinogenesis. PCP at 60 mg/kg given to male B6C3F1 mice as a single gavage dose resulted in significant elevation of 8-hydroxydeoxyguanosine (8-0H-dG) adducts in liver nuclear DNA at 6 hr but not at 24 hr (Sai-Kato et al., 1995). Curiously, TCHQ given to male B6C3F] mice as a single i.p. dose of 50 mg/kg produced no elevation of 8-0H-dG adducts in liver nuclear DNA at 6 hr (Dahlhaus et at., 1994). One possible explanation of these findings is that when TCHQ was administered i.p., adducts were formed rapidly and DNA repair was complete by the 6-hr mark, whereas when PCP was administered by gavage, the rate of absorption or metabolism to TCHQ slowed the rate of adduct formation, which was therefore incomplete at 6 hr postexposure. PCP is an uncoupler of oxidative phosphorylation in intact mitochondria, an effect observable at micromolar PCP concentrations (Weinbach, 1957). The ratio of phosphate to oxygen consumed (P : 0 ratio) was approximately halved when rat liver mitochondria were incubated in the presence of 5 f.1M PCP (Weinbach and Garbus, 1965), while a concentration of 0.5 f.1M produced a substantial increase in oxygen consumption consistent with the action of an uncoupler (Arrhenius et at., 1977). It is perhaps conceivable that mitochondrial uncoupling could play a role in the hepatotoxicity or immunotoxicity of PCP; this hypothesis has not been tested. PCP inhibits the binding of T4 to thyroid hormone receptors. An ICso of approximately 0.05 f.1M was found by van Raaij et at. (1991 a) for PCP displacement of T 4 from rat serum binding sites in vitro. Depression of serum T4 levels in PCP-exposed animals is likely to be a consequence of negative feedback originating from PCP occupation of the receptors. In a yeast strain engineered to express the human progesterone receptor (hPR), PCP antagonized the action of 10 nM progesterone with an ICso of approximately 0.7 f.1M. Binding of 1 nM synthetic progesterone to hPR was inhibited by PCP with an ICso of approximately 0.4 f.1M (Tran et at., 1996). PCP inhibits hepatic phenol sulfotransferase with high specificity and inhibits several other metabolizing enzymes and cytochrome P450 activities with lower efficiency. The ICso for PCP inhibition of 4'-hydroxypropranolol sulfation by canine hepatic cytosol was approximately 0.2 f.1M; in a partially purified phenol sulfotransferase fraction, the ICso was approximately 0.1 f.1M (Christ and Walle, 1989). An ICso value two orders of magnitude larger (approximately 20 f.1M) was found for PCP inhibition of l-chloro-2,4-dinitrobenzene sulfation by purified equine liver glutathione S-transferase (Moorthy and Randerath, 1996). ICso values of 12 to 25 f.1M were observed for PCP inhibition of O-acetylation of N-hydroxyarylamines in liver cytosol from hamster and rat (Shinohara et at., 1986). Higher 1Cso values have been observed for inhibition of N -acetyltransferase, N, O-acetyltransferase, and epoxide hydrolase
65.11 Human Health Effects
activities in liver extracts from one or more rodent species (Moorthy and Randerath, 1996; Shinohara et aI., 1986). Each of the hepatic enzyme activities inhibited by PCP is implicated in both activating and deactivating pathways in the metabolism of one or more class of xenobiotics. Studies in rats designed to elucidate potentially carcinogenic products of xenobiotic metabolism have found an inhibitory effect of PCP pretreatment or cotreatment on hepatic DNA adduct formation produced by N-hydroxy-2-acetylaminofluorene or dinitrotoluene (Kedderis et aI., 1984; Meerman et aI., 1981), hepatic unscheduled DNA synthesis produced by 2-acetylaminofluorene (Monteith, 1992), hepatic cytogenetic changes produced by safrole (Daimon et aI., 1997-98), initiation and promotion of hepatic foci and tumors produced by l'-hydroxysafrole (Boberg et aI., 1987), and growth of hepatic foci promoted by N -hydroxy-2-acetylaminofluorene (Kroese et aI., 1988). In mice, pretreatment with PCP inhibited DNA adduct and hepatoma formation produced by l'-hydroxy-2',3'-dehydroestragole (Fennell et aI., 1985). All of the anti-DNA-reactive and antitumorigenic actions of PCP described in this paragraph are likely to be explained, at least in part, by inhibition of sulfotransferase activity. The above evidence suggests that PCP at subcarcinogenic doses inhibits the metabolism of many xenobiotics to DNAreactive compounds and therefore would have a mitigating effect overall on the initiation of carcinogenesis by those compounds. On the other hand, i.p. administration of PCP (>99% pure) to mice at 20 mg/kg daily on each of the four days on which tamoxifen was administered increased the sulfotransferase-dependentDNA adduct formation produced by tamoxifen (Randerath et aI., 1994). Also, there is some evidence for a promotional effect of PCP that could, at sufficiently high doses, counteract the anti-initiating effect. Umemura et al. (1999) found that in male B6C3Fj mice fed 99% pure PCP at 300 ppm diet for 25 weeks following pre-initiation with diethylnitrosamine, promotion of liver tumors occurred (discussed in Section 65.5.2). When co-administered to rats chronically with 2-hydroxyethylnitrosourea, technical-grade (86% pure) PCP at 500 ppm diet (dose estimated as 22 mg/kg-day based on NTP, 1999) increased the incidence of acute myelocytic leukemia in males from 20% to 60% (Mirvish et aI., 1991). However, the technical-grade contaminants, including TCDD (25 ppb diet) and 2,3,7,8-tetrachlorodibenzofuran (TCDF, 670 ppb diet), might have been relevant to the observed promoting action. Promotional potential may also be indicated by the ability of PCP (>99% pure) to interfere with gap junctional communication in cultured rat liver epithelial cells (Sai et aI., 1998). In that study, the ICso for PCP was >40 J.!M (the highest concentration tested) for 4-hr incubation and 30-40 J.!M for 24-hr incubation. Although no statistical analysis was presented, the data shown indicated a threshold above 10 J.!M for 4-hr incubation and < 10 J.!M (the lowest concentration tested) for 24-hr incubation. The fact that the extent of inhibition increased between 4 and 24 hr suggests that the mechanism is indirect, perhaps a buildup of damage arising as a consequence of PCP inhibition of oxidative phosphorylation, buildup that
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might not occur in vascularized tissue in vivo. Therefore, in the event that PCP treatment in vivo is able to interfere with gap junctional communication, it is likely that much higher concentrations would be required to produce a given level of inhibition.
65.11 HUMAN HEALTH EFFECTS 65.11.1 SHORT-TERM EXPOSURE 65.11.1.1 Occupational Case Reports The occupational case report literature on PCP includes evidence that short-term inhalation of vapors in an unventilated area or prolonged dermal contact can result in hemotoxicity, including aplastic anemia and thrombocytopenic purpura (Hay and Singer, 1991; Rugman and Cos stick, 1990), toxicity to the nervous system, including excessive sweating, tachycardia, tachypnea, anorexia, fever, and death (Bergner et aI., 1965; Gordon, 1956; Menon, 1958), and hepatic and pulmonary damage combined with pancreatitis (Cooper and Macauley, 1982). 65.11.1.2 Nonoccupational Case Reports In a hospital nursery, poisoning of neonates occurred when diapers and bed linens were laundered with PCP. Twenty exposed infants manifested symptoms potentially associated with neurotoxicity (excessive sweating, tachycardia, tachypnea, respiratory distress, and metabolic acidosis); autopsy of the two fatal cases revealed hepatic and renal pathology (Armstrong et aI., 1969; Robson et aI., 1969). Three cases of dermal toxicity (pemphigus vulgaris and urticaria) were reported to be associated with short-term residential exposure to PCP-treated wood; serum PCP levels in the cases at presentation were 47-114 J.!g/l. 65.11.2 SUBCHRONIC AND CHRONIC EXPOSURE 65.11.2.1 Case Reports The case report literature on the toxicity of subchronic exposures to PCP is primarily indicative of hemotoxicity (Roberts, 1990). Roberts (1983) described six cases of aplastic anemia or red cell aplasia associated with exposure to PCP. In one of the cases of aplastic anemia, the patient, a 21-year-old male, had handled wet lumber processed with a commercial product containing 3% PCP and 1.5% tetrachlorophenol during the year prior to onset of clinical symptoms. Handling resulted in dermal and oral (hand-to-mouth) exposure to the wood preservative. The patient died of related causes 5 months after clinical onset. In another case, a 27-year-old male worked at building a log cabin one day per month. Using cloth gloves, he applied a 5% PCP solution with a brush after diluting a 40% solution with fuel oil. Symptoms were established within 9 months of the initial PCP exposure. Multiple medical measures were taken, but death occurred 6 months later. The other cases summarized
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by Roberts (1983) include a 24-year-old male exposed to PCP while working in construction who developed aplastic anemia and Hodgkin's disease 2 years after initial exposure, resulting in death; a 21-year-old male who applied PCP to furniture for 2 days, allowed the furniture to remain in his room, and developed aplastic anemia 1 month later; a 73-year-old male who repeatedly applied undiluted PCP to wood in a poorly ventilated barn, wearing no protective clothing, and developed pure red cell aplasia 4 years after initial exposure, resulting in death; and a 47-year-old male who dipped window frames into PCP at work and developed pure red cell aplasia 2 years after initial exposure, with death due to acute leukemia 2 years later. Rugman and Cos stick (1990) reported one case of fatal aplastic anemia in a 28-year-old male who applied PCP to timber over a period of months while renovating an old building. De Maeyer et al. (1995) described the cases of two women who suffered multiple miscarriages following introduction of PCP-treated wood products into their homes. Serum PCP levels recorded around the time of the miscarriages were :::62 IJ-g/I in one woman and :::31 IJ-g/I in the other. Roberts (1997) reported symptoms and laboratory findings for a family of five who had burned PCP-treated logs in their fireplace and whose home had been sided with PCPtreated lumber. The family reported intense irritation of the eyes and respiratory tract, recurrent infections, and neuropsychiatric symptoms. Laboratory findings included antinuclear antibody positivity, high serum rheumatoid factor titer, low serum complement 4, and elevated urinary porphyrins. 65.11.2.2 Case-Control Studies: Cancer In a case-control study, Pearce et al. (1986a) investigated whether the incidence of non-Hodgkin's lymphoma (NHL) in New Zealand was associated with the performance of occupational tasks entailing likely exposure to pentachlorophenol or other chlorophenols. The case population consisted of male patients with a confirmed diagnosis of NHL other than lymphosarcoma or reticulosarcoma during the years 1977-1981; the number of cases in the final sample was 83. Two sets of controls were chosen, one from the general population and one from patients with other types of cancer. A telephone questionnaire was used to elicit information concerning occupation and potential occupational exposures. The criterion selected for statistical significance was p < 0.1. The relative risk for fencing contractors was nonsignificantly elevated against the cancer controls and significantly elevated against the population controls; for the latter, the odds ratio (OR) was 6.1 and the 90% lower confidence limit (LCL) was 1.5. The relative risk for fencing as a farmer was significantly elevated against the cancer controls (OR = 1.9,90% LCL = 1.1) and nonsignificantly elevated against the population controls. For employment in the pelt department of a meat works (a source of potential exposure to 2,4,6,-trichlorophenol), the relative risk was significantly elevated against the popUlation controls (OR = 4.1, 90% LCL = 1.1) and non significantly elevated against the cancer controls. There was no increased relative risk of NHL for sawmill/timber
merchants or for persons in the category "potential chlorophenol exposure at sawmill or timber merchant." The investigators suggested that some nonchlorophenol exposure (such as a tumor virus) may have been responsible for the elevated NHL risk observed in meat workers. Pearce et al. (1987) expanded the case-control study of Pearce et al. (1986a) to include lymphosarcoma and reticulosarcoma diagnoses among the NHL cases; a total of 183 male NHL cases were included in the final sample. Controls were drawn from patients diagnosed with other types of cancer. The criterion selected for statistical significance was p < 0.1. The relative risk for fencing work was nonsignificantly elevated (OR = 1.4,90% LCL = 1.0), as was the relative risk for employment in the pelt department of a meat works (OR = 1.9,90% LCL = 0.9). For employment in any meat works job, the elevation was statistically significant (OR = 1.8, 90% LCL = 1.2). Once again there was no increased risk of NHL for sawmill/timber merchants or for persons in the category "potential chlorophenol exposure at sawmill or timber merchant." In both studies by Pearce and co-workers (1986a, 1987), sawmill workers were expected to have had the most PCP exposure but did not exhibit an elevated relative risk for NHL. Although NHL was found to be associated with fencing work, the investigators concluded that this finding was not relevant to any potential association between PCP and NHL. This can be understood from the investigators' observation that the principal method for preserving fencing timber in New Zealand is by vacuum-pressure impregnation with chromated copper arsenate and that since 1955, less than 1% of posts have been preserved with PCP. Although NaPCP was probably applied to some percentage of fencing-post timber as an antisapstain agent, any resultant exposure to fencing workers was still expected to be less than that for exposure at a sawmill or timber merchant (Pearce et al., 1986a). Nevertheless, it is difficult to assess the relevance of the fencing-work associations because neither absolute nor relative PCP exposure estimates were provided. It would be helpful also to have more information about the plausibility of a link between NHL and chromated copper arsenate. Hardell et al. (1981, 1994) reported results for 105 confirmed cases of NHL in males diagnosed in Umea, Sweden during the years 1974-1978. Living and deceased controls were drawn from national registries of population and death. A self-administered questionnaire was used to elicit information concerning chemical exposures and occupation. The criterion selected for statistical significance was p < 0.05. A non significantly elevated relative risk was found for sawmill workers (OR = 1.5,95% LCL = 0.7). A significantly elevated relative risk was found for high-grade PCP exposure, i.e., more than 1 week continuously or one month total (OR = 8.8; 95% LCL = 3.4). Pearce et al. (1986b) interviewed 76 male cancer patients in New Zealand who had been diagnosed with multiple myeloma (MM) during the years 1977-1981 and looked for associations with occupational exposures to chlorophenols. Controls were drawn from patients diagnosed with other types of can-
65.11 Human Health Effects cer. A telephone questionnaire was used to elicit information concerning occupation and potential occupational exposures. The criterion selected for statistical significance was p < 0.05. Relative risk for MM was nonsignificantly elevated for fencing work (OR = 1.6, 95% LCL = 0.9), fencing work with selftreated posts (OR = 1.1,95% LCL = 0.2), work at a saw mill or timber merchant (OR = 1.1,95% LCL = 0.5), potential occupational exposure to chlorophenols at a sawmill or timber merchant (OR = 1.4,95% LCL = 0.5), and meat works employment (OR = 1.3,95% LCL = 0.7). Smith et al. (1984) interviewed 82 male cancer patients in New Zealand diagnosed with soft-tissue sarcoma (STS) during the years 1976-1980 and looked for associations with occupational exposures to chlorophenols. Controls were drawn from patients diagnosed with other types of cancer. A telephone questionnaire was used to elicit information concerning occupation and potential occupational exposures. The criterion selected for statistical significance was p < 0.1. Nonsignificantly elevated relative risks for STS were found for fencing contractors (OR = 1.9, 90% LCL = 0.5), saw mill or timber merchants (OR = 1.3, 90% LCL = 0.6), employment in the pelt department of a meat works (OR = 4.7, 90% LCL = 0.6), and tannery or meat-works pelt department work (OR = 7.2, 90% LCL = 1.0). The only significantly elevated relative risk was for meat works employment (OR = 2.8, 90% LCL = 1.3). There was no elevation of relative risk linked to the job classification with the highest potential exposure to PCP, "potential chlorophenol exposure at saw mill or timber merchant." Eriksson et al. (1990) performed a case-control study based on confirmed cases of STS in males diagnosed in Uppsala, Sweden during the years 1978-1986. Assessment of exposure to PCP and other chlorophenols was determined by questionnaire and follow-up interview. The final case group size was 78 alive and 140 deceased. Controls were drawn from the national population registry. The criterion selected for statistical significance was p < 0.05. Relative risk was significantly elevated for highgrade exposure to PCP (i.e., at least 1 week continuously or at least 1 month total) in the absence of exposure to phenoxy acetic acid herbicides (OR = 3.9, 95% LCL = 1.2).
Potential Confounding by TCDD/pCDD Contamination In laboratory tests conducted in the late 1970s, TCDD levels in the range 0.25 to 1.1 ppb were measured in commercial PCP (IPCS, 1987). Thus, it is expected that this range approximates the TCDD contamination associated with PCP exposures discussed in case reports and epidemiological studies of that time period and earlier. Exposure to TCDD has been linked to STS in several epidemiologic studies. Fingerhut et al. (1991) examined a cohort consisting of essentially all U.S. chemical workers occupationally exposed to TCDD; a significantly increased risk of cancer mortality, mostly due to STS, was found in the subcohort with at least 1 year of exposure and at least 20 years latency. STS was found to be associated with self-reported exposures to TCDD and more highly chlorinated pCDDs in a Swedish case-control study (Eriksson et al., 1990). Although that study found larger odds ratios associated with exposures to
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PCP than to TCDD and higher pCDDs, the design did not allow assessment of exposure to TCDD and pCDDS among those with self-reported exposures to those agents relative to those with self-reported exposures to PCP.
65.11.2.3 Cohort Stndies Mortality Ramlow et al. (1996) conducted a mortality study of PCP manufacturing workers, part of a larger cohort of Dow Chemical Company workers in Michigan with potential exposure to pCDDs. The final PCP cohort consisted of 770 workers employed between 1940 and 1989, followed through 1989. Relative risks were based on comparison with age- and periodspecific death rates for white males in the U.S. population. In cohort members with at least 15 years latency, a nonsignificant elevation was observed for mortality from lymphopoietic cancer other than Hodgkin's disease, leukemia, or aleukemia (SMR = 2.0, 95% LCL = 0.54). The only statistically significant finding for the group with 15 years latency was for mortality from gastric or duodenal ulcer (SMR = 5.6, 95% LCL = 1.8). Dividing the entire cohort into low and highexposure categories revealed significantly elevated mortality in the high-exposure group for cancer of the kidney (SMR = 4.2, LCL = 1.4) and nonmalignant digestive system disease (including gastric or duodenal ulcer and liver cirrhosis). Also in the high-exposure group, mortality from lymphopoietic cancer other than leukemia was nonsignificantly elevated (SMR = 2.6, 95% LCL = 0.98). Immunotoxicity and Clinical Pathology Klemmer et al. (1980) performed a clinical study of wood-treatment workers, pest-control operators (PCOs), farmers, and nonoccupationally exposed controls in Hawaii. A single, cross-sectional measurement of blood PCP concentration was used as a surrogate measure of long-term exposure. The mean blood PCP concentration in the wood-treatment workers ("PCP-exposed") was 2.7 mgfl. The mean blood PCP concentration in the PCOs, farmers, and controls ("non-PCP-exposed") was 0.26 mgfl. Routine clinical testing of blood was performed; results were reported for only 45% (189/422) of the available study population, those for whom a complete set was available. Of these, 17 were wood-treatment workers, 155 were PCOs or farmers, and 17 were controls. In the PCP-exposed group there was significant elevation of the number of immature leukocytes (p < 0.005), plasma cholinesterase levels (p < 0.05), and alkaline phosphatase levels (p < 0.05); the observed values were within ranges considered normal. The study'S ability to detect an effect of PCP exposure was limited by the fact that the "nonPCP-exposed" population had serum PCP levels far higher than what can be considered background exposure. McConnachie and Zahalsky (1991) performed a crosssectional study on peripheral blood lymphocytes (PBLs) from 38 individuals in 10 families living in PCP-treated log homes in Indiana and Kentucky. The cohort consisted of 21 males and 17 females, 8-60 years of age (mean, 30 years). Exposure duration was 1-13 years (mean, 7.4 years). The elapsed
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time between the last exposure and performance of PBL testing was 0-9 years (mean, 4 years). There were 39 male and 81 female controls, 11-67 years of age. Controls had no known hematologic disorders or illness. No exclusions or data corrections were made for other potentially confounding factors such as smoking. The results of measuring antibody expression, T-cell activation, functional immunity, B-cell regulation, autoantibody expression, and natural killer cell function provided a reasonably coherent body of evidence suggesting that production of autoimmunity and depression of the T-cell proliferative response are associated with residence in log cabins treated with PCP. For the most part, the results were consistent with observations that T-cell response and general immunocompetence were suppressed in experimental animals treated with technicalgrade PCP manufactured in the same decade (1980s) that the cabins were constructed. The findings suggest an immunotoxic effect of subchronic exposure to technical-grade PCP that is not reversed upon removal of exposure. Information on the postexposure time course of the observed outcomes would aid in the interpretation of this study. Colosio et al. (1993) evaluated clinical and immunologic parameters in a cross-sectional study of 32 workers with prolonged exposure to PCP in a northern Italian wood factory. Of these, 14 subjects (the high-exposure group) had been engaged in brush application of PCP for at least 10 years. The remaining 18 subjects (the low-exposure group) had only indirect exposure to PCP in the work place. The control group consisted of 37 subjects who worked in a marble factory in the same valley. Mean plasma PCP concentrations in the highexposure, low-exposure, and control groups were 289,145, and 9 !-1g/l, respectively. Routine clinical chemistry testing of blood and urine was performed. Fasting levels of serum bile acids (SBAs) and urinary excretion of D-glucaric acid (a degradation product of SBAs), porphyrins, and 6-,B-hydroxycortisol were determined as additional measures of liver function. Tests of humoral and cellular immunity were also performed. In the high-exposure group, plasma levels of some SBAs were found to be significantly elevated relative to the controls, while the T-cell response to mitogen was significantly reduced relative to the low-exposure group as well as the controls. The results demonstrated a potentially adverse effect on the liver associated with long-term occupational exposure to technical-grade PCP. The observation of depressed T-cell response to mitogen was consistent with the results of treating animals with technicalgrade PCP. However, suppression of humoral immunity was invariably produced in the animal studies of technical-grade PCP while none was observed in this study. The finding in this study of depressed T-cell activation in occupationally exposed workers is consistent with a similar finding in residents of PCP-treated log homes, reported by McConnachie and Zahalsky (1991). Daniel et al. (1995) identified 188 patients with >6 months exposure to PCP-containing pesticides. The investigators, who were based in a hospital immunology department, did not indicate the circumstances under which the patients had originally presented themselves. Presumably the patients had symptoms
of suspected immune dysfunction, although those with chronic infections, rheumatic diseases, or other chronic diseases were excluded from the study population. In vitro responses to mitogenic and allogenic stimulation were determined in 163 patients, lymphocyte subpopulations in 157 patients, plasma neopterin levels in 118 patients, and plasma cytokine and cytokine receptor levels in 100 patients. By the criteria for impaired lymphocyte stimulation established within the study, a significantly greater fraction of patients with blood PCP > 10 !-1g/l exhibited impairment than those with blood PCP :s 10 fJ-g/I (91/133 vs. 15/30, p < 0.05). In addition, IL-8 levels were reported to correlate with blood PCP levels. Finally, all 11 patients with both impaired lymphocyte stimulation and an abnormally low CD4/CD8 ratio had blood PCP levels > 10 !-1g/l. Although the results of this study are intriguing, the presentation of data was far too incomplete to engender confidence in the results. Also, in the absence of a comparison group consisting of persons chosen from the general population, it is not possible to assess the clinical relevance of the study findings. These and other problems with the study were discussed by Triebig (1997). Neurotoxicity A longitudinal study of nerve conduction velocity (NCV) in the peripheral nervous system was performed by Triebig et al. (1987) in 10 chemical-company workers in Germany. The 7 men and 3 women had been in contact with PCP for a duration of 4-24 years (mean, 16 years). NCV values measured in 1980 and 1984 and were in the low-normal range; mean serum PCP levels were 368 !-1g/1 in 1980 and 346 !-1g/l in 1984. The only conclusion that can be drawn from the results is that four additional years of PCP exposure apparently did not diminish NCV in subjects with ongoing occupational exposure to PCP. The study did not address the question of whether PCP exposure was at all related to the clinically unremarkable finding of low-normal NCV values. Peper et al. (1999) identified 15 women who met the following criteria: residential exposure to wood-preserving chemicals for > 5 years, serum PCP > 25 !-1g/l, and serum lindane > 0.1 !-1g/l. This cohort of PCP-exposed women was drawn from a population of 2000 women presenting with infertility or gynecological problems. The exposure group was matched pairwise (for age, education, demographics, and education) with a control group of 15 women drawn from the same population. Subjective complaints were ascertained by questionnaire. Although a variety of the self-reported complaints surveyed were significantly higher in the exposure group, the investigators found that none was meaningfully correlated with serum PCP level. Standardized tests of cognitive function, psychomotor speed, attention, learning, and memory were administered. The PCP exposure group did significantly less well than the controls on several subtests. A rank correlation analysis of subtest score vs. serum PCP level revealed significant negative correlations for four subtests: reading speed, naming speed, paired associations, and visual retention. Unfortunately, the raw data were not supplied and the investigators apparently did not examine to data to discover whether an effect threshold could be
References discerned (i.e., a serum PCP level below which the correlations would be absent). Given the potential importance of the finding that neurobehavioral test scores were correlated with serum PCP, the study warrants repeating with a larger sample size, a higher ratio of controls to exposed, and a more detailed analysis of the relationship between serum PCP and the outcome variables. Endocrine Toxicity Gerhard et at. (1999) conducted a study of endocrine parameters in a cohort of 171 women presenting with infertility, a gynecological problem, or an endocrine problem with gynecologic ramifications. Serum PCP levels> 20 J.1g/1 were measured for 65 of the women; these were considered to be the exposed population. The control group consisted of 105 women with serum PCP < 20 J.1gll, matched for age, underlying condition, and geographical region. It is difficult to understand what was meant by matching on the basis of "underlying condition" because in the exposure group relative to the controls, primary infertility was under-represented (26% vs. 58%) while habitual abortions (22% vs. 8%) and alopecia (22% vs. 0.9%) were over-represented. The investigators reported that the exposed and control populations had significantly different levels of various hormones. However, all differences in mean values were small and almost certainly clinically unimportant. For example, mean (±SD) values of FSH were 7.8 ± 12.2 mE/ml in the exposure group and 8.1 ± 7.5 mE/ml in the controls, while the normal range was given as 1-10 mE/m!. Despite the low p value reported for the comparison (0.005), it is doubtful that the observed difference was meaningful. At a minimum, the investigators should have investigated the reproducibility of intergroup differences in measurements made at different times. Reproductive Toxicity Dimich-Ward et at. (1996) identified 19,675 births in British Columbia between 1952 and 1988 as the children of a cohort of 9512 fathers with potential exposure to PCP prior to the births. The fathers were part of a larger cohort of men who had worked at least one year in a sawmill where chlorophenate wood preservatives had been used. A nested case-referent (i.e., case-control) strategy was adopted. Five referents were matched for birth year; gender was either matched or treated as a covariate. The investigators indicated that the referents "were selected from all the offspring at risk when the case occurred." Presumably this meant all births recorded in British Columbia during the appropriate year. Chlorophenate exposure for three cumulative exposure periods (two preconception, one postconception) and maximal annual preconception exposure were estimated for each father for each birth. The investigators calculated odds ratios for five major indicators of reproductive health and 18 categories of congenital anomalies. The criterion for statistical significance was p < 0.05 (two-tailed test). None of the reproductive health indicators (prematurity, small size, low birthweight, stillbirth, or neonatal death) was significantly associated with any measure of chlorophenate exposure. Two of the congenital-anomaly categories were positively associated (p < 0.05) with one
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exposure measure: increased incidence of anenceephaly/spina bifida was associated with maximal preconception exposure (OR = 1.11) and increased incidence of genital-organ anomalies was associated with cumulative exposure during pregnancy (OR = 1.05). One of the congenital-anomaly categories was negatively associated (p < 0.05) with one exposure measure: decreased incidence of other limb was associated with maximal preconception exposure. Considering the number of comparisons made (18 x 4 = 72), 1 or 2 positive and 1 or 2 negative associations with two-tailed p values <0.05 would be expected to arise by chance (2.5% of 72 = 1.8). Thus, the above results are consistent with no effect of chlorophenates. However, congenital anomalies of the eye exhibited positive associations with three exposure measures, and two of these were highly significant (p < 0.005): increased incidence of eye anomalies with cumulative exposure 3 months prior to conception (OR = 2.01) and during pregnancy (OR = 1.21). Analysis by subcategories revealed that the associations pertained primarily to cataracts and the relative risk for cataracts was significantly greater in those with the most exposure (75th percentile) than those with the least (25th percentile). The results of Dimich-Ward et at. (1996) suggest that congenital cataracts may be related to paternal chlorophenate exposure occurring prior to conception and during pregnancy. If this relationship turns out to be real, it will be important to discover whether the causative agent is PCP, its technical-grade impurities, or some other chlorophenate. Despite the slant of the investigators, the study results do not necessarily indicate an effect on the male reproductive system. Alternative explanations exist, including the possibility that the causative agent contaminated the residential environment and was absorbed by the mother, impacting her reproductive function or affecting the fetus directly.
ACKNOWLEDGMENT The author evaluated many of the studies discussed in this chapter while a Staff Toxicologist in the Medical Toxicology Branch of the California Environmental Protection Agency's Department of Pesticide Regulation (DPR) during the years 1993 to 1996. She wishes to thank two former colleagues in the Medical Toxicology Branch, Charles N. Aldous, Ph.D. and Earl Meierhenry, D.V.M., PhD. for the valuable gift of their counsel. Dr. Aldous performed initial data review on the major PCP toxicology studies, identifying any deficiencies in methodology or in the reporting of results. At numerous times during the course of evaluating PCP toxicology studies, the author called upon Dr. Meierhenry to contribute his insight and expertise as a veterinary pathologist.
REFERENCES Ahlborg, U. G., Lindgren, J.-E., and Mcrcier, M. (1974). Metabolism of pentachlorophenol. Arch. Toxieol. 32, 271-281. Ahlborg, U. G., and Larsson, K. (1978). Metabolism of tetrachlorophenols in the rat. Arch. Toxicol. 40,63-74. Ahlborg, U. G., Larsson, K., and Thunberg, T. (1978). Metabolism of pentachlorophenol in vivo and in vitro. Arch. Toxieo!. 40, 45-53. Armstrong, R. w., Eichner, E. R., Klein, D. E., Barthel, W. F., et al. (1969). Pentachlorophenol poisoning in a nursery for newborn infants. n. Epidemiologic and toxicologic studies. 1. Pediatr. 75, 317-325.
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Sewall, C. H., Flagler, N., Vanden Heuvel, J. P., Clark, G. C., Tritscher, A. M., Maronpot, R. M., and Lucier, G. W. (1995). Alterations in thyroid function in female Sprague-Dawley rats following chronic treatment with 2,3,7,8tetrachlorodibenzo-p-dioxin. Toxieol. Appl. Pharmacol. 132,237-244. Shinohara, A. Saito, K., Yamazoe, Y., Kamataki, T., and Kato, R. (1986). Inhibition of acetyl-coenzyme A dependent activation of N -hydroxyarylamines by phenolic compounds, pentachlorophenol and I-nitro-2-naphthol. Chem. BioI. Interact. 60,275-285. Simmon, V. F., and Kauhanen, K. (1978). "In Vitro Microbiological Mutagenicity Assays of Pentachlorophenol." Final report, Stanford Research Institute (SRI) Project LSU-5612, U.S. Environmental Protection Agency, National Environmental Research Center, Cincinnati, OH. 13 pp. As cited in NTP, 1989 (op. cit.). Simmon, V. F., Riccio, E. S., and Peirce, M. V. (1979). "In Vitro Microbiological Genotoxicity Assays of Pentachlorophenol and 2,4,5-T Acid." Final report, Stanford Research Institute (SRI) Project LSU-7558. Sipes, G. 1., and Gandolfi, J. A. (1991). Biotransformation of toxicants. In "Casarett and Doull's Toxicology: The Basic Science of Poisons," (M. A. Amdur, J. Doull, and C. D. Klaassen, eds.), 4th ed., pp. 88-126. Pergamon, New York. Smith, A. H., Pearce, N. E., Fisher, D. 0., Giles, H. J., Teague, C. A., and Howard, J. K. (1984). Soft tissue sarcoma and exposure to phenoxyherbicides and chlorophenols in New Zealand. J. Natl. Cancer Inst. 73, 1111-1117. Tennant, R. w., Margolin, B. H., Shelby, M. D., Zeiger, E., Haseman, J. K, Spalding J., Caspary, W., Resnick, M., Stasiewicz, S., Anderson, B., and Minor, R. (1987). Prediction of chemical carcinogenicity in rodents from in vitro genetic toxicity assays. Science 236, 933-941. Tran, D. Q., Klotz, D. M., Ladlie, B. L., Ide, C. F., McLachlan, J. A., and Arnold, S. F. (1996). Biochem. Biophys. Res. Commun. 229, 518-523. Triebig, G., Csuzda, 1., Ktekeler, H. J., and Schaller, K. H. (1987). Pentachlorophenol and the peripheral nervous system: A longitudinal study in exposed workers. Brit. J. Indust. Med. 44,638-641. Triebig, G. (1997). Untitled letter. Arch. Environ. Health 52, 148. Uhl, S., Schmid, P., and Schlatter, C. (1986). Pharmacokinetics of pentachlorophenol in man. Arch. Toxicol. 58, 182-186. Umemura, T., Kai, S., Hasegawa, R., Sai, K., Kurokawa, Y., and Williams, G. M. (1999). Pentachlorophenol (PCP) produces liver oxidative stress and promotes but does not initiate hepatocarcinogenesis in B6C3F 1 mice. Carcinogenesis 20,1115-1120. United States Environmental Protection Agency (U.S. EPA) (1984a). Creosote, pentachlorophenol, and inorganic arsenicals; notice of intent to cancel; notice of determination; notice of availability of position document. Federal Reg. 49(136), 28666-28689; OPP-30000/28F; PH-FRL-2630-4. United States Environmental Protection Agency (U.S. EPA) (1984b). "Wood Preservative Pesticides: Creosote, Pentachlorophenol, and Inorganic Arsenicals," Position Document 4, Office of Pesticides and Toxic Substances, Washington. United States Environmental Protection Agency (U.S. EPA) (1986). "Creosote, Pentachlorophenol, and Inorganic Arsenicals; Amendment of Notice of Intent to Cancel Registrations; Notice." Federal Reg. 51(7), 1334-1348; OPP-30000128H; FRL-2952-6. United States Environmental Protection Agency (U.S. EPA) (1987). Pentachlorophenol; Amendment of notice of intent to cancel registrations. Federal Reg. 52(1), 140-148; OPP-30000/28M; FRL-3137-3. United States Environmental Protection Agency (U.S. EPA) (1988). "Recommendations for and Documentation of Biological Values for Use in Risk Assessment." EPN600/6-87/008, PB88-179874, Environmental Criteria and Assessment Office, Office of Health and Environmental Assessment, Cincinnati. United States Environmental Protection Agency (U.S. EPA) (199Ia). Federal Reg. 56 FR 3600 (1/30/91); 56 FR 3526 (1/30/91); 56 FR 30266 (7/1/91).
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United States Environmental Protection Agency (U.S. EPA) (199Ib). Federal Reg. 56 FR 3600 (1/30/91); 56 FR 30266 (7/1/91). United States Environmental Protection Agency, Science Advisory Board (U.S. EPA SAB) (1991). "Report of Science Advisory Board's Review of Issues Concerning the Health Effects of Ingested Pentachlorophenol," Environmental Health Committee, EPA-SAB-EHC-91-002. van Ommen, B., Adang, A. Muller, F., and van Bladeren, P. J. (1986). The microsomal metabolism of pentachlorophenol and its covalent binding to protein and DNA. Chem.-Biol. Interactions 60, 1-11. van Raaij., J. A. G. M., van den Berg, K. J., and Notten, W. R. F. (199Ia). Hexachlorobenzene and its metabolites pentachlorophenol and tetrachlorohydroquinone: Interaction with thyroxine binding sites of rat thyroid hormone carriers ex vivo and in vitro. Toxieol. Let!. 59, 101-107. van Raaij., J. A. G. M., van den Bcrg, K. J., Engel, R., Bragt, P. c., and Notten, W. R. F. (1991 b). Effects of hexachlorobenzene and its metabolites pentachlorophenol and tetrachlorohydroquinone on serum thyroid hormone levels in rats. Toxicology 67,107-116. van Raaij., J. A. G. M., Frijters, C. M. G., and van den Berg, K. J. (1993). Hexachlorobenzene-induced hypothyroidism: Involvement of different mechanisms by parent compound and metabolite. Biochem. Pharmacol. 46, 1385-1391. van Raaij., J. A. G. M., Frijters, C. M. G., Wong Yen Kong, L., van den Berg, K. J., and Notten, W. R. F. (1994). Reduction of thyroxine uptake into cerebrospinal fluid and rat brain by hexachlorobenzene and pentachlorophenol. Toxicology 94, 197-208. Villena, E, Montoya, G., Klaasen, R., Fleckenstein, R., and Suwalsky, M. (1992). Morphological changes on nerves and histopathological effects on liver and kidney of rats by pentachlorophenol (PCP). Comp. Biochem. Physiol. 101, 353-363. Waidyanatha, S., Lin, P.-H., and Rappaport, S. M. (1996). Characterization of chlorinated adducts of hemoglobin and albumin following administration of pentachlorophenol to rats. Chem. Res. Toxicol. 9,647-653. Wang, Y. J., and Lin, J. K. (1995). Estimation of selected phenols in drinking water with in situ acetylation and study on the DNA damaging properties of polychlorinated phenols. Arch. Environ. Contam. Toxicol. 28, 537-542. Weinbach, E. C. (1957). Biochemical basis for the toxicity of pentachlorophenol. Proc. Natl. Aead. Sci. USA 43, 393-397. Weinbach, E. C., and Garbus, J. (1965). The interaction of uncoupling phenols with mitochondria and with mitochondrial protein. J. BioI. Chem. 240, 1811-1819. Welsh, J. J., Collins, T. E X., Black, T. N., Graham, S. L., and O'Donnell, M. w., Jr. (1987). Teratogenic potential of purified pentachlorophenol and pentachloroanisole in subchronically exposed SpragueDawley rats. Fd. Chem. Toxic. 25, 163-172. Wester, R. c., Maibach, H. 1., Sedik, L., Melendres, J., Wade, M., and DiZio, S. (1993). Percutaneous absorption of pentachlorophenol from soil. Fundam. Appl. Toxicol. 20, 68-71. Witte, 1., Juhl, U., and Butte, W. (1985). DNA-damaging properties and cytotoxicity in human fibroblasts of tetrachlorohydroquinone, a pentachlorophenol metabolite. Mutation Res. 145,71-75. WHO (World Health Organization). (1987). "Pentachlorophenol. Environmental Health Criteria 71," IPCS, International Programme on Chemical Safety, World Health Organization, Geneva. Young, J. F., and Haley, T. J. (1978). A pharmacokinetic study of pentachlorophenol poisoning and the effect of forced diuresis. CZin. Toxicol. 12,41-48. Yuan, J. H., Goehl, T. J., Murrill, E., Moore, R., Clark, J., Hong, H. L., and Irwin, R. D. (1994). Toxicokinetics of pentachlorophenol in the F344 rat: Gavage and dosed feed studies. Xenobiotica 24, 553-560. Ziemsen, B., Angerer, J., and Lehnert, G. (1987). Sister chromatid exchange and chromosomal breakage in pentachlorophenol (PCP) exposed workers. Int. Arch. Occup. Environ. Health 59,413-417.
CHAPTER
66 Symmetrical and Asymmetrical Triazine Herbicides James T. Stevens, Charles B. Breckenridge Novartis Crop Protection
James Simpkins University of North Texas
J. Charles Eldridge Wake Forest University
66.1 INTRODUCTION
presence of the carbonyl group. Metribuzin adheres to the class by containing a thiomethyl-substituent on the ring. These herbicides are formulated into various water and lipid soluble products for commercial sales. Most of these formulations are identified with trade names and major manufacturers in Table 66.1.
Triazines have been used extensively as selective herbicides in agriculture in the United States and other parts of the world for more than 35 years (Stevens and Sumner, 1991). The triazines inhibit photosynthesis (Gysin and Knuesli, 1960). Even after more than three decades of use, certain of these triazine herbicides remain agronomically and commercially important, especially for the pre-emergent control of broadleaf weeds. They have become important "mixing partners" for many of the newer herbicides since they offer a broad spectrum of weed control (Gressel et aI., 1982). Because of their continued widespread use, the safety of this class of chemistry has been continually reassessed.
66.1.2 USES
66.1.1 CHEMISTRY AND FORMULATIONS
66.1.3 HAZARD IDENTIFICATION
These inhibitors of photosynthesis include the asymmetrical triazines or triazinones, such as metribuzin, and the symmetrical triazine herbicides. The major commercial symmetrical triazines are further divided into chloro-s-triazines: simazine, atrazine, propazine, cyanazine; the thiomethyl-s-triazines: ametryn, prometryn, terbutryn; and the methoxy-s-triazine prometon. The symmetrical triazines (s-triazines) have a chlorine, sulfur, or oxygen atom at the 2-position of the ring and are usually substituted in the 4- and 6-positions with alkylaminogroup (Fig. 66.1). Cyanazine contains a 2-cyano-isopropylamino-substituent at the 4-position on the ring. The asymmetrical triazine, metribuzin, retains the triazine ring, but since the nitrogen atoms are unevenly spaced, aromaticity is maintained by the
Hazards were identified for all the triazine and triazinone herbicides based upon protocols established under the Federal Insecticide, Fungicide, and Rodenticide Act (U.S. EPA, 1982) and the studies were conducted according to Good Laboratory Practices (U.S. EPA, 1979).
Handbook of Pesticide Toxicology Volume 2. Agents
Mode of action of the triazine and triazinone herbicides is inhibition of photosynthetic electron transport in most plants and generally have low toxicity to animals. The mode of action and uses of these chemicals are shown in Table 66.2.
66.1.4 ACUTE STUDIES Triazine herbicides are acutely relatively nontoxic, they are not remarkably irritating to the skin or eye, nor are they generally skin sensitizers (Table 66.3). The exceptions are atrazine, which is a skin sensitizer, and cyanazine, which is acute by the oral route.
1511
Copyright © 2001 by Academic Press. All right" ofreproduction in any form reserved
1512
CHAPTER 66
Triazine and Triazinone Herbicides S-Chloro-triazine
S-Thiomethyl-triazine
Others
S-
Cl
Na N
Na N
~NANRN-<
~NANRN--<
>Ylv
N,~S -
Ametryn
Atrazine Cl
N Na ~N~NRNr-
N
Metribuzin
---
...
S-
O-
Simazine Cl
NaN Na N Na N >-NANRN--< >-NANRN--< >-NANRN--< Propazine
Prometryn
Prometon
S-
Cl
N Na --\ANRN
±
+
Na N --\ANRN
Terbuthylazine
Terbutryn
Cl
...
. ..
...
NaN
~NANRN~C=N Cyanazine
Figure 66_1
Structures of symmetrical and asymmetrical triazine herbicides.
66.1.5 SUBCHRONIC/CHRONIC TOXICITY
The United States Environmental Protection Agency guidelines (U.S. EPA, 1982, 1998a) for subchronic studies for pesticides specify treatment of rats, mice, or dogs with the chemical for various lengths of time. Rat and mouse studies are 90 days in duration, and lifetime studies are typically 24 and 18 months,
respectively. In dogs, the studies are usually conducted for 90 days, 1 year, or 2 years. In all cases, animals are divided into test groups, 10 to 50 rats or mice and 4 to 6 dogs. At least four test groups are used in each study, one group receiving no chemical (controls) and three groups receiving low, medium, or high concentrations of the chemical in their diets. In these studies, urinalysis, hematology, and clinical chemistry
Table 66.1 Chemical Key Trade Name, Major Manufacturer, and Formulations for the Selected Triazines and Triazinone Major
Water soluble
Emulifiable
Chemical
Trade name
manufacturer
granule/powder
concentrate
Other
Ametryn
EVIK®
Novartis
80W
NAa
NA
Atrazine
AATREX®
Novartis
Nine-O
41 (90%)
NA
Cyanazine
BLADEX®
DuPont
NA
4L (90%)
NA
Metribuzin
SENCOR®
Miles
DF (75%)
4 (75%)
NA 5PS (5%)
Prometon
PRAMITOL®
Novartis
NA
25E (25%)
Prometryn
CAPAROL®
Novartis
WG(80%)
4L (90%)
NA
Propazine
Milo-Pro*
Griffin
80 WP (80%)
NA
NA
Simazine
PRlNCEp®
Novartis
80WG(80%)
4L (90%)
NA
Terbuthylazine
Gardoprim®
Novartis
WG
NA
NA
Terbutryn
Terbutrex*
Makhteshim Agan
50WP
L
NA
aNA = not available.
66.1 Introduction
1513
Table 66.2 Mode of Action and Uses for the Selected Triazines and Triazinone Mode of herbicide Chemical
Reference
action
Uses
Ametryn
Tomlin (1997a)
Selective systemic
Control of annual grasses and broad-leaved weeds in pineapples
Atrazine
Tomlin (1997b)
Selective systemic
Control of annual grasses and broad-leafs in corn, sorghum and sugar cane
Cyanazine
Tomlin (1997c)
Selective systemic
General weed control in beans, maize and peas
Metribuzin
Tomlin (1997d)
Selective systemic
Control of many grasses and broad-leaved weeds in soya beans, maize and corn
Prometon
Tomlin (1997e)
Non-selective systemic
Controls most grasses, many broad-leaved weeds and brush
Prometryn
Tomlin (1997f)
Selective systemic
Pre-emergent use in cotton, sun flowers, peanuts and vegetables
Propazine
TomIin (1997g)
Selective systemic
Preplant and pre-emergent use for weeds in sorghum
Simazine
Tomlin (1997h)
Selective systemic
Control of annual grasses and broad-leaved weeds in pome fruit, stone fruit, cane fruit, citrus, vines, strawberries, nuts, olives, pineapples, beans, peas, corn, aparagus, hops, alfalfa, coffee, rubber, oil palms, tea, and turef
Terbuthylazine
Tomlin (1997i)
Systemic
Broad-spectrum pre- or postemergence weed control in corn, sorghum, vines, fruit trees, citrus, coffee, oil palm, cocoa, olives, potatoes, peas, beans, sugar cane, rubber, and tree nurseries
Terbutryn
Tomlin (1997j)
Selective systemic
Pre-emergence weed control in cereals, sugar cane, and sunflowers; postemergence uses in cereals, sugar cane, and corn
parameters are evaluated, and gross and microscopic pathological examinations are performed on up to 50 tissues. Maximally tolerated doses are tested in order to demonstrate toxicity (up to 1000 mg/kg/day in the diets). In this fashion, it is possible to determine whether a chemical damages or alters any organ or tissue, and to establish levels of the chemical which produce no observable effects (the NOEL), and the lowest level at which effects are noted (the LOEL). The responses of repeated exposure of rats and dogs to the selected triazine herbicides are presented in Table 66.4. With the exception of rats and dogs fed terbuthylazine and cyanazine, the NOEL values were generally 2.5 mg/kg/day or higher, and LOEL values were 15 mg/kg/day or higher. The most common observation was not a specific organ or tissue,
but a reduction in body weight gain. Microscopically, the liver was the most common target organ. 66.1.6 DEVELOPMENTAL AND REPRODUCTIVE TOXICITY
Developmental toxicity studies, formerly called teratology studies, are required to be performed both in rats and rabbits; a two-generation reproduction study is conducted in rats (U.S. EPA, 1998a). The results of such studies conducted with the selected triazine herbicides are presented in Table 66.5. The triazine herbicides, with the exception of cyanazine, did not produce developmental or reproductive effects. Cyanazine produced developmental effects in the rat and rabbit at the high-
Table 66.3 EPA Acute Hazard Classification of the Technical Grade For Selected Triazine Herbicidesa
Triazine technicala Group s-Cl
Chemical
Eye
Skin
Irritation
Irritation
Oral
Dermal
Inhalation
LDSO (mg/kg)
LDsO (mglkg)
LCSO (mg/L)
Sensitization
Signal
potential
word Caution
Atrazine
Nonirritating
3090
>3100
>5
Sensitizer
Simazine
Nonirritating
Mild
>5000
>3100
>5.5
Nonsensitizer
Caution
Propazine
Mild
Nonirritating
>7000
>3100
>2.0
Nonsensitizer
Caution
Terbuthylazine
Nonirritating
Mild
1590-2000
>2000
>5.3
Nonsensitizer
Caution
Cyanazine
Nonirritating
Nonirritating
182-334
>2000
0.81
Nonsensitizer
Warning
Ametryn
Nonirritating
Nonirritating
1160
>2020
>5.1
Sensitizer
Caution
Prometryn
Slight irritation
Nonirritating
4550
>2020
>5.1
Nonsensitizer
Caution
Terbutryn
Nonirritating
Nonirritating
2500
>2000
>2.2
Nonsensitizer
Caution
s-OCH3
Prometon
Irritating
Mild
1518-4345
>2020
>3.2
Nonsensitizer
Caution
Asym.
Metribuzin
Nonirritating
Nonirritating
1090-1206
>20,000
>0.65
Nonsensitizer
Caution
s-SCH3
aThis table lists only technical products; formulated products used for agriculture may have more restrictive labeling due to the formulants used. Commercial formulations of prometon (Danger, Corrosive) and the 4L formulation of prometryn (Warning) carry more restrictive signal words due to formulants.
1514
CHAPTER 66
Triazine and Triazinone Herbicides
Table 66.4 Hazard Assessment for Repeat Exposure to the Selected Triazine Herbicides Triazine
mg/kg/day
Group
Chemical
s-CI
Atrazine Simazine Propazine Terbuty lazine Cyanazine Ametryn
Species/study
Target organ LOELb
Tissue or system
Rat/90-day oral
3.4
33
Body weight
Dog/52-week oral
7.5
25
Heart/myocardium
Rat/90-day oral
12.6
126
Body weight
Dog/90-day oral
7.5
134
Body weight
Rat/90-day oral
13
50
Body weight
Dog/90-day oral
7
25
Body weight
Rat/90-day oral
2.1
7.1
Dog/52-week oral
0.4
1.6
Body weight
Rat/90-day oral
1.5
15.0
Body weight
Dog/52-week oral
0.8
Rat/90-day oral
8.6
86
Dog/52-week oral
5.0
50
7.5
Body weight
Hematology effects Liver Liver
Prometryn
Rat/90-day oral
Terbutryn
Rat/90-day oral
50
140
Dog/26-week oral
10
25
Stomach Body weight
Dog/l04-week oral
Prometon Asymmetric
NOEL"
Metribuzin
50 3.7
>500 37.5
75% mortality Liver, kidney, bone marrow Body weight
Rat/90-day oral
5
15
Dog/52-week oral
5
20
Body weight
Rat/l04-week oral
5
15.0
Body weight, liver, kidney
Dog/I04-week oral
2.5
37.5
Body weight, liver, kidney
a No
observable effect level. bLowest observable effect level.
est doses tested. Effects noted at doses that were toxic to the mothers were cyclopia and diaphragmatic hernia in rabbits, and an apparent increase in the incidence of skeletal variations (i.e., anomalies) in rats (U.S. EPA, 1994).
66.2 MUTAGENICITY Weisburger (1975) noted that certain chemical carcinogens are capable of interacting directly with genetic material such as DNA. Based upon this association, several short-term tests to identify the alteration of genetic material or mutation were introduced into hazard testing for crop protection chemicals. These include tests to examine the possible (1) interaction with genes (gene mutation tests), (2) interaction with the chromosome (clastogenic tests), and (3) direct interaction with DNA (classified as other tests). The results for the selected triazine herbicides are presented in Table 66.6. All of the triazines were negative in the specific tests listed in Table 66.6. The overall mutagenic potentials of atrazine, simazine, and cyanazine have been reviewed by Brusick (1994), Hauswirth and Wetzel (1998), and Bogdanffy et al. (1999), respectively. The weight of the evidence indicates that these schloro-triazines are not mutagenic. Metribuzin is also negative in the standard battery of mutagenicity studies.
66.2.1 ONCOGENICITY ASSESSMENT The triazine and triazinone herbicides have been evaluated in lifetime animal bioassays. In these studies, groups of mice and rats are fed selected concentrations of the test chemical in their diet for 18 and 24 months, respectively. The levels of the test chemical administered in the diet are generally selected from repeat dose feeding studies at least 90-day in duration, and normally used to establish the NOEL, LOEL, and the maximum tolerated dose (MTD) (Farber, 1987). The MTD is defined as the highest concentration of test chemical that can be administrated without causing the death of the animal; often a 10% reduction in body weight gain has been used as criteria for establishing the MTD (Foran et aI., 1997). Following lifetime feeding of the chemical at the prescribed levels, veterinary pathologists microscopically examine approximately 50 tissues from each animal for the presence of tumors or other evidence of tissue damage. The results of such oncogenicity studies conducted in the mice are presented in Table 66.7. None of the selected triazines showed any evidence of the induction of tumors in either male or female mice despite high feeding levels, ranging from 76 to 1400 mg/kg/day doses, that were equal to or exceeding the MTD. The chloro-s-triazines, atrazine, cyanazine, propazine, and simazine, all resulted in either an increased incidence or an earlier onset of mammary tumors when administered to female S-D rats at high feeding
1515
66.2 Mutagenicity Table 66.5 Summary of the Results of Rat and Rabbit Developmental and a Two-Generation Rat Reproduction Study with Triazine Herbicides Triazine Group
Chemical
s-Cl
Atrazine
Simazine
Propazine
Terbuthylazine
Cyanazine
S-S-CH3
Ametryn
Prometryn
Terbutryn
S- OCH 3
Assymet.
Prometon
Metribuzin
Study/Species
Developmental/
Toxicity
reproduction
observed
mglkg/day
HDTa
Developmental/rabbit
None
Reproductive/rat
None
+Body wt. gain +Body wt. gain +Body wt. gain +Body wt. gain +Body wt. gain +Body wt. gain +Body wt. gain +Body wt. gain +Body wt. gain +Body wt. gain +Body wt. gain +Body wt. gain
Developmental/rat
Positive
t Malformations
Developmental/rabbit
Positive
t
Reproductive/rat
None
Developmental/rat
None
Developmental/rabbit
None
Reproductive/rat
None
Developmental/rat
None
Developmental/rabbit
None
Reproductive/rat
None
Developmental/rat
None
Developmental/rabbit
None
Reproductive/rat
None
Developmental/rat
None
LOELb
NOEU
700
70
10
75
75
5
35
35
600
300
30
200
75
5
30
6
600
100
50
50
5
30
30
5
5
5
15
3
75
5
3.5
0.6 10
0.4 0.3 >5
4
2
1
15
5
1.5
250
50
5
Body wt. gain
60
60
10
Malformations
Developmental/rat
None
+Body wt. gain +Body wt. gain
Developmental/rabbit
None
v
Reproductive/rat
None
+Body wt. gain
100
10
Developmental/rat
None
v
Body wt. gain
250
250
Developmental/rabbit
None
72
72
12
Reproductive/rat
None
5
5
>5
500
500
50
75
75
10
+Body wt. gain +Body wt. gain
>250
Developmental/rat
None
.;- Body wt. gain
Developmental/rabbit
None
+Ossified stemabra
Reproductive/rat
None
y
Body wt. gain
150
150
15
Developmental/rat
None
+Body wt. gain
360
120
36
Developmental/rabbit
None
v
Body wt. gain
25
25
Reproductive/rat
None
+Body wt. gain
75
25
Body wt. gain
100
100
>100
135
45
>45
Developmental/rat
None
Developmental/rabbit
None
Reproductive/rat
None
v
+Body wt. gain +Body wt. gain
15
3.5
7.5
1.5
a Highest
dose tested. bLowest observable effect level. cNo observable effect level. Table 66.6 Results of Mutagenicity Studies with the Selected Triazine Herbicides Triazine
Gene mutation Mouse
Clastogenic
DNA
Dominant
lymphoma
Micronucleus
repair
lethal Negative
Group
Chemical
Ames
REC
s-CI
Atrazine
Negative
Negative
Simazine
Negative
s-SCH3
Other
E. Coli
Negative
Negative
Negative
Negative
Negative
Negative
Propazine
Negative
Negative
Negative
Negative
Negative
Terbuthylazine
Negative
Negative
Negative
Negative
Negative
Cyanazine
Negative
Negative
Negative
Negative
Ametryn
Negative
Negative
Negative
Negative
Prometryn
Negative
Negative
Negative
Negative
Negative
Terbutryn
Negative
Negative
Negative
s-OCH3
Prometon
Negative
Negative
Negative
Asymmetric
Metribuzin
Negative
Negative
Negative
Negative
Negative Negative Negative
1516
CHAPTER 66
Triazine and Triazinone Herbicides
Table 66.7 Results of Carcinogenicity Studies in the Mouse Feeding level Triazine
(mg/kg/day)
Cancer
Group
Chemicala
potential
HDTb
s-CI
Atrazine
Negative
386e
Simazine
Negative
Propazine
Negative
544450e
Terbuthylazine
Negative
76 e
Other effects
Reference
t t
Body weight gain and thrombi in both sexes
Hauswirth and Wetzel (1998)
132
Body weight gain in both sexes
Hauswirth and Wetzel (1998)
15
450
Cardiac fibrosis and focal degeneration
IRIS (1997a)
15
76
t
Stevens et al. (1994)
LOEU
NOEU
38
38
5.7
Body weight gain in both sexes;
hematologic changes in males s-SCH3
s-OCH3
Asymmet.
1.4
Cyanazine
Negative
Ametryn
Negative
143 100e
0.5
50
Prometryn
Negative
429
1.4
143
Terbutryn
Negative
Prometon
Negative
42ge 1400e
Metribuzin
Negative
480
3.6
429
>429
70
700
120
120
t t t
Body weight gain in both sexes
V.S. EPA (1994)
Body weight gain in both sexes
Ahrens (1994a)
Body weight gain in both sexes
Ahrens (1994b)
No effects observed
Jessup (1980)
t
Body weight gain and
Ahrens (1994c)
t
mortality in both sexes
t
Hematocrit,
t
Liver, kidney and spleen weight
IRIS (1997b)
aCD 1 mice were tested for all chemical except terbuthylazine where Tif: MAGf was used. bHighest dose tested. CLowest observable effect level. dNo observable effect. eMaximum tolerated dose exceeded.
Table 66.8 Results of Carcinogenicity Studies in the Rat Triazine
Tumor
Group
Chemicala
response
s-CI
Atrazine
Mammary (S-D)a
Simazine
Mammary (S-D)
Feeding level (mg/kg/day) HDTb 50e 50e
NOEU
LOEU
Other effects
Reference
3.5
20
t t
Body weight gain in both sexes
Stevens et al. (1999)
Body weight gain in both sexes;
Stevens et al. (1994)
0.5
5.3
hematologic effects and
t mortality in females
s-SCH3
Propazine
Mammary (S-D)
Terbuthylazine
Negative (Tif: RAI)
Cyanazine
Mammary (S-D)
Ametryn
Negative (S-D)
50e 33 e 2.5 e 100e
5.8
50
0.2
1.1
0.2
1.0
2.5
25
t t t
Body weight gain in both sexes
Stevens et al. (1994)
Body weight gain in both sexes
Stevens et al. (1994)
Body weight gain in females;
V.S. EPA (1994)
t
hyperactivity in the males
t
Body weight gain in both sexes;
Stevens et al. (1994)
hematological effects in females Prometryn
Negative (S-D)
Terbutryn
Mammary, Thyroid,
75' 150e
5.0
37
0.1
15
t t
Body weight gain in both sexes
Stevens et al. (1994)
Body weight gain in both sexes
Stevens et al. (1994)
t t
Body weight gain in both sexes
Stevens et al. (1994)
Body weight gain;
IRIS (1997b);
t
liver and kidney weight;
Ahrens (1994d)
Liver (S-D) s-OCH3
Prometon
Negative (S-D)
75
Assymet
Metribuzin
Negative (Wistar)
15
25 5.0
15
uterine and mammary gland pathology aStrain of rat tested (S-D = Sprague-Dawley). bHighest dose tested. CLowest observable effect level. dNo observable effect level. eMaximum tolerated dose exceeded.
66.2 Mutagenicity
Estrogen LH Ovulatory LH Threshold
-""---
DDP E
E
3 months Age of Female
Figure 66.2 Representation of effect of normal reproductive aging on the ovulatory LH surge in S-D female rat.
levels, whereas male rats did not respond (Hauswirth and Wetzel, 1998). These finding are presented in Table 66.8. The thiomethyl- and methoxy-s-triazines, as well as asymmetrical triazine, metribuzin, were not carcinogenic, even at feeding levels exceeding the MTD; the exception was terbutryn where an increased incidence of mammary, thyroid, and liver tumors were observed in female S-D rats at feeding levels that exceeded the MTD. The female S-D rat has limitations in evaluating the effects of chemicals on the endocrine system because of the high degree of spontaneous tumor formation seen in the pituitary and mammary gland. At about 9 to 12 months of age the S-D rat begins to experience prolonged periods of estrus (Eldridge et aI., 1996; Simpkins et aI., 1998). Most laboratory rats, and S-D rats prior to 9-12 months, spend about 20 to 25% of their time in estrus. In the S-D rat, after this period, the animals spend increasing amounts of time in estrus. By 12 months of age, they often reach 40% of the time in estrus and eventually display persistent estrus (Eldridge et aI., 1998). This normal process of reproductive aging in the female S-D rat is depicted in Fig. 66.2. The failure of reproduction in the female S-D rat is most probably related to a deficient neuroendocrine control of gonadotropin releasing hormone (GnRH) secretion from the hypothalmus. With decreasing release of GnRH, the pituitary secretion of lutenizing hormone (LH) gradually decreases until it is inadequate to stimulate ovulation. As a result of the failed ovulation and the resulting prolonged periods of estrus, the rats experience prolonged exposure to estrogen and prolactin produced by the ovary and pituitary, respectively (Simpkins et aI., 1998). Both of these hormones are known to enhance the growth rate of mammary tumors in rats (Cutts and Noble, 1964). The reproductive aging process observed in the female S-D rat appears to be species specific. Other strains of rats, like the Fischer 344 rat, do not demonstrate this deficiency and do not have a high spontaneous incidence of mammary or pituitary tumors (Eldridge et aI., 1998). Detailed studies on atrazine have shown that Fischer 344 rats administered high doses of atrazine do not develop either an increased incidence or an early onset of mammary tumors (Thakur et aI., 1998; Wetzel et aI., 1994), unlike the find-
1517
ings noted in similarly treated female S-D rats (Hauswirth and Wetzel, 1998; Stevens et aI., 1994; Wetzel et aI., 1994). Furthermore, when the major internal source of estrogen was removed from the female S-D rats by surgical removal of the ovaries at the beginning of the study, no mammary tumors were found after 2 years of atrazine treatment (Stevens et al., 1999). Examination of the reproductive cycles of intact female S-D rats fed high doses of atrazine over their lifetimes showed that prolonged periods of estrus occurred earlier in the treated group than in control group (Hauswirth and Wetzel, 1998). Subsequent studies showed that high doses of atrazine administered to female S-D rats reduced the magnitude of the LH, resulting in a failure of ovulation to occur (Simpkins et aI., 1998). However, low doses of atrazine had no effect on the LH surge, the estrous cycle, or the rate of appearance of mammary tumors (Simpkins et aI., 1998), indicating that even in female S-D rats there is a threshold dose below which there are no adverse effects on reproductive processes. Finally, when high-dose atrazine-treated animals were given GnRH, the hormone that is responsible for triggering the LH surge, the LH surge was restored. This finding suggests that the pituitary LH releasing mechanisms function normally in atrazine-treated animals (Cooper et aI., 1995). Presumably then the effect of atrazine treatment at high doses is at the level of the hypothalamus. High doses of chloro-s-triazines appear to accelerate the development of mammary tumors in the S-D rat; This phenomenon occurs in a strain of rat that is already prone to spontaneously developing mammary tumors because of an inherent age-dependent deficiency in the regulation of the estrous cycle. The earlier appearance of mammary tumors in female S-D rats treated with high doses of atrazine is attributed to an increased exposure to endogenous estrogen and prolactin, secondary to the lengthening of the estrous cycle. Removal of endogenous estrogen in female S-D rats by ovariectomy prevents the appearance of mammary tumors, even in animals that have received high doses of atrazine for 2 years. In the S-D female there is a dose of atrazine (approximately 2.5 mg/kg) that has no effect on the estrous cycle or mammary tumor incidence and/or onset. The mammary tumor response to high doses of atrazine is unique to the female S-D rat; the response was not observed in male or female CD-l mice, Fischer 344 rats, or male S-D rats. 66.2.2 TOXICOLOGY IN HUMANS The reproductive aging process observed in female S-D rats is not relevant to human female since reproductive senescence in women is characterized principally as a ovarian failure. Human menopause commences with a decrease in endogenous estrogen, instead of an increase that is characteristic of the female S-D rat in a state of persistent estrus. The lack of relevance of these data to humans is supported by about 40 years of manufacturing and use history for atrazine and other triazine herbicides; To date there is no evidence linking atrazine ex-
1518
CHAPTER 66
Triazine and Triazinone Herbicides
posure to any human health effects (Loosli, 1995; Neuberger, 1996; Sathiakumur et aI., 1992).
66.2.3 SUMMARY RISK CHARACTERIZATION Evaluation of hazard profiles of the triazine herbicides revealed that these products are relatively nontoxic acutely, are well tolerated when administered to animals over a long duration of time, are not developmental or reproductive toxins, and are not mutagenic nor carcinogenic in mice or male rats. The exception is cyanazine, which is more acutely toxic, is weakly mutagenic, and results in developmental toxicity, presumably because of the presence of the cyanomoiety. The chloro-s-triazines appear to produce an earlier onset or an excess of mammary tumors in female S-D rats at high doses. Because of the unique nature of reproductive aging in female S-D rats, the carcinogenic response in this strain of rat is not considered relevant for human risk assessment.
REFERENCES Ahrens, W. H. (1994a). Ametryn. In "Herbicide Handbook," 7th ed., pp. 12-14. Weed Science Society of America, Champaign, IL. Ahrens, W. H. (1994b). Prometryn. In "Herbicide Handbook," 7th ed., pp. 245-247. Weed Science Society of America, Champaign, IL. Ahrens, W. H. (1994c). Prometryn. In "Herbicide Handbook," 7th ed., pp. 243245. Weed Science Society of America, Champaign, IL. Ahrens, W. H. (1994d). Metribuzin. In "Herbicide Handbook," 7th ed., pp. 200-203. Weed Science Society of America, Champaign, IL. Bakke, J. E., Larson, J. R., and Price, C. E. (1972). Metabolism of atrazine and 2-hydroxyatrazine by the rat. J. Agrie. Food Chem. 20, 603-607. Bogdanffy, M. S., O'Connor, J. C., Hansen, J. E, Gaddamidi, v., Van Pelt, C. S., and Cook, J. C. (1999). Chronic toxicity and oncogenicity bioassay in rats with the chloro-S-triazine herbicide cyanazine. J. Toxieo!. Environ. Health, Accepted for publication. Brusick, D. J. (1994). An assessment of the genetic toxicity of atrazine: Relevance to health and effects. Mutation Res. 317, 133-144. Cooper, R. L., Parrish, M. B., McElroy, W. K., Rehnberg, G. L., Hein, J. E, Goldmann, J. M., Stoker, T., and Tyrey, L. (1995). Effect of atrazine on the hormonal control of the ovary. The Toxicologist 15, 294. Cutts, J. H., and Noble, R. L. (1964). Estrone-induced mammary tumors in the rat-H. Effect of alteration in hormonal environment on tumor induction, behavior, and growth. Cancer Res. 24, 1124-1130. Eldridge, J. c., Stevens, J. T., Wetzel, L. T., Tisdel, M. 0., Breckenridge, C. B., McConnell, R. E, and Simpkins, J. W. (1996). Atrazine: Mechanisms of hormonal imbalance in female SD rats. Fund. App!. Toxieo!. 24(12), 2-5. Eldridge, J. c., McConnell, R. E, Wetzel, L. T., and Tisdel, M. O. (1998). Appearance of mammary tumors in atrazine-treated female rats: Probable mode of action involving strain-related control of ovulation and estrous cycling. In "Triazine Herbicides: Risk Assessment" (L. G. Ballantine, J. E. McFarland, and D. S. Hackett, eds.), Chap. 32, pp. 414-423. American Chemical Society, Washington, DC. Farber, T. M. (1987). "Pesticide Assessment Guidelines, Subdivision F, Position Document: Selection of a Maximum Tolerated Dose (MTD) in Oncogenicity Studies." Toxicology Branch, Hazard Evaluation Division, Office of Pesticides Programs, V.S. Environmental Protection Agency, NTIS PB88116736. Foran, J., and the ILSI Risk Science Working Group on Dose Selection. (1997). Principles for the selection of doses in chronic rodent bioassays. Environ. Health Perspeet. 105(1), 18-20.
Gressel, J., Ammon, H. V., Fogelfors, H., Gasquez, J., Kay, Q. O. N., and Kees, H. (1982). Discovery and distribution of herbicide-resistant weeds outside of North America. In "Herbicide Resistance in Plants" (H. M. LeBaron and J. Gressel, eds.), pp. 31-46. Wiley, New York. Gysin, H., and Knuesli, E. (1960). Chemistry and herbicidal properties of triazine derivatives. In "Advances in Pest Control Research" (R. Metcalf, ed.), Vo!. Ill, pp. 289-358. Wiley Interscience, New York. Hauswirth, J. w., and Wetzel, L. T. (1998). Toxicity characteristics of the 2-chlorotriazines, atrazine and simazine. In "Triazine Herbicides: Risk Assessment" (L. G. Ballantine, J. E. McFarland, and D. S, Hackett, eds.), pp. 370--383. American Chemical Society, Washington, DC. Jessup, D. C. (1980). "Terbutryn Technical Two-Year Carcinogenicity Study in Mice." Project Rep. 382-005, International Research and Development Corporation. Mattawan, MI (unpublished). IRIS (1997a). Propazino. CAS RN 139-40-2. EPA. http://Narero.epa.gov/ ngispgrn3/iris/substlO 187 .htrn IRIS (1997b). Metribuzin. CASRN 21087-64.9. EPA. http://Narero.epa.gov/ ngispgrn3/iris/substl. 0075htm Loosli, R. (1995). Epidemiology of atrazine. Rev. Environ. Contam. Toxieo!. 143,47-57. Neuberger, J. S. (1996). Atrazine and/or triazine herbicides exposure and cancer: An epidemiologic review. J. Agromed. 3(2), 9-30. Sathiakumur, N., Delzell, E., Austin, H., and Cole, P. (1992). A follow-up study of agricultural chemical production workers. Am. J. Ind. Med. 21,321-330. Simpkins, J. w., Eldridge, J. c., and Wetzel, L. T. (1998). Role of strain-specific reproductive patterns in the appearance of mammary tumors in atrazinetreated rats. In "Triazine Herbicides: Risk Assessment" (L. G. Ballantine, J. E. McFarland, and D. S. Hackett, eds.), Chap. 31, pp. 399-413. American Chemical Society, Washington, DC. Stevens, J. T., and Sumner, D. D. (1991). Herbicides. In "Handbook of Pesticide Toxicology" (w. J. Hayes and E. R. Laws, eds.), Vo!. 3, pp. 1317-1408. Academic Press, San Diego. Stevens, J. T., Breckenridge, C. B., Wetzel, L. T., Gillis, J., and Luempert, Ill, L. C. (1994). Hypothesis for mammary tumorigenesis in SpragueDawley rats exposed to certain triazine herbicides. J. Toxieo!. Environ. Health 43, 139-153. Stevens, J. T., Breckenridge, C. B., Wetzel, L. T., Thakur, A. J., Liu, c., Werner, C., Luempert, Ill, L. c., and Eldridge, J. C. (1999). A risk characterization for atrazine: Oncogenicity profile. J. Toxieo!. Environ. Health 56,69-109. Sumner, D. D., Luempert, Ill, L. c., and Stevens, J. T. (1995). Agricultural chemicals: The impact of regulation under FIFRA on science and economics. In "Primer on Regulatory Toxicology" (C. Chenzelis, J. Holson, and S. Gad, eds.), pp. 133-163. Raven Press, New York. Thakur, A. J., Wetzel, L. T., Voelker, R. w., and Wakefield, A. E. (1998). Results of a two-year oncogenicity study in Fischer 344 rats with atrazine. In "Triazine Herbicides: Risk Assessment" (L. G. Ballantine, J. E. McFarland, and D. S. Hackett, eds.), pp. 384-398. American Chemical Society, Washington, DC. Tomlin, C. D. S. (1997a). Ametryn. In "A World Compendium: The Pesticide Manual," 11th ed., pp. 35-36. British Crop Protection Council, Famham, Surrey. Tomlin, C. D. S. (1997b). Atrazine. In "A World Compendium: The Pesticide Manual," 11th ed., pp. 55-57. British Crop Protection Council, Famham, Surrey. Tomlin, C. D. S. (l997c). Cyanazine. In "A World Compendium: The Pesticide Manual," 11th ed., pp. 280--282. British Crop Protection Council, Famham, Surrey. Tomlin, C. D. S. (1997d). Metribuzin. In "A World Compendium: The Pesticide Manual," 11th ed., pp. 840--841. British Crop Protection Council, Famham, Surrey. Tomlin, C. D. S. (1997e). Prometon. In "A World Compendium: The Pesticide Manual," 11th ed., pp. 1010--1011. British Crop Protection Council, Famham, Surrey. Tomlin, C. D. s. (1997f). Prometryn. In "A World Compendium: The Pesticide Manual," 11th ed., pp. 1011-1012. British Crop Protection Council, Famham, Surrey.
References
Tomlin, C. D. S. (1997g). Propazine.In "A World Compendium: The Pesticide Manual," 11th ed., pp. 1024-1026. British Crop Protection Council, Farnham, Surrey. Tomlin, C. D. S. (1997h). Simazine.In "A World Compendium: The Pesticide Manual," 11 th ed., pp. 1106-1108. British Crop Protection Council, Farnham, Surrey. Tomlin, C. D. S. (1997i). Terbuthylazine. In "A World Compendium: The Pesticide Manual," 11th ed., pp. 1168-1170. British Crop Protection Council, Farnham, Surrey. Tomlin, C. D. S. (1997j). Terbutryn.In "A World Compendium: The Pesticide Manual," II th ed., pp. 1170-1172. British Crop Protection Council, Farnham, Surrey. V.S. Environmental Protection Agency (EPA) (1979). "Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA); Good Laboratory Practice Standards-Final Rules. 40 CFR Part 160." Fed. Reg. 44(91),27362-27407. V.S. Environmental Protection Agency (EPA) (1982). "Pesticide Assessment Guidelines, Subdivision F. Hazard Evaluation: Human and Domestic An-
1519
imals." Environmental Protection Agency-540/9-82-025. Available from NTIS, Springfield, VA. V.S. Environmental Protection Agency (EPA) (1994). Atrazine, simazine, and cyanazine; Notice of initiation of special review. Fed. Reg. 59(225), 6041260443. V.S. Environmental Protection Agency (EPA) (1997). Propazine. Pesticide tolerance petition filing. Fed. Reg. 63(193), 53657-53660. V.S. Environmental Protection Agency (EPA) (1998a). "Pesticide Assessment Guidelines," http://www. epa. gov/opptsfrs/OPPTS Harmonizedl870 Health Effects Test Guidelines /Series (accessed 3/99). V.S. Environmental Protection Agency (EPA). (1998b). Prometryn: Pesticide tolerances. Fed. Reg. 63(37),9494-9499. Weisburger, J. H. (1975). In "Toxicology, The Basic Science of Poisons" (L. J. Casarett and J. Duoll, eds.), p. 333. MacMillan, New York. Wetzel et al. (1994).
CHAPTER
67 Phenylurea Herbicides Jing Liu Oklahoma State University
67.1 INTRODUCTION Substituted phenylurea herbicides are a group of pesticides used for general weed control in agricultural and nonagricultural practices, for example, along railroads, utilities' rights-of-way, and in industrial areas. The first phenylurea herbicide, N,Ndimethyl-N'-( 4-chlorophenyl)-urea, was introduced in 1952 by Du Pont under the common name of monuron. In subsequent years, many more derivatives of this class of compounds have been marketed. The phenylurea herbicides are now manufactured and distributed under the names of anisuron, buturon, chlorbromuron, chlortoluron, chloroxuron, difenoxuron, diuron, fenuron, fluometuron, isoproturon, linuron, methiuron, metobromuron, metoxuron, monuron, neburon, parafluron, siduron, tebuthiuron, tetrafluron, and thidiazuron. Of these, diuron and fluometuron are the most commonly used in the United States. Isoproturon has been reported to be the most widely used phenylurea herbicide in other countries, for example, India. The herbicidal action of these compounds is based on their ability to inhibit photosynthesis. Typical phenylurea herbicides, such as diuron, fluometuron, and isoproturon, are photosystem 11 inhibitors. Photosystem 11 is a multisubunit enzyme complex which uses light energy to catalyze the photooxidation of water to reducing equivalents and oxygen. The reaction center in photosystem 11 is composed of the proteins Dl, D2, CP43, CP47, and the light-harvesting complex 11 (Rhee et aI., 1998). Substituted phenylurea herbicides inhibit photodependent electron transfer by binding to the Dl protein (Arnaud et aI., 1994). Degradation of phenylurea herbicides in nature can be a relatively slow process. These pesticides can be decomposed by ultraviolet (UV) irradiation or by acidic or alkaline conditions. There are four basic types of reactions in the photochemistry of the substituted phenylurea herbicides; they are photolysis of the -C-X bond on aromatics (X = Cl, Br), photoeliminations (Norrish-Type 11 reactions), photooxidations, and photorearrangements (Kotzias and Korte, 1981). Biological degradation of the compounds in plants and soil is carried out by microflora and microfauna. Vroumsia and co-workers (1996) reported that Rhizoctonia solani (agonomycetes) was the most efficient microorganism tested at Handbook of Pesticide Toxicology Volume 2. Agents
degrading diuron, chlortoluron, and isoproturon. The basic reactions of the biological metabolism of phenylurea herbicides are N-demethylation followed by oxidation of the aromatic ring. The compounds are gradually transformed by microorganisms to 3-arylureas, which are then metabolized to arylamines, carbon dioxide, and ammonia (Cernakova, 1995; Engelhard et aI., 1972). Human cytochrome P450 3A4 (CYP3A4) expressed in yeast has been reported to catalyze the metabolism of chlortoluron (Mehmood et aI., 1995). The metabolism was absolutely dependent on NADPH. Chlortoluron was degraded by CYP3A4 into four major metabolites, hydroxylated-N-monodemethylated, hydroxylatedring methylated, N-didemethylated, and N-monodemethylated products. Other cytochrome P450-mediated reactions have not been reported. While over 20 different phenylureas have been marketed for use as herbicides, little information is available on the toxicity of most of these compounds. More specific information on three of the most common phenylureas, diuron, fluometuron, and isoproturon, is provided.
1521
67.2 DIURON Synonyms: N' -(3,4-Dichlorophenyl)-N,N-dimethylurea; 3-(3,4-Dichlorophenyl)-I, 1dimethylurea; DCMU; DMU Chemical abstract number (CAS#): 330541 Molecular formula: C 9 HIQCI,NzO (233.1) Chemical structure:
Trade names and available formulations: Karmex, Karmex DL, Diuron 80WP, Diuron 4L. Direx 4L, Dj-on, Diurex, Duirol, Dailon, Rout, Diater, Unidron, Crisuron, and Cekiuron.
Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
1522
CHAPTER 67 Phenylurea Herbicides
67.2.1 PHYSICAL AND CHEMICAL PROPERTIES
Diuron is a white, odorless, crystalline solid with a melting point of 158-159°C, and a boiling point of 180-190°C. Diuron has a water solubility of about 42 ppm (mg/l) at 25°C. Under room temperature and neutral pH, hydrolysis of diuron is negligible. Diuron is stable to oxygen and moisture (Worthing, 1983). 67.2.2 USAGE
Diuron was introduced in 1954 by E. I. Du Pont de Nemours & Co. (Inc.) under the trademark "Karmex" and is mainly used as a pre-emergence herbicide for general weed control on noncrop lands. It is also used as a soil sterilant. The industrial use of diuron, for example, along railroad rights-of-way, represents 57% of the total usage. As of 1995, the estimated usage of diuron in the United States was about 2-4 million pounds (Aspelin, 1997). Diuron is also used selectively before emergence on crops such as asparagus, citrus, pineapple, sugarcane, and cotton. Tolerance for diuron residues was established at 1 ppm on fruits and vegetables in Canada (Chapman, 1967). The occupational exposure limit (TWA, i.e., 8-hourtime-weighted average) for diuron in workplace air was established at 10 mg/m 3 by ACGIH in the United States (International Labor Office, 1980), which indicates that occupational intake at a rate of 1.4 mg/kg/day is considered safe (Stevens and Sumner, 1991). 67.2.3 ABSORPTION, METABOLISM, AND EXCRETION
1967). It has been reported (Boyd and Krupa, 1970) that protein content in the diet can influence the acute toxicity of diuron. For example, the LD50 of diuron in weanling rats fed a proteindeficient diet (i.e., about 14% of the normal protein intake) was 0.4 ± 0.1 g/kg, whereas the LD50 was 2.4 ± 1.4 g/kg in weanling rats fed a protein-enriched diet. Weanling rats fed normal laboratory chow exhibited intermediate sensitivity (LD50 = 1.0 ± 0.2 g/kg). Signs of acute toxicity following near-lethal dosages of diuron in rats included drowsiness, ataxia, decrease and subsequent increase in reflexes, irritability, and bradypnea. Diarrhea, diuresis, shedding of bloody tears, and nosebleed were also noted. Animals treated with diuron exhibited significant loss of body weight, along with a decrease in food and water intake. Hypothermia, glucosuria, proteinuria, and aciduria were detected at 24 hours after exposure. Respiratory failure was the immediate cause of death. The intensity of the signs of toxicity was dose-dependent. Recovery in surviving animals began at 24 hours, and most of the signs disappeared by 72 hours after exposure (Boyd and Krupa, 1970). No signs of skin irritation or sensitization were noted in dermally exposed guinea pigs (Hodge et aI., 1967). Pathological examination revealed local gastroenteritis, gastric ulcers, and capillary-venous congestion of the gastrointestinal mucosa. Stress reactions were also seen in the adrenal and thymus glands and in the spleen. Young animals may exhibit signs of kwashiorkor, such as developmental retardation of the adrenal and thymus glands, gastrointestinal tracts, and especially the testes (Boyd and Krupa, 1970). Other pathological changes included an enlarged, congested spleen (Hodge et aI., 1967). 67.2.4.2 Subacute Toxicity
Diuron is readily absorbed through the gastrointestinal tract in rats and dogs. Tissue levels of diuron were positively correlated with dosage. No apparent storage of diuron in tissues was noted (Hodge et aI., 1967). In mammals, diuron is mainly metabolized by dealkylation of the urea methyl groups. Hydrolysis of diuron to 3,4-dichloroaniline and oxidation to 3,4dichlorophenol as well as hydroxylation at carbon 2 and/or 6 of the benzene ring, have also been reported. The predominant metabolite of diuron in urine was N -(3,4-dichlorophenyl)-urea. Diuron is also partially excreted unchanged in feces and urine (Boehme and Ernst, 1965; Hodge et aI., 1967). Metabolites found in mammals were qualitatively similar to those found in soil and plants wherein dealkylation was also the major metabolic pathway (Dalton et aI., 1966; Geissbuhler et aI., 1963). 67.2.4 TOXICITY TO LABORATORY ANIMALS 67.2.4.1 Acute Toxicity
The oral LD50 (14 days) for diuron in male rats was 3.4 glkg with 95% confidence limits of 2.9-4.0 g/kg (Hodge et aI.,
Rats treated with 1 g/kg of diuron daily for 10 days did not show any lethality but exhibited weight gain retardation. Pathologic changes in the spleen and bone marrow were noted at autopsy 3 and 11 days after the final dose (Hodge et aI., 1967). 67.2.4.3 Subchronic and Chronic Toxicity
Rats fed a diuron-containing diet (200, 400, 2000, 4000, or 8000 ppm) for 30 days showed growth retardation and anemia at dosages of 4000 and 8000 ppm. Red blood cell counts and hemoglobin levels were also reduced. Lethality occurred only at 8000 ppm. Congestion and an increase in spleen weight were also observed with the highest dose (Hodge et aI., 1967). In 90-day feeding studies (Hodge et aI., 1967), 50 and 250 ppm diuron did not cause any toxic effects in rats of either sex: Female rats treated with 500 ppm diuron showed cyanotic discoloration and less weight gain, while the males exhibited literally no signs of toxicity. Reductions in red blood cell counts and hemoglobin levels accompanied by a compensatory bone marrow hyperplasia were observed in rats fed 2500 or 5000 ppm diuron. Growth retardation and decreased food consumption were also noted. All these effects of diuron were greater in females than in males.
67.2 Diuron Two-year feeding studies in rats and dogs revealed no significant adverse effects at the dietary levels of 25, 125, or 250 ppm diuron except for an inconsistent and sporadic slight anemia. Growth depression was seen with 250 ppm diuron (Hodge et aI., 1967). 67.2.4.4 Hematotoxicity Female Sprague-Dawley rats fed 250, 500, or 1000 mg/kg diuron in the diet for 14 months exhibited biochemical and morphological changes in the circulatory system (Wang et aI., 1993). Relative spleen weight was significantly increased in a dose-dependent manner. Hemoglobin levels and erythrocyte counts were significantly reduced, while methemoglobin concentration and white blood cell counts were increased. Hemoglobin adduct of the released parent aromatic amine, 3,4DCA, was detected at dose-related levels in animals fed 500 or 1000 mg/kg diuron. Increased pigmentation (hemosiderin) in the spleen was seen histologic ally, reflecting a response to the hemolytic anemia and methemoglobinemia induced by the herbicide. Morphological examination of the red blood cells revealed changes such as erythrocytes with the shape of a spindle or with a centrally stained area associated with abnormal hemoglobin, polychromatic erythrocytes, and hypochromic erythrocytes with a large area of central pallor presumably due to the decreased hemoglobin content. 67.2.4.5 Genotoxicity and Carcinogenesis It has been reported that a single dose of 170 or 340 mg/kg
diuron given intraperitoneally induced the formation of micronuclei in bone marrow cells in mice at 30 and 48 hours after the treatment (Agrawal et aI., 1996). Seiler (1978), however, reported that diuron was incapable of inducing micronuclei in erythrocytes when given as a single dose by gavage at 1 or 2 g/kg in mice. In other mutagenicity tests, such as the testicular DNA synthesis inhibition test and the Ames test, diuron exhibited mutagenic activity (Seiler, 1978). Diuron was a suspect genotoxicant, directly or after S-9 activation, at the lowest detected concentrations (LCD) of 900 or 112.5 Ilgll, respectively, in the Vibrio jischerifMutatox™ test (Canna-Michaelidou and Nicolaou, 1996). Antony and co-workers (1989) reported that topical application of diuron at the rate of 250 mg/kg, three times a week for 3 weeks followed by multiple application of a known skin tumor promotor (12-0-tetradecanoyl phorbol 13-acetate, TPA) initiated neoplastic transformation and development of skin tumors in mice. Multiple skin applications of diuron alone for up to 52 weeks, however, did not show any tumor-inducing activity. There was no evidence in two-year bioassays that diuron was carcinogenic in rats or dogs (Hodge et aI., 1967). Mice given 464 mg/kg diuron daily from 7 to 28 days of age followed by 1000 ppm diuron daily in the diet for 18 months showed no signs of increased tumor formation (Reinhold, 1987).
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67.2.4.6 Teratogenicity In a study in which the formulation Karmex® (containing 80% diuron) was given by gastric intubation to pregnant rats from gestation days 6-15 at levels of 125, 250, or 500 mg/kg/day, only the highest dosage reduced both maternal and fetal body weights. Wavy ribs were seen at the dosages of 250 and 500 mg/kg, and delayed ossification of the calvarium was noted in fetuses whose dams received 125 mg/kg diuron (Khera et aI., 1979). Diuron showed no teratogenic activity in mice, however (Reinhold, 1987). A multigeneration reproductive toxicity study in rats given 125 ppm diuron in the diet revealed no significant changes in reproductive performance endpoints. Post-weaning growth of the F2b and F3a generations was moderately affected, however (Hodge et aI., 1967). 67.2.4.7 Biochemical Effects Diuron, a dihalogenated substituted urea herbicide, has been reported to be a more potent inducer of hepatic metabolizing enzymes [e.g., benzo(a)pyrene mono-oxygenase (BP-MOO), 7-ethoxycoumarin O-deethylase (ECOD), and 7-ethoxyresorufin O-deethylase (EROD)] compared to those phenylurea herbicides with one or no halogen substitutions, for example, chlortoluron and isoproturon (Schoket and Vincze, 1985, 1986, 1990). Schoket and co-workers (1987) found that repeated diuron exposures (116 LDso for 3 days) decreased the plasma halflife of antipyrine significantly, indicating hepatic cytochrome P4S0 isozymes were induced. Close correlations (r = 0.980.99) were found between the induction of BP-MOO, ECOD, and EROD and the increase of antipyrine metabolism after diuron treatment. Moreover, hepatic enzymes such as cytochrome P4S0, BP-MOO, microsomal epoxide hydrolase, glutathione S-transferase, and UDP-glucuronyltransferase were all induced by diuron in a dose-related manner (oral dosing 1120 to 114 LDso) in rats (Schoket and Vincze, 1990). Dose-related induction of hepatic microsomal enzymes was also seen in rats fed a diuron-containing diet (100, 250, 500, 1000, and 2000 ppm) for 13 weeks (Kinoshita and DuBois, 1970). Maximum induction occurred within the first 3 weeks of feeding and then decreased afterwards. Moreover, a sex difference was noted in the response of the animals, that is, male rats were more sensitive than females to the enzyme-inducing activity of diuron. 67.2.4.8 Aquatic Toxicity Diuron used for weed control in water may interfere with the growth of fish and food-chain microfauna, such as Daphnia (Crosby and Tucker, 1966). The LCso of diuron in Daphnia magna or Daphnia pulex at 24 or 48 hours is 1.4 mg/1. The LCsos of diuron in the warm-water fish Lepomis macrochirus and the cold-water fish Oncorhynchus kisutch are 7.4 and 16.0 mg/l at 48 hours, respectively (Ramamoorthy and Baddaloo, 1995).
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CHAPTER 67
Phenylurea Herbicides
67.3 FLUOMETURON
67.3.3 TOXICITY TO LABORATORY ANIMALS 67.3.3.1 Acute Toxicity
Synonym: 1,I-dimethyl-3-[3-(trifluoromethyl)phenyIJurea Chemical abstract number (CAS#): 2164-17-2 Molecular formula: ClOHllF3N20 (232.2) Chemical structure:
Trade names and available formulations: Cotoran, Cotoran 4L, Cotoran 85DF, Meturon 4L, Ciba 2059, Cottonex, Lanex, Herbicide C-2059, Pakhtaran
67.3.1 PHYSICAL AND CHEMICAL PROPERTIES
Fluometuron is a colorless, crystalline, sandlike material with a melting point of 163-164.5°C and a vapor pressure of 5 x 10-7 mm Hg at 20°e. Its solubility in water and acetone at 20°C is about 105 mg/l (ppm) and 105 g/l (ppt), respectively. Fluometuron is soluble in most organic solvents (lARC Monographs, 1983; Worthing, 1983).
67.3.2 USAGE
Fluometuron has been used as a herbicide in the United States for more than three decades. Since being introduced as a commercial chemical in 1960 by Ciba-Geigy AG under the trademark "Cotoran" (Worthing, 1983), fluometuron has been widely used to control broadleaf weeds and grasses on agricultural crops, for example, cotton and sugarcane. The amount of fluometuron used in the United States in 1976 and 1978 was estimated to be 5.3 and 2.9 million pounds, respectively (IARC Monographs, 1983). As of 1995, the approximate quantity of fluometuron used annually in U.S. agricultural practices was about 5-9 million pounds (Aspelin, 1997). The tolerance for fluometuron in or on raw agricultural commodities, cottonseed, and sugarcane was set at 0.1 ppm in the United States (U.S. Environmental Protection Agency, 1980a). In or on sugarcane bagasse, a tolerance of 0.2 ppm was established (U.S. Environmental Protection Agency, 1980b). Maximum occupational exposure to fluometuron was established at the level of 5 mg/m3 in workplace air in the USSR (International Labor Office, 1980).
The oral LDso of fluometuron was about 8.9 g/kg in rats of both sexes (Ben-Dyke et aI., 1970), 0.9 and 2.4 g/kg in male and female mice, and greater than 10 g/kg in rabbits and dogs (Spencer, 1968). The dermal LDso in rats and rabbits was> 2 and 10 glkg, respectively (Ben-Dyke et al., 1970; Worthing, 1983). LCsos (96 hours) for rainbow trout, crucian carp, and bluegill were 47, 17, and 96 mg/l, respectively (Worthing, 1983). Animals treated with lethal dosages of fluometuron exhibited signs of depression, gasping, hyperpnea, lacrimation, and peripheral vasoconstriction (National Cancer Institute, 1980). Toxicity of fluometuron in 6- to 9-month-old desert sheep has been seen with a single oral dose of 0.8 or 4 g/kg (Mohamed et aI., 1995). Signs of toxicity appeared within 15 minutes after exposure and included depression, salivation, grinding of the teeth, chewing movement of the jaws, mydriasis, dyspnea, incoordination of movements, and drowsiness. Similar signs of toxicity were also seen in animals treated with repeated daily dosages of 25 or 200 mg/kg fluometuron. Laboratory testing revealed increased activities of serum alanine aminotransferase (ALT), aspartate transaminase (AST), and lactate dehydrogenase. Blood urea nitrogen was also elevated. Total serum protein and calcium were significantly decreased. 67.3.3.2 Subchronic Toxicity
In 90-day feeding studies where rats or mice of both sexes were treated with 250, 500, 1000,2000,4000,8000, or 16,000 ppm fluometuron, less weight gain was seen with the three highest doses in both male and female rats (National Cancer Institute, 1980). Deaths occurred in male rats fed 8000 and 16,000 ppm fluometuron and in females receiving 16,000 ppm. Various degrees of spleen enlargement were observed in rats of both sexes treated with 2000 ppm or more of fluometuron. Dose-related pathological changes in rats included mild to severe congestion of the red pulp with corresponding atrophy of the white pulp and depletion of the lymphocytic elements. There were essentially no signs of toxicity observed in mice in these subchronic studies except for a moderate decrease (about 10%) in body weight gain in both sexes at levels of 4000 ppm and greater. 67.3.3.3 Mutagenicity
Fluometuron given as a single dose by gavage in mice exhibited mutagenic activity in both the testicular DNA synthesis inhibition test and the erythrocyte micronucleus test at levels of 1 and 2 g/kg. In the in vitro Ames test, fluometuron also showed mutagenic activity (Seiler, 1978). 67.3.3.4 Carcinogenicity
Mice of both sexes (7 weeks old) fed a diet containing 500 or 1000 mg/kg fluometuron for 103 weeks showed similar incidences of both neoplastic and non-neoplastic lesions compared
67.4 Isoproturon to the control animals. A nonsignificant increase in the incidences of hepatocellular adenomas or carcinomas and tumors in the hematopoietic system (e.g., lymphoma and leukemia) was seen in male mice. Carcinogenicity studies in rats fed 125 or 250 mg/kg ftuometuron-containing diet for 103 weeks were negative (IARC Monographs, 1983; National Cancer Institute, 1980).
67.4 ISOPROTURON Synonyms: (N.N-dimethyl- N' -[4-(l-methylethyl)phenyl]urea; 3-(4-isopropylphenyl)I,I-dimethylurea Chemical abstract number (CAS#): 34123-59-6 Molecular formula: C 12 H I8N20 (206.3) Chemical structure:
Trade names and available formulations: Arelon, Graminon
67.4.1 PHYSICAL AND CHEMICAL PROPERTIES Isoproturon is a colorless powder with a melting point of 155156°C and a vapor pressure of 2.5 x 10-8 mm Hg at 20°e. Its water solubility is about 55 ppm at 20°e. Isoproturon is soluble in most organic solvents and is stable to light, acids, and alkalis (Worthing, 1983). 67.4.2 USAGE Isoproturon was introduced as a herbicide by Hoechst AG under the tradename of "Arelon," by Ciba-Geigy AG under the tradename of "Graminon", and by Rhone-Poulenc as Phytosanitaire. Isoproturon is used to control selectively germinating broadleaf and grass weeds in sugarcane, citrus, cotton, and asparagus. It is widely used abroad. About 3 million pounds of isoproturon are used annually in India. 67.4.3 TOXICITY TO LABORATORY ANIMALS 67.4.3.1 Basic Findings The acute oral LDso of isoproturon in rats was estimated at 1.8-2.4 g/kg, while the acute dermal LDso was >3.2 g/kg (Worthing, 1983). An acute oral LDso of 1 g/kg in rats has also been reported (Behera and Bhunya, 1990).
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In 90-day feeding studies, the no observed adverse effect level was reported to be 400 mg/kg/day in rats and 50 mg/kg/day in dogs (Worthing, 1983). The 96-hour LCsos of isoproturon for carp and rainbow trout were 193 and 240 mg/l, respectively (Worthing, 1983). 67.4.3.2 Subacute and Subchronic Toxicity Repeated dermal exposure of isoproturon technical (IPT; 250, 500, or 1000 mg/kg/day) and its wettable powder formulation (IPF; 750, 1500, or 2250 mg/kg/day) in rats of both sexes for 21 days produced no clear overt signs of toxicity (Dikshith et aI., 1990). Increased organ:body weight ratios (e.g., liver, kidney, adrenal, and spleen) were seen more in the females with both IPT and IPF, especially with the highest dosages. Hematological studies showed that all three dosages of IPT decreased red blood cell counts in males, while only the highest dosage caused a slight reduction in erythrocyte counts in females. Hemoglobin levels in both sexes were reduced by all three dosages of isoproturon (IPT). Neutrophils were decreased and lymphocytes were increased by all IPT exposures. IPF, on the other hand, did not produce any hematological changes. Activities of ALT and AST and protein content in the liver and serum were also altered by IPT or IPF in one or both sexes of the rats. For example, females had an increase in serum AST activity, a decrease in liver ALT, and a significant reduction in serum protein after repeated IPT. Repeated IPF exposure reduced serum ALT activity in male rats and lowered total serum proteins in female rats. No overt signs of toxicity were observed in rats treated orally with 200, 400, or 800 mg/kg/day of isoproturon for 42 and 60 days (Sarkar et aI., 1995). The highest dosage of the chemical, however, did significantly decrease body weight. Moreover, a dose-dependent increase in the liver weight was noted. Isoproturon also increased the weights of the kidney and the heart. Histopathological changes included hepatocellular degeneration and focal necrosis in the liver, glomerular and tubular degeneration in the kidney, and hemosiderosis in the spleen. 67.4.3.3 Genotoxicity Isoproturon and its structural analogs ftuometuron and monuron have all been reported to be genotoxic (Behera and Bhunya, 1990; Garrett et aI., 1986; Seiler, 1978). Behera and Bhunya (1990) found that 100, 150, and 200 mg/kg of isoproturon given intraperitoneally to adult Swiss albino mice induced various types of chromosomal aberrations in bone marrow cells, for example, chromatid gaps, chromatid breaks, acentric fragments, chromatid exchanges, ring chromosomes, and metacentric chromosomes. The highest dosage of isoproturon also induced the formation of micronuclei in bone marrow cells. Pregnant rats treated orally with 180 mg/kg of isoproturon daily from gestation day 6 to gestation day 20 also exhibited chromatid breaks in bone marrow cells (Srivastava and Raizada, 1995).
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CHAPTER 67
Phenylurea Herbicides
67.4.3.4 Reproductive Toxicity
67.4.3.7 Biochemical Effects
Isoproturon has been found to induce abnormalities in sperm shape in a dose-dependent manner in mice (Behera and Bhunya, 1990). The sperm would either have hammer, mushroom, or amorphous shaped heads, or hook and beak shaped acrosomal ends. Sarkar and co-workers (1997) also reported the potential toxic effects of isoproturon on the male reproductive system in rats. When isoproturon was given orally to rats 6 days/week at the rate of 200, 400, or 800 mg/kg/day for 10 weeks, the highest dosage decreased epididymal sperm counts and the percentage of motile sperm and increased the percentage of abnormal sperm (e.g., distorted heads, atypical tails, bent necks or midpieces). Degeneration and desquamation of germinal layer cells were observed in the testis. Tubular lumens of the testis and the epididymis exhibited reduced numbers of spermatids and spermatozoa, respectively (Sarkar et aI., 1995). The activities of androgen biosynthesis-related enzymes, for example, glucose-6-phosphate dehydrogenase and ~5-3,B-hydroxy steroid dehydrogenase, were reduced in a dose-related manner (Sarkar et aI., 1997). Overall, isoproturon has been suggested to have the potential to cause maturational malformation of sperm cells and retarded spermatogenesis in rats.
Isoproturon given by gavage to rats at the dose of 116 LD50 for 3 consecutive days induced significantly the activities of hepatic enzymes, for example, NADPH-cytochrome c reductase, microsomal epoxide hydrolase, 7-ethoxycoumarin O-deethylase, aldrin epoxidase, UDP-glucuronyl-transferase, and glutathionS-transferase (Schoket and Vincze, 1985, 1986). Antipyrine plasma half-life, however, was not affected by the isoproturon treatment, suggesting that hepatic cytochrome P-450 isozymes, for example, benzo(a)pyrene monooxygenase and 7ethoxyresorufin O-deethylase, were not induced by isoproturon (Schoket et aI., 1987).
67.4.3.5 Fetotoxicity and Teratogenicity Isoproturon (45, 90, or 180 mg/kg/day) given orally to pregnant rats from gestation days 6-20 caused no observable fetotoxic and/or teratogenic effects. The numbers of implantations and resorptions, fetal body weights, and external, visceral, and skeletal structures were all comparable to those in the controls (Srivastava and Raizada, 1995). With higher dosages of isoproturon (225, 450, and 900 mg/kg/day) given orally from gestation days 6-15 (Sarkar and Gupta, 1993a), however, doserelated depression and drowsiness of the dams were observed. Though there was no lethality associated with isoproturon exposures, decreased maternal body weight was noted during later pregnancy (gestation days 15-20) with the dosages of 450 and 900 mg/kg. Litter size, fetal weights, and crown-rump and transumbilicallengths were all decreased by 450 mg/kg or more of isoproturon. Moreover, there was a significant increase in the frequency of fetal resorptions and the number of fetuses with retarded growth. Again, no major visceral and/or skeletal malformations were observed. 67.4.3.6 Neurological Effects Neurotoxic effects of isoproturon in mice have been reported by Sarkar and Gupta (1993b). They found a single oral dose of isoproturon (0.5, 1.0, or 2.0 g/kg) potentiated both pentobarbitaland barbital-induced sleeping time. Spontaneous and forced locomotor activity were reduced by 1.0 and 2.0 mg/kg isoproturon. In addition, isoproturon exhibited anticonvulsant activity against electroshock and pentylenetetrazol-induced convulsions. The mechanisms of these inhibitory effects of isoproturon on the central nervous system are unclear.
REFERENCES Agrawa1, R. c., Kumar, S., and Mehrotra, N. K. (1996). Micronucleus induction by diuron in mouse bone marrow. Toxicol. Lett. 89, 1-4. Antony, M., Shukla, Y., and Mehrotra, N. K. (1989). Tumor initiatory activity of a herbicide diuron on mouse skin. Cancer Lett. 48, 125-128. Amaud, L., Taillandier, G., Kaouadji, M., Ravanel, P., and Tissut, M. (1994). Photosynthesis inhibition by phenylureas: A QSAR approach. Ecotoxicol. Environ. Safety 28, 121-133. Aspelin, A. L. (1997). "Pesticides Industry Sales and Usage: 1994 and 1995 Market Estimates," EPA 733-R-97-002, Office of Prevention, Pesticides and Toxic Substances (7503W), United States Environmental Protection Agency, Washington, DC. Behera, B. C., and Bhunya, S. P. (1990). Genotoxic effect of isoproturon (herbicide) as revealed by three mammalian in vivo mutagenic bioassays. Indian 1. Exper. Bioi. 28, 862-867. Ben-Dyke, R., Sanderson, D. M., and Noakes, D. N. (1970). Acute toxicity data for pesticides. World Rev. Pest. Contr. 9(3), 119-127. Boehme, C., and Emst, W. (1965). The metabolism of urea-herbicides in the rat. 2. Diuron and linuron. Food Cosmet. Toxico!. 3, 797-802. [In German]. Boyd, E. M., and Krupa, V. (1970). Protein-deficient diet and diuron toxicity. 1. Agr. Food Chem. 18(6), 1104-1107. Canna-Michaelidou, S., and Nicolaou, A.-S. (1996). Evaluation of the genotoxicity potential (by Mutatox™) test of ten pesticides found as water pollutants in Cyprus. Sci. Total Environ. 193,27-35. Cemakova, M. (1995). Biological degradation of isoproturon, chlortoluron and fenitrothion. Folia Microbio!. 40(2), 201-206. Chapman, R. A. (1967). "Tolerances for Residues of Pesticide Chemicals," T.LL. No. 290. Food and Drug Directorate, Department of National Health and Welfare, Ottawa. Crosby, D. G., and Tucker, R. K. (1966). Toxicity of aquatic herbicides to Daphnia magna. Science 154, 289-290. Dalton, R. L., Evans, A. W., and Rhodes, R. C. (1966). Disappearance of diuron in cotton field soils. Weeds 14, 31-33. Dikshith, T. S. S., Raizada, R. B., and Srivastava, M. K. (1990). Dermal toxicity to rats of isoproturon technical and formulation. Vet. Hum. Toxico!. 32(5), 432-434. Engelhard, G., Wallnofer, P. R., and Plapp, K. (1972). Identification of N,Odimethylhydroxylamine as microbial degradation product of the herbicide linuron. Appl. Microbiol. 23, 664-666. Garrett, l\. E., Stack, H. E, and Waters, M. D. (1986). Evaluation of the genetic activity profiles of 65 pesticides. Mut. Res. 168, 301-325. Geissbuhler, H. c., Haselback, c., Aebi, H., and Ebner, L. (1963). The fate of N'-(4-chlorophenoxy)-phenyl-N,N-dimethylurea (C-1983) in soils and plants. Ill. Breakdown in soils. Weed Res. 3, 277-297. Hodge, H. C., Downs, W. L., Panner, B. S., Smith, D. w., and Maynard, E. A. (1967). Oral toxicity and metabolism of diuron (N-(3,4-dichlorophenyl)N',N'-dimethylurea) in rats and dogs. Food Cosmet. Toxicol. 5, 513-531. IARC Monographs (1983). Fluometuron. IRAC Monographs on the Evaluation of the Carcinogenic Risk of Chemicals to Man 30, 245-253.
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International Labor Office (1980). "Occupational Exposure Limits for Airborn Toxic Substances," 2nd (revised) ed., Occupational Safety and Health Series, No. 37, pp. 106; 118-119. Geneva. Khera, K. S., Whalen, c., Trivett, G., and Angers, G. (1979). Teratogenicity studies on pesticidal formulations of dimethoate, diuron and lindane in rats. Bull. Environ. Cantam. Taxiea!. 22, 522-529. Kinoshita, F. K., and DuBois, K. P. (1970). Induction of hepatic microsomal enzymes by Herban, diuron and other substituted urea hericides. Taxieol. App!. Pharmaeo!' 17,406-417. Kotzias, D., and Korte, F. (1981). Photochemistry of phenylurea herbicides and their reactions in the environment. Eeotoxieol. Environ. Safety 5, 503-5 I 2. Mehmood, Z., KeIIy, D. E., and KeIIy, S. L. (1995). Metabolism of the herbicide chlortoluron by human cytochrome P450 3A4. Chemosphere 31(11112), 4515-4529. Mohamed, O. S. A., Ahmed, K. E., Adam, S. E. I., and Idris, O. F. (1995). Toxicity of cotoran (f1uometuron) in desert sheep. Vet. Hum. Toxieol. 37(3), 214-216. National Cancer Institute (1980). "Bioassay of Fluometuron for Possible Carcinogenicity." Carcinogenesis Technical Report Series. CAS No. 216417-2, NCI-CG-TR-195, NTP-80-11. Ramamoorthy, S., and Baddaloo, E. G. (1995). Aquatic toxicity data. In "Handbook of Chemical Toxicity Profiles of Biological Species. Volume I: Aquatic Species," pp. 165,172,251,284. CRC Press, Boca Raton, PL. Reinhold, V. N. (1987). Diuron, a review. Dangerous Properties af Industria! Materia! Report 7(5),49-55. Rhee, K. H., Morris, E. P., Barber, J., and Kuhlbrandt, W. (1998). Threedimensional structure of the plant photosystem 11 reaction centre at 8 A resolution. Nature 398(19),283-286. Sarkar, S. N., Chattopadhyay, S. K., and Majumdar, A. C. (1995). Subacute toxicity of urea herbicide, isoproturon, in rats. Indian 1. Exper. Bio!. 33, 851-856. Sarkar, S. N., and Gupta, P. K. (l993a). Fetotoxic and teratogenic potential of substituted phenylurea herbicide, isoproturon, in rats. Indian J. Exper. Bio!. 31, 280-282. Sarkar, S. N., and Gupta, P. K. (I 993b). Neurotoxicity of isoproturon, a substituted phenylurea herbicide, in mice. Indian J. Exper. Bio!. 31, 977-981. Sarkar, S. N., Majumdar, A. c., and Chattopadhyay, S. K. (1997). Effect of isoproturon on male reproductive system: Clinical, histological and histoenzymological studies in rats. Indian 1. Exper. Bia!. 35, 133-138.
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Schoket, B., and Vincze, I. (1985). Induction of rat hepatic drug metabolizing enzymes by substituted urea herbicides. Acta Pharmaeol. Toxieo!. 56, 283288. Schoket, B., and Vincze, I. (1986). Induction of rat hepatic microsomal epoxide hydrolase by substituted urea herbicides. Acta Pharmacal. Taxieo!. 58, 156158. Schoket, B., and Vincze, I. (1990). Dose-related induction of rat hepatic drug-metabolizing enzymes by diuron and chlorotoluron, two substituted phenylurea herbicides. Taxiea!. Letl. 50, 1-7. Schoket, B., Zilahy, Z., Molnar, J., and Vincze, I. (1987). Comparative investigation of antipyrine half-life and induction of cytochrome P-450 dependent monooxygenases in rats treated with phenylurea herbicides. In Vivo 1, 185188. Seiler, J. P. (1978). Herbicidal phenylalkylureas as possible mutagens I. Mutagenicity tests with some urea herbicides. Mut. Res. 58,353-359. Spencer, E. Y. (1968). "Guide to the Chemicals Used in Crop Protection," 5th ed. Research Branch Agriculture Canada Publication 1093, University of Western Ontario, London, Ontario. Srivastava, M. K., and Raizada, R. B. (1995). Developmental toxicity of the substituted phenylurea herbicide isoproturon in rats. Vet. Hum. Toxieo!. 37(3), 220-223. Stevens, J. T., and Sumner, D. D. (1991). Herbicides. In "Handbook of Pesticide Toxicology. Vo!. 3. Classes of Pesticides" (W. J. Hayes, Jr. and E. R. Laws, Jr., eds.), pp. 1349-1350. Academic Press, San Diego, CA. U.S. Environmental Protection Agency (l980a). "Protection of Environment." US Code of Federal Regulations, Title 40, part 180.229. U.S. Environmental Protection Agency (I 980b). "Food and Drugs." US Code of Federal Regulations, Title 21, part 561.240. Vroumsia, T., Steiman, R., Seigle-Murandi, F., Benoit-Guyod, J. L., and Khadrani, A. (1996). Biodegradation of three substituted phenylurea herbicides (chlortoluron, diuron, and isoproturon) by soil fungi. A comparative study. Chemosphere 33(10),2045-2056. Wang, S. W., Chu, C. Y., Hsu, J. D., and Wang, C. J. (1993). Haemotoxic effect of phenylurea herbicides in rats: Role of haemoglobin-adduct formation in splenic toxicity. Food Chem. Toxic. 31(4), 285-295. Worthing, C. R. (1983). "The Pesticide Manual: A World Compendium," 7th ed., pp. 226,281,329. BCPC.
CHAPTER
68 Protoporphyrinogen Oxidase Inhibitors Franck E. Dayan, Joanne G. Romagni, and Stephen O. Duke United States Department of Agriculture
68.1 INTRODUCTION Protoporphyrinogen oxidase (Protox)-inhibiting herbicides have been used since the 1960s and now represent a relatively large (ca. 10%) and growing segment of the herbicide market. Before their molecular target, Protox, was discovered (Matringe et aI., 1989a, b), this category of herbicides was often termed "diphenyl ether-type herbicides." At that time, almost all of the Protox inhibitors were diphenyl ethers. This nomenclature led to some confusion in herbicide classification, because other "diphenyl ether" herbicides have an entirely different molecular site of action (i.e., inhibition of acetyl CoA carboxylase). Now, many other structural classes of Protox inhibitors are commercialized. In general, the newer products are more potent Protox inhibitors, resulting in lower application rates than the older herbicides of this class. Some of them appear to be analogs of the substrate or a substrate/product transition state of the enzyme (Reddy et aI., 1998). There are several previous reviews of the Protox inhibitors, both before (B6ger, 1984; Gilham and Dodge, 1987; Kunert et aI., 1987; Matsunaka, 1976) and after (Dayan and Duke, 1996, 1997; Duke et aI., 1991; Nandihalli and Duke, 1993; Reddy et aI., 1998; Scalla and Matringe, 1994) their target site was known. The emphasis in most of these previous reviews was on the mode of action. An entire book dealing mainly with Protox inhibitors is available (Duke and Rebeiz, 1994). This review updates and supplements these previous publications.
68.2 COMMERCIALLY AVAILABLE PROTOX INHIBITORS 68.2.1 DIPHENYL ETHER PROTOPORPHYRINOGEN OXIDASE INHIBITORS Nitrofen was the first Protox-inhibiting herbicide to be introduced for commercial use in 1964. This diphenyl ether (DPE) Handbook of Pesticide Toxicology Volume 2. Agents
herbicide was eventually recognized as a relatively weak inhibitor of Protox (Nandihalli et aI., 1992), but was a lead compound of an entire class of structurally related herbicides that were much more active. Several DPE herbicides (see Table 68.1) have been successfully commercialized (Anderson et aI., 1994). Although many of these older commercialized DPEs have a p-nitrophenyl substitution, newer DPE-like herbicides more commonly contain p-trifiuoromethyl phenyl substitutions. These new herbicides are heterocyclic phenyl ethers (structurally related to DPEs), where one of the phenyl rings of ether is replaced by an aromatic heterocycle. Recently reported examples include 6-aryloxy-1 H -benzotriazoles (Condon et aI., 1995), aryloxyindolin-2(3H)-ones (Karp et aI., 1995), 5-aryloxybenzisoxazoles (Wepplo et aI., 1995), 6-aryloxyquinoxalin-2,3-diones (Anderson et aI., 1994), benzheterocycles (Lee et aI., 1995), and benzoxazines (Sumida et aI., 1995). Pyrazolyl, pyridyl, and furyl rings have also been investigated as the heterocyclic component (Anderson et aI., 1994; Armbruster et aI., 1993; Clark, 1994; 1996; Sherman et aI., 1991). 68.2.2 NON-DIPHENYL ETHER PROTOPORPHYRINOGEN OXIDASE INHIBITORS Whereas the first generation of Protox inhibitors (with the exception of oxadiazon) was primarily based on the DPE backbone, numerous non-oxygen-bridged compounds with this same site of action have been discovered (Table 68.2). These compounds invariably consist of heterobicyclic structures with one phenyl ring attached to a heterocyclic ring. The linkage can consist of a carbon-carbon bridge, as with the isoxazole carboxamides (Dayan et aI., 1997a; Hamper et aI., 1995), but more often consists of a carbon-nitrogen bridge, as with the phenyl imides (Lyga et aI., 1991; Mito et aI., 1991; Sato et aI., 1987), triazolinones (Amuti et aI., 1997; Dayan et aI., 1997b, c; Theodoridis, 1997; Theodoridis et aI., 1992, 1995), oxadia-
1529
1530
CHAPTER 68
Protoporphyrinogen Oxidase Inhibitors
Table 68.1 Commercially Available DPE and DPE-like Protox-Inhibiting Herbicides a Common name
V.S. trade name
Primary source
Chemical structure Cl
"C-O-O-O-NO'
Acifluorfen
BASF
Bifenox
Rhone-Poulenc
Complete®
Fluoroglycofen
COOH
Rohm and Haas
o Cl
11
0
11
C-NH-S-CH3
,~--6-a-Q-.o, ~
Fomesafen
Zeneca
Lactofen
Valent
Oxyfluorfen
Rohm and Haas
aInformation from Ahrens (1994).
zolones (Dickmann et aI., 1997), and pyrazoles (Prosch et aI., 1997). There are several reasons for the high level of interest among agrochemical companies in the discovery and development of new Protox inhibitors. For one thing, weeds have not shown a propensity to develop resistance to these herbicides. It is likely that Protox inhibitors may soon supplant acetolactate synthase inhibitors as the preferred herbicides for broad-spectrum weed control in soybean fields. Additionally, the market niche for Protox inhibitor is beginning to expand to weed control in monocot crops, with the marketing of newer compounds such as carfentrazone, JV 485, and oxadiargyl (Table 68.3). Finally, Protox seems to be a particularly good target that can be inhibited by structurally diverse classes of herbicides, allowing for the development of a new chemistry not yet patented by other companies.
68.3 AGRICULTURAL USE 68.3.1 CROPS AND WEEDS Protox inhibitors have historically been used for broad-spectrum weed control in soybean fields (Table 68.3). Dayan and Duke (1996) proposed that the market share of Protox in-
hibitors could be increased by developing compounds for use in monocot crops. Several non-DPE herbicides based on the triazolinone and the oxadiazole structures have been, or will soon be, registered for weed control in cereal crops. At the moment, most of the compounds, with selectivity for cereals and small grains (bifenox, carfentrazone, and fluoroglycofen), are currently not available in the D.S. market. Finally, compounds with the highest biological activity, such as oxyfiuorfen and azafenidin, have also been developed for use as nonselective herbicides in noncrop areas and nurseries. As a group, Protox-inhibiting herbicides control both monocotyledonous and dicotyledonous weeds. The market for weed control in soy bean fields is intense because a similar spectrum of weed control can be obtained with herbicides with other sites of action. In particular, acetolactate synthase (ALS) inhibitors, such as the imidazolinones and sulfonyl ureas, have become important tools for soy bean growers. The situation may be changing, though, as many weeds have become resistant to the ALS-inhibiting herbicides by evolving herbicide-insensitive forms of the enzyme. In contrast, no cases of evolved resistance to Protox inhibitors have been reported. Possible reasons for this will be discussed in Section 68.4.3. As a result, agrochemical companies are actively pursuing the development of new Protox inhibitors in order to replace ALS-inhibiting herbicides as resistant weeds become more widespread in soybean fields.
68.4 Behavior in Plants
1531
Table 68.2 Non-Oxygen-Bridged Protox-Inhibiting Herbicides Common name
Trade name
Azafenidina
Carfentrazoneb
Primary source
Chemical structure
Dupont
Affinity®
FMC
rY<°A~COOH
FIumiclorac c
Resource®
Valent
v\(N-yCI o
JV-485 d
Monsanto
Oxadiargyle
Rh6ne-Poulenc
Oxadiazonc
Ronstar®
Rh6ne-Poulenc
Sulfentrazone f
Authority®
FMC Corporation
F
et al. (1997). (1995b). c Ahrens (1994). dprosch et al. (1997). eDickmann et al. (1997). f Anonymous (1995a). a Amuti
b Anonymous
68.3.2 MODE OF APPLICATION For the most part, Protox-inhibiting herbicides are applied postemergence during the early stages of weed development. The rate of application ranges widely among Protox inhibitors. Most of the DPE herbicides (such as acifluorfen and lactofen) are applied at rates of 100 to 500 g active ingredient per hectare. Fluoroglycofen is more active and can be applied at rates 10fold lower. The most active of these herbicides can be applied at rates as low as 1 g active ingredient per hectare. With a few exceptions, these compounds do not have preemergence activity and have little residual activity in soils.
However, more recently discovered structures such as sulfentrazone have excellent preemergence activity (Table 68.3) (Theodoridis et aI., 1992). The preemergence-applied Protox inhibitors are active at concentrations ranging from 2 to 4 kg active ingredient per hectare for oxadiazon to 100 g active ingredient per hectare for sulfentrazone.
68.4 BEHAVIOR IN PLANTS Protox inhibitors cause rapid photobleaching and light-dependent desiccation of foliage. The symptoms observed on the
1532
CHAPTER 68
Protoporphyrinogen Oxidase Inhibitors
Table 68.3 Targeted Crops and Mode of Application of Commercially Available ProtoxInhibiting Herbicides Main crop
Applicationa
Acifluorfen a
Soybean, peanut, rice
Postemergence
Azafenidin b
Perennial crops/forestry
Preemergence
Bifenoxo
Small grain
Preemergence/postemergence
CarfentrazoneC
Cereal crops
Postemergence
Flumiclorac"
Soybean and maize
Preemergence/postemergence
Fluoroglycofena
Cereal crops
Postemergence
Fomesafen"
Soybean
Postemergence
JV-485 d
Winter wheat
Preemergence
Lactofena
Soybean
Postemergence
Oxadiargyle
Rice/sugarcane
Preemergence
Oxadiazon a
Grasses and omamentals
Postemergence
Oxyfluorfen"
Vegetable crops
Preemergence/postemergence
Sulfentrazone f
Soybean, sugarcane,
Preemergence
Common name
tobacco a Ahrens
(1994). et al. ( 1997). C Anonymous (I 995b ). dprosch et al. (1997). eDickmann et al. (1997). f Anonymous (l995a). b Amuti
foliage of DPE-treated plants include leaf cupping, crinkling, bronzing, and necrosis (John son et aI., 1978). The lesions on the foliage are due to loss of membrane integrity that leads to cellular leakage (Fig. 68.1) (Dayan et aI., 1997c, 1998; Kenyon et aI., 1985; Lee et aI., 1995). Other physiological responses include inhibition of photosynthesis; evolution of ethylene, ethane, and malondialdehyde; and, finally, bleaching of chloroplast pigments (Kenyon et aI., 1985). Protox inhibitors are known to cause temporary injury to the foliage of treated 700 600
~
500
68.4.1 ABSORPTION, TRANSLOCATION, AND METABOLISM
Most postemergence-applied Protox inhibitors are readily absorbed through the leaves (Dayan et aI., 1996, 1997b; Ritter and Coble, 1981a). Root uptake of foliar active compounds is generally poor (Ritter and Coble, 1981b). Once absorbed into the foliage, little translocation is observed. Most of the DPE herbicides (such as acifluorfen, lactofen, and fluoroglycofen) are not translocated beyond the point of absorption. However, some of them, such as fomesafen, can be readily translocated by the xylem. Some studies demonstrate that absorption and translocation of DPE herbicides may be affected by temperature and humidity (Ritter and Coble, 1981a; Wills and McWhorter, 1981). Soil active compounds, such as sulfentrazone, are readily taken up by the roots and translocated through the xylem along the transpiration flow (Wehtje et aI., 1997). Radiolabeled pulsechase studies show that nearly all of the herbicide taken up by the roots is translocated to the foliage within 24 h (Fig. 68.2) (Dayan et aI., 1996, 1997c). Interestingly, the related sensitive (coffee senna) and resistant (sicklepod) weeds absorbed and translocated sulfentrazone in the same manner (Fig. 68.2). The selectivity of Protox inhibitors lies primarily in the differential rate of metabolism. In the case of DPE herbicides, such as acifluorfen and fluorodifen, the resistance of soy bean is achieved by metabolic cleavage of the ether bridge, followed
. . r - - - - -(c)
(a)
'E
crops (particularly soybean). However, crops generally recover rapidly, and yields are not affected (Vidrine et aI., 1993, 1994, 1996; Walker et aI., 1992). Though not desirable, crop injury on soy bean is not unusual and farmers have become accustomed to this phenomenon as the use of ALS-inhibiting herbicides has become more widespread.
~
E
400
13
200
,2; .?:- 300 .s; ::J
-g o
100
t)
o
8 16 24 32 40 48 0
8 16 24 32 40 48 0
8 16 24 32 40 48
Time (h ) Figure 68.1 Herbicidal activity of carfentrazone-ethyl and deethylated carfentrazone metabolite as measured by electrolyte leakage from leaf discs of (a) soybean (Glycine max), (b) velvetleaf (Abutilon theophrasti), and (c) ivyleaf mominglory (Ipomoea hederacea). Leaf discs were incubated in the presence of no inhibitor (el, 10 ;tM carfentrazone-ethyl (_), or the free-acid metabolite (A) for 20 h in darkness and then exposed to continuous light. The arrows indicate the beginning of light exposure and the dotted lines indicate maximum conductivity from boiled samples. Data are the average of three replications ± SD (from Dayan et aI., 1997b).
Figure 68.2 Uptake and translocation of 14C-sulfentrazone in (a) coffee senna (Senna occidentalis) and (b) sicklepod (Cassia obtusifolia). Autoradiograms were obtained by incubating the roots of both plants in a solution containing 14C-sulfentrazone for 3 hand transfening the plants to nonlabeled solutions for another 21 h. R, root; C, cotyledons; 1,2, and 3 refer to the first, second, and third true leaves. Notice that no radioactivity remained in the roots and that most of the radioactivity accumulated in the fully expanded leaves (I and 2).
68.4 Behavior in Plants
by rapid conjugation of the phenyl rings to cysteine, homoglutathione, and glucose (Fig. 68.3) (Frear and Swanson, 1973; Frear et aI., 1983). In the case of non-DPE structures, such as sulfentrazone, resistance is achieved via rapid oxidative degradation followed by conjugation (Dayan et aI., 1996, 1997c). The nature of the conjugated metabolites was not determined in this study, but 90% of the herbicide was transformed into extremely water-soluble metabolites within 24 h from the time of application (Dayan et aI., 1996, 1997c). Sensitive weeds are apparently unable to oxidize sulfentrazone (Dayan et aI., 1996). Further studies on the biological activity of the primary metabolites of sulfentrazone showed that the initial oxidative step leading the hydroxylation of the methyl group on the triazolinone ring did not reduce the biological activity of sulfentrazone. Further oxidation of the moiety led to the formation of less active compounds (Dayan et al., 1998).
1533
68.4.2 MODE OF ACTION Early investigations on the mode of action of Protox inhibitors determined that light was needed for herbicidal activity (Matsunaka, 1969); yet, photosynthesis was not involved in the mechanism (Duke and Kenyon, 1986). There was good evidence that the observed peroxidative damage was due to a process involving a photodynamic pigment [reviewed by Scalla and Matringe (1994)]. Matringe and Scalla (1988) demonstrated that the chlorophyll precursor protoporphyrin IX (Proto) accumulated in acifluorfen-treated plant tissues. Proto is a photodynamic pigment (Cox et al., 1982), and its content in Protox inhibitor-treated plant tissues correlated well with peroxidative damage (e.g., Becerril and Duke, 1989). Matringe et al. (1989a, b) discovered that Proto accumulated in response to the inhibition of the enzyme responsible for its synthesis (Fig. 68.4). This apparently contradictory situation is mirrored by the accumulation of Proto in humans with the genetic disease variegate porphyria that results from a deficiency of Protox activity (Deybach et aI., 1981). The paradox of inhibition of an enzyme leading to the accumulation of its catalytic product is explained by altered compartmentalization of porphyrin intermediates (Jacobs et aI., 1991; Lee et aI., 1993). Inhibition of Protox induces an uncontrolled accumulation of protoporphyrinogen IX (Protogen), which leaks out of the chloroplast outer membrane into the cytoplasm where it is converted into the highly photodynamic Proto (Fig. 68.5). In the presence of light, this
Glutamic acid
Chlorophyll
t
o-aminolevulinic acid
Chlorophyllide
t
(b)
CHF2
O
~
(0)
--...
CH 0H 1(;N"N>- 2
I
Cl
h
Coproporphyrinogen
Mg-Proto
COOH
.(0)
" k I
o
Conjugates - . - -
Protogen ,CHF2
t:>-COOH
Cl
h
COOH
I
Figure 68.3 Examples of hydrolytic and oxidative metabolic degradation of various Protox inhibitors in plants. (a) Metabolism of the diphenyl ether aciftuorfen in soybean plants (Frear et aI., 1983) and (b) metabolism of the triazolinone carfentrazone (Anonymous, 1995b).
-1 Protox ~ I Dipbenyl ethers Oxadiazoles N-pbenylimides Triazolinones
t Proto
~ Heme
Figure 68.4 Simplified pathway of chlorophyll and heme biosynthesis in plants showing the enzymes known to be sensitive to chemical inhibition (boxes). ALA dehydratase, aminolevulinic acid dehydratase; PBG deaminase/Uro III cosynthase, porphobilinogen deaminase/uroporphyrinogen III cosynthase; Protox, protoporphyrinogen oxidase (redrawn from Reddy et aI., 1998).
1534
CHAPTER 68
Protoporphyrinogen Oxidase Inhibitors
Light Hatblelde DroplM
Oxldatlve Degradation
Figure 68.5 Mode of action of Protox-inhibiting herbicides. The initial step consists of the foliar absorption and translocation of the herbicide. Inhibition of Protox (localized in the outer membrane of chloroplasts) causes an unregulated accumulation of Protogen, which is oxidized (either enzymaticaIIy or chemically) to Proto. Proto is energized by light and causes formation of reactive singlet oxygen species that can lead to oxidative membrane degradation (redrawn from Dayan and Duke, 1997).
photosensitized Proto generates a highly reactive singlet oxygen that induces lipid peroxidation of the relatively unprotected plasma membrane (Fig. 68.5) (Devine et al., 1993; Lee et aI., 1993). This phenomenon explains the light-dependent nature of the mode of action of Protox inhibitors. A large number of Protox inhibitors exists, and in spite of their structural differences, they apparently all compete for the same site on the enzyme, suggesting that the binding pocket is very promiscuous. Figure 68.6 demonstrates the competitive binding between the natural substrate and various synthetic herbicides belonging to structurally different chemical classes. The regression of the lines intersecting with each other at the level of the y axis on a double reciprocal plot is typical of competitive binding. These graphs also enable the calculation of binding constants for each inhibitor. As would be expected, the inhibition of Protox is proportional to the ability of each compound to bind to that particular site on Protox. Early studies on DPE
C
.8 :2
68.4.3 MODES OF RESISTANCE
Plants can be resistant to herbicides via physical, physiological, and/or biochemical mechanisms. Most often, resistance is achieved via slow uptake and translocation, rapid metabolic degradation, and/or resistance at the molecular site. The complex mode of action of Protox inhibitors provides several more unusual sites for possible herbicide resistance (Fig. 68.7) (Duke et aI., 1997). Reduced uptake and translocation of Protox-inhibiting herbicides through the shoots may account for the resistance of some species (Matsumoto et aI., 1994) but does not play a major role in resistance to those that are soil applied (Dayan et aI., 1996, 1997b, c). On the other hand, metabolic degradation of Protox inhibitors seems to plays a key role in crop resistance to these herbicides. Resistance of soybean to acifluorfen and two phenyl triazolinones is due primarily to rapid metabolic degradation of the herbicides (Dayan et aI., 1996, 1997b, c; Frear et aI., 1983). However, these herbicides act so rapidly that metabolic degradation does not provide a large safety margin, and some crop damage, often referred to as "bronzing," is com-
(c)
(b)
"0
herbicides indicated that these molecules compete for the binding site by mimicking half of the natural substrate Protogen (Nandihalli et aI., 1992). However, no such clear resemblance to Protogen can be easily established with non-oxygen-bridged Protox inhibitors (Dayan et aI., 1997a-c; Reddy et aI., 1998). Two-dimensional quantitative structure-activity relationship (QSAR) analyses have been somewhat successful in predicting the herbicidal activities of these compounds (Nandihalli et aI., 1992; Reddy et aI., 1995, 1998). However, equations derived from three-dimensional (3-D) QSAR have been more reliable at predicting the activity of various structurally related groups of Protox-inhibiting herbicides, as well as differentiating active from inactive stereoisomers (Dayan et aI., 1999).
2
c
c:
'Cjj
e' Q. Cl
E
Oxidative Degradation
0.1 0.2 0.3 04 0.5
1/nM Free
1/nM Free
1/nM Free
Figure 68.6 Competitive binding between the radiolabeled acifiuorfen (DPE herbicide), Protogen (the natural substrate of Protox), and various Protox inhibitors on isolated etioplasts. The graphs illustrate the competitive nature of the binding between acifiuorfen (0) and (a) Protogen (e), (b) a triazolinone-type Protox inhibitor (_), and (c) an isoxazole carboxamide-type Protox inhibitor (A) (redrawn, respectively, from Matringe et aI., 1992; Dayan et aI., 1997b; Dayan et aI., I 997a).
Figure 68.7 Schematic of possible mechanisms of resistance to Protoxinhibiting herbicides. Potential sites of resistance are in boldface numbers: 1, inhibition of uptake or sequestration of the herbicide; 2, rapid metabolic degradation of the herbicide; 3, herbicide-resistant Protox; 4, degradation of extraplastidic Protogen and/or Proto; 5, inactivated herbicide-resistant, extraplastidic Protox; 6, quenching of singlet oxygen and other toxic oxygen species (redrawn from Dayan and Duke, 1996).
68.5 Environmental Impact
mono The extent of injury might be greater with those herbicides that have longer soil persistence. There are no cases of natural resistance in whole plants associated with herbicide-insensitive chloroplastic Protox. Nevertheless, herbicide-resistant Protox has been isolated from tobacco (Ichinose et aI., 1995) and soybean cell cultures selected with Protox inhibitors (Pornprom et aI., 1994). Overexpression of mitochondrial Protox in photomixotrophic tobacco cells lines can result in resistance to Protox inhibitors (Watanabe et aI., 1998). This result suggests that if sufficient uninhibited mitochondrial Protox is available, Protogen leaking from inhibited plastids can be rapidly converted to heme by the mitochondria before it can accumulate as damaging extraorganellar Proto. Whole plants have not been regenerated from any of these cell lines. Rice appears to be more naturally resistant to the oxidative stress induced by Protox inhibitors relative to the targeted weeds (Matsumoto et aI., 1994). That is, the plant generates Proto in response to the herbicides, but appears to have the ability to cope with the resulting singlet oxygen and the toxic compounds (hydroxyl radical, lipid peroxides, etc.) resulting from it. We have reported that a similar mechanism may be involved in the differential sensitivity of soybean cultivars to sulfentrazone (Dayan et aI., 1997c). Other species (e.g., mustards) and older tissues of some species that are sensitive in the seedling stage are apparently resistant due to enzymatic degradation of Protogen to nontoxic compounds (Jacobs et a!., 1996). The complex mechanism of action of Protox inhibitors provides several potential mechanisms for evolved resistance in weeds (Fig. 68.7). Yet, no cases of evolved resistance in the field have been verified. This could be due, in part, to the relatively short-lived selection pressure of these fast-acting, foliar-applied herbicides. However, the recent development of more persistent soil-active Protox inhibitors will increase the selection pressure, increasing the probability of the evolution of resistance. No resistant mutants were obtained when hundreds of thousands of mutagenized Arabidopsis thialiana seeds were tested with a Protox inhibitor (M. Yamamoto and S. O. Duke, unpublished). Recent patents (e.g., Ward and Volrath, 1995) propose to produce Protox inhibitor-resistant crops through a variety of molecular genetics approaches. To date, the only published successful genetically engineered crop resistant to Protox inhibitors was obtained by transforming tobacco with a herbicideresistant form of Protox from bacteria (Choi et aI., 1998).
1535
and sulfentrazone). Nevertheless, the excellent broad-spectrum preemergence activity associated with greater soil persistence of sulfentrazone, relative to other Protox-inhibiting herbicides, makes this herbicide unique in its class for the moment. The combination of relatively high soil adsorption and rapid microbial degradation strongly limits soil leaching of most Protox-inhibiting herbicides. None of the Protox inhibitors has volatilization problems (vapor pressure lower than 10-7 mm Hg at 25°C), and none has caused drift-related injury to nontarget crops when properly applied. It is important to note that soil quality may affect leaching of fluoroglycofen, and metabolites of lactofen may be highly mobile in soil (Ahrens, 1994). In the case of soil active Protox inhibitors, mobility may be of some concern. The herbicide bifenox is quite mobile in spite of its relatively high soil sorption (Table 68.4); fortunately, its half-life is fairly short. Sulfentrazone is not as mobile, but its sorption appears to be affected by pH and soil mobility may be a problem at pH greater than its pKa (6.6) (Anonymous, 1995a; Grey et aI., 1997). Typical soils of soybean fields can be as high as pH 7.5. Such pH might lead to significant levels of leaching. This problem may be compounded by the relatively long half-life of sulfentrazone (Table 68.4). 68.5.2 DEGRADATION IN THE ENVIRONMENT
Protox inhibitors are not known to be a threat to the environment. The principal form of degradation is associated with microbial activity, though some of these herbicides (e.g., DPEs) are also susceptible to photodegradation (Table 68.4). The halflives of this class of herbicides vary greatly and are affected by soil quality. Half-life can be very short (i.e., less than a week for lactofen) but can be as long as 280 days (e.g., sulfentrazone) (Table 68.4) (Anonymous, 1995a). In an aquatic environment, photodegradation of most Protox inhibitors is very rapid «5 days). Some Protox inhibitors, such as JV-485 (Prosch et a!., 1997) and fomesafen (Ahrens, 1994), have low water solubility and are, therefore, considered a low risk to groundwater or surface water runoff and, consequently, to aquatic wildlife. Others, such as oxadiargyl (Dickmann et aI., 1997), dissipate rapidly from water and do not persist in the aquatic environment. Oxadiazon, oxyfluorfen, and lactofen (Ahrens, 1994) are strongly adsorbed by the soil and, therefore, do not leach into groundwater, presenting a diminished toxicological risk to aquatic wildlife.
68.5 ENVIRONMENTAL IMPACT 68.5.3 ECOTOXICOLOGY 68.5.1 INTERACTION WITH SOIL
As mentioned previously, the crop selectivity of Protox inhibitors is, for the most part, limited to soybean. Although no significant limitation on crop rotations has been associated with foliar-applied Protox inhibitors, there are some limitations with the more persistent Protox inhibitors (e.g., fomesafen
With the exception of acifiuorfen, the majority of Protox inhibitors are not harmful to avian wildlife (Table 68.5). In general, they present a low risk to the environment (Table 68.4) and terrestrial animals (Table 68.6). Of the herbicides tested against insects, fiumiclorac, fiuoroglycofen, lactofen, oxyfiuorfen, and JV-485 presented extremely low risk of toxicity.
1536
CHAPTER 68
Protoporphyrinogen Oxidase Inhibitors
Table 68.4 Fate of Protox-Inhibiting Herbicides in the Environment Common name Acifluorfena
Sorption (Koc, mIlg)
II3
Azafenidinb Bifenoxa
nla
10,000
CarfentrazoneC
nla
FlumicIoracG Fluoroglycofena
Fomesafena JV-485 d
Lactofena
Strong
1,364 60 250 10,000
Oxadiargyle Oxadiazona Oxyfluorfena Sulfentrazone f
nla
3,200 100,000 160--1928
Degradation
Mobility
Volatilization
Photo/microbiol.
Negligible
Negligible
Photo/microbiol.
Negligible
Negligible
14-60 25-40
Microbiol.
Significant
Negligible
7-14
Photo/microbiol.
Low
Negligible
60
Photo/hydrol.
Negligible
Negligible
Photo/microbiol.
Moderate
Negligible
1-6 7-21
Photo
Moderate
Negligible
100
Microbiol.
Negligible
Negligible
40-70
Microbiol.
Negligible
Negligible
3-7
Microbiol.
Negligible
Not available
Not available
Low
Negligible
Photo/microbiol.
Negligible/low
Low
Microbiol.
Moderate
Negligible
Half-life (days)
40 60 30-40 IIO-280
a Ahrens
(1994). et al. (1997). C Anonymous (I 995b). dprosch et al. (1997). eDickmann et al. (1997). f Anonymous (l995a). 8K. N. Reddy, M. A. Locke, and L. A. Gaston (1997), personal communication. b Amuti
Fomesafen was the only inhibitor tested that exhibited moderate toxicity to bees. Aquatic wildlife is generally more susceptible to Protox inhibitors than avian wildlife. Photodegradation times in water for most of these is very short «5 days). Some Protox inhibitors, such as JV-485 (Prosch et aI., 1997) and fomesafen (Ahrens, 1994), have low water solubility and are, therefore, considered a low risk to ground water or surface water runoff and, consequently, to aquatic wildlife following soil application. Others, such as oxadiargyl (Dickmann et aI., 1997), dissipate rapidly from water and do not persist in the aquatic environment. When applied directly in water, oxadiazon, oxyfiuorfen, and lactofen are strongly adsorbed by the soil and, therefore do not leach into the ground water, presenting a diminished toxicological risk to aquatic wildlife (Ahrens, 1994). Many Protox-inhibiting herbicides, such as carfentrazone and flumiclorac, are highly toxic to algae, but are only moderately toxic to fish. Others, such as sulfentrazone and fomesafen, are essentially nontoxic to fish (Ahrens, 1994; Anonymous, 1995a). Finally, bifenox and oxyfiuorfen are highly toxic to aquatic wildlife (Ahrens, 1994).
get site of these herbicides was identified, Protox inhibitors have been shown to have little acute toxicity. It is now known that the target site is Protox, the last common enzyme in the biosynthesis of heme and chlorophylls. Protox-inhibiting herbicides appear to be as inhibitory to mammalian mitochondrial Protox as to that of chloroplast (e.g., Matringe et aI., 1989b). These compounds cause an increase in porphyrin levels in animals when administered by oral doses (see Section 68.6.3). However, the herbicides appear to be effectively metabolized and/or excreted (Adler et aI., 1977; Hunt et aI., 1977; Leung et aI., 1991), and thus porphyrin levels return to normal within a few days. Interestingly, many of the primary mammalian metabolites formed are the same as photochemical degradation products (Hunt et aI., 1977). Even under exaggerated dietary doses (> 100 x recommended field rates), there appears to be little bioaccumulation risk in animals (e.g., Leung et aI., 1991). In general, for healthy individuals, these compounds are not considered to pose any significant toxicological risk.
68.6 MAMMALIAN TOXICOLOGY
All compounds have been tested under a series of mutagenicity studies, and the overwhelming weight of evidence supports the conclusion that, with the exception of oxyfiuorofen and sulfentrazone, Protox inhibitors are not genotoxic (Table 68.6). Teratology studies conducted on rat and rabbit, have demonstrated that the majority of the compounds are not teratogenic. Chronic toxicity/oncogenicity studies of carfentrazone (Anonymous, 1995b) are still in progress at the time of this writing.
68.6.1 SKIN AND ORAL
All Protox-inhibiting herbicides have passed toxicological evaluations prior to registration. Table 68.6 provides toxicological information on most commercial Protox inhibitors. Although most of these approvals were granted before the molecular tar-
68.6.2 TERATOGENICITY AND MUTAGENICITY
68.6 Mammalian Toxicology 1537 Table 68.5 Ecotoxicity of Protoporphyrin Oxidase-Inhibiting Herbicides Fish
Avian Common name
Test species
Acifluorfena
Bobwhite quail
(mg/kg) 325 4,187
Mallard duck Azafenidin h Bifenoxa Carfentrazonec Flumic10raca Fluoroglycofena Fomesafen a JV-485 d Lactofena Oxadiargyle Oxadiazona
Sulfentrazone f
Test species
(96 h)
Bluegill sunfish
62
Rainbow trout
17
Bobwhite quail
>2,500
Bluegill sunfish
48
Mallard duck
>2,500
Rainbow trout
33
Mallard duck
>5,000
Bluegill sunfish
0.64
Pheasant
>5,000
Rainbow trout
0.87
Bobwhite quail
>2,250
Bluegill sunfish
2.0
Mallard duck
>5,620
Rainbow trout
1.6 17.4
Bobwhite quail
>2,250
Bluegill sunfish
Mallard duck
>5,620
Rainbow trout
Bobwhite quail
>1,075
Bluegill sunfish
Mallard duck
>5,000
Rainbow trout
1.5
>2,000
Bluegill sunfish
>5,000
Rainbow trout
Bobwhite quail
>2,130
Bluegill sunfish
>0.045 mg a.i.ll
Mallard duck
>2,130
Rainbow trout
>0.045 mg a.i./l
Bobwhite quail
>2,510
Bluegill sunfish
>560
Mallard duck
>5,620
Rainbow trout
Bobwhite quail
>2,000
Bluegill sunfish
Below detection
Rainbow trout
Below detection
6,000 >1,000
Catfish
6,000
nJa nJa nJa nJa Honey bee
>106 f.J.g/bee
Honey bee
> 100 f.J.glbee
Honey bee
>50 f.J.g/bee
Earthworm
> 1,170 mg/kg soil
680
Honey bee
>160 f.J.g/bee
>0.1
::0:15.4
Rainbow trout
Oral LD50
23
Bobwhite quail
Bobwhite quail
Test species
1.1
Mallard duck
Mallard duck Oxyfluorfen"
Others LCso (mgfl)
Oral LD50
nJa nJa
>9
Bobwhite quail
>2,200
Bluegill sunfish
0.2
Honey bee
>IO,OOOppm
Mallard duck
>4,000
Rainbow trout
0.4
Fiddler crab
>1,000 mg/l
Bobwhite quail
>5,620
Bluegill sunfish
Mallard duck
>5,620
Rainbow troul
93.8
nJa
>130
a Ahrens
(1994). et al. (1997). c Anonymous (1995b). dProsch et al. (1997). eDickmann et al. (1997). f Anonymous (l995a). b Amuli
68.6.3 EFFECTS ON MAMMALIAN PORPHYRIN METABOLISM In healthy individuals, approved Protox-inhibiting herbicides are not considered to be of significant toxicological risk due to their effects on porphyrin metabolism. To date, no health problems have been associated with human consumption of crops treated with these compounds (Duke and Rebeiz, 1994). Mammalian Protox is as sensitive to Protox-inhibiting herbicides as chloroplastic Protox (Birchfield and Casida, 1997; Krijt et al., 1994; Scalla and Matringe, 1994), and these compounds can cause greatly elevated levels of porphyrins in animals administered with oral doses of these compounds (Krijt et aI., 1994). However, these herbicides are either not readily absorbed by the body during digestion and/or are rapidly degraded
by metabolism (Adler et aI., 1977; Hunt et aI., 1977; Leung et aI., 1991). In mammals, there are remarkable species differences in the levels of porphyrin accumulation resulting from exposure to Protox inhibitors, and developmental toxicity correlates with Proto accumulation (Kawamura et al., 1996). Rats and mice seem to be particularly sensitive to Protox inhibitors. Variegate porphyria, a human disease characterized by accumulation of Proto and other porphyrins, is caused by a deficiency of Protox. Variegate porphyria-like symptoms can be generated in mice with high doses of herbicidal Protox inhibitors (Krijt et aI., 1997). However, neither of two structurally divergent Protox inhibitors (oxadiazon and oxyfiuorfen) affected Protox activity of the brain and liver. How this relates to human risk a<;sessment is unknown. In general, however, relatively high doses
1538
CHAPTER 68
Protoporphyrinogen Oxidase Inhibitors
Table 68.6 Mammalian Toxicity of Protoporphyrin Oxidase-Inhibiting Herbicides
Common name Acifluorfena
Azafenidinb Bifenoxa
CarfentrazoneC FlumicIoraca Fluoroglycofena Fomesafena JV-485 d Lactofena Oxadiargyle Oxadiazona Oxyfluorfena Sulfentrazone f
Test
Test species
LDSO (mg/kg)
Acute oral
Rat
Acute dermal
Rabbit
Acute oral
Rat
>5,000
Acute dermal
Rabbit
>2,000
Acute oral
Rat
>5,000
Mouse
>4,556
Acute dermal
Rabbit
>2,000
Acute oral
Rat
>5,000
Acute dermal
Rat
>4,000
Acute oral
Rat/mouse
>5,000
Acute dermal
Rat
>2,000
1,540
Teratogenicity Mutagenicity Negative
(mglkg/day) nla
>2000mg/kg
Acute oral
Rat
Acute dermal
Rabbit
Acute oral
Rat
24,000
Acute dermal
Rabbit
>1,683
Acute oral
Rat
>5,000
Acute dermal
Rat
>5,000
Acute oral
Rat
>5,000
Acute dermal
Rabbit
>2,000
1,500
Negative
nla
nla
None at 200
Negative
Negative
Negative
> 1500 (rat)
Negative
Negative (rat/rabbit)
Negative
Negative
Negative
Negative
Negative
Negative
>5,000
Acute oral
Rat
>5,000
Acute dermal
Rat
>2,000
Acute oral
Rat
>5,000
Acute dermal
Rat
>8,000
Acute oral
Rat/dog
>5,000
Acute dermal
Rabbit/rat
>5,000
Acute oral
Rat
>2,000
Acute dermal
Rabbit/rat
>2,000
Negative
nla
nla
nla
Positive
Toxic at 150
Negative
Toxic at 25
a Ahrens
(1994). et al. (1997). c Anonymous (1995b). dprosch et al. (1997). eDickmann et al. (1997). f Anonymous (1995a). b Amuti
of herbicides are required to elicit an effect, and porphyrin levels return to normal within days after withdrawal of the herbicide. Finally, Protox inhibitors have been proposed as pharmaceuticals for use in tumor phototherapy (HaIling et al., 1994). Some Protox inhibitors may preferentially accumulate in tumors, resulting in sufficient differences between tumor Proto accumulation and that in adjacent tissues for exploitation in phototherapy.
68.6.4 METABOLIC DEGRADATION IN ANIMALS
There are few studies that document metabolism of DPE herbicides and other Protox inhibitors in animals and wildlife. In
general, the primary form of metabolite excretion is through urine and feces (Hunt et aI., 1977; Leung et aI., 1991). A variety of animals, including rats, rabbits, goats, sheep, cattle, and chickens, have been tested. General classes of metabolic degradation of these compounds by animals include nitro reduction, deesterification, and conjugation to glutathione, cysteine, and carbohydrates. All commercial diphenyl ether herbicides contain a p-nitrophenyl substituent. In animals, this moiety is readily reduced to an amine group. Some DPE herbicides contain a carboxyester group at the meta position on the nitrophenyl ring. This ester is readily hydrolyzed to produce a very polar carboxylic acid derivative. DPEs are also easily inactivated by cleavage of the ether bridge, followed by conjugation to glutathione (Aizawa and Brown, 1999). Other conjugated metabolites have been identified with glucuronic acid. Most of the pri-
References mary metabolites are the same as those formed in plants (see Fig. 68.3a). The DPE herbicides fluoroglycofen ethyl and bifenox are readily deesterified in animals. In fact, the initial steps of fluoroglycofen ethyl metabolism lead to the formation of acifluorofen (Aizawa and Brown, 1999). In animals, bifenox, following deesterification, forms 5-(2,4-dichlorophenoxy)-2-nitrobenzoic acid. The 4-hydroxyphenyl ether metabolite of nitrofen was identified as a major derivative in rabbits (Bray et aI., 1953). This metabolite was excreted primarily as a glucuronide conjugate, with small amounts of free 4,4' -dihydroxydiphenyl ether also present. Another study demonstrated that rabbits further degraded these metabolites to yield their corresponding hydroxy derivatives (Matsunaka, 1976). Studies of the metabolic degradation of oxyflurofen in animals showed that most of the metabolic products were excreted in the feces, with small amounts remaining in the urine (2% in males; 4% in females) (Adler et aI., 1977). A common metabolite consisted of an amino derivative of nitro-substituted primary metabolite 4-[2-chloro-4-(trifluoromethyl)phenoxy]-2ethoxybenzenamine. This amino derivative was further degraded with the consecutive conversion of the amino group to an acetamido group to yield another common metabolite (N -[4[2-chloro-4-( trifluoromethy1)phenoxy ] -2-hydroxy -phenyl] acetamide). All non-oxygen-bridged Protox inhibitors (i.e., those structurally different from DPEs) appear to follow a similar metabolic degradation pattern. In rats and goats, most of the herbicide metabolites are found in urine, with small amounts excreted in feces and milk. In chickens, approximately 95% of the metabolites are eliminated in excreta, with small amounts (0.09%) eliminated in the eggs (Leung et al., 1991). The carboxyester group of the triazolinone herbicide carfentrazone ethyl is initially metabolized to a carboxylic acid group. Further metabolites identified in rats and lactating goats included hydroxymethylpropionic acid and cinnamic acid derivatives (Aizawa and Brown, 1999). In mammals, the propionic acid metabolites undergo further oxidation of the methyl group by the cytochrome P-450. Finally, the cinnamic acid conjugate may be further metabolized to yield a benzoic acid derivative (Aizawa and Brown, 1999). Metabolism of the triazolinone herbicide sulfentrazone has been tested in rats, goats, and hens. The primary metabolite (88-95%) is 3-hydroxymethyl sulfentrazone. Other metabolites include 3-desmethyl sulfentrazone and 2,3-dihydroxymethyl sulfentrazone. Overall, triazolinone herbicides such as sulfentrazone are rapidly metabolized, with most of the compound being excreted within 3-5 days (Leung et aI., 1991).
68.7 CONCLUDING COMMENTS Protox-inhibiting herbicides may play a more important role in the future agrochemical market for several reasons. These compounds are effective at very low application rates and have
1539
generally good ecotoxicology and human toxicology profiles at recommended application rates. Many are highly compatible with the trend toward no-tillage agriculture. Furthermore, unlike with some of the other herbicide classes, weeds have not been able to evolve resistance at this particular site of action. As a result, these herbicides might replace comparable products currently available for broad-spectrum weed control in soybean fields to which weeds are rapidly becoming resistant. The success of Protox inhibitors is dependent on broadening their use to include other major crops such as maize and in enhancing the resistance of crops for which these herbicides are currently being marketed. Following the promise shown by the commercialization of genetically engineered glyphosateresistant soybean (Cole, 1994), intensive research to generate crops resistant to Protox inhibitors is in progress (e.g., Choi et aI., 1998). Although we are aware of no evidence of any significant environmental or toxicological risks of approved Protox inhibitor herbicides, the fact that all mitochondrial Protox forms tested so far are highly sensitive might provide a clue for toxicologists (mammalian and environmental) to find overlooked effects. However, the relatively low dose rates required for herbicidal activity may be far below the dose needed to adversely affect porphyrin metabolism in animals in field situations or humans as a result of exposures in food, water, air, or during application. This appears to the case with at least one other herbicide (glufosinate) that is equally effective on the plant and animal forms of the enzyme glutamine synthetase (Lydon and Duke, 1998).
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Protoporphyrinogen Oxidase Inhibitors
Bray, H. G., James, S. P., Thorpe, W. V., and Wasdell, M. R. (1953). The metabolism of ethers in the rabbit. Biochem. J. 54, 547-551. Choi, K. w., Han, 0., Lee, H. J., Yun, Y. C., Moon, Y. H., Kim, M., Kuk, Y. 1., Han, S. U., and Guh, J. O. (1998). Generation of resistance to the diphenyl ether herbicide, oxyfluorfen via expression of the Bacillus subtilis protoporphyrinogen oxidase gene in transgenic: on tobacco plants. Biosci. Biotechnol. Biochem. 62, 558-560. Clark, R. D. (1994). Synthesis of protoporphyrinogen oxidase inhibitors. Am. Chem. Soc. Symp. Ser. 559, 34-47. Clark, R. D. (1996). Synthesis and QSAR of herbicidal 3-pyrazoyl a, a, atrifluorotolyl ethers. J. Agric. Food Chem. 44, 3643-3652. Cole, D. J. (1994). Introduction of herbicide resistant crops. Pestic. Outlook 5, 32-36. Condon, M. E., Alvarado, S. 1., Arthen, F J., Birk, J. H., Brady, T. E., Crews, A. D., Marc, J. H., Karp, G. M., Lavanish, J. M., Nielsen, D. R., and Lies, T. A. (1995). 6-aryloxy-1H-benzotriazoles. Am. Chem. Soc. Symp. Ser. 584, 123-135. Cox, G. S., Krieg, M., and Whitten, D. G. (1982). Self-sensitized photooxidation of protoporphyrin IX derivatives in aqueous surfactant solutions: Product and mechanistic studies. 1. Am. Chem. Soc. 104,6930-6937. Dayan, F. E., and Duke, S. O. (1996). Porphyrin-generating herbicides. Pestic. Outlook 7,22-27. Dayan, F E., and Duke, S. O. (1997). Phytotoxicity of protoporphyrinogen oxidase inhibitors: Phenomenology, mode of action and mechanisms of resistance. In "Herbicide Activity: Toxicology, Biochemistry and Molecular Biology" (R. M. Roe, J. D. Burton, and R. J. Kuhr, eds.), pp. 11-35. IOS Press, The Netherlands. Dayan, F. E., Armstrong, B. M., and Weete, J. D. (1998). Inhibitory activity of sulfentrazone and its metabolic derivatives on soybean (Glycine max) protoporphyrinogen oxidase. J. Agric. Food Chem. 46, 2024-2029. Dayan, F. E., Duke, S. 0., Reddy, K. N., Hamper, B. C., and Leschinsky, K. L. (1997a). Effect of isoxazole herbicides on protoporphyrinogen oxidase and porphyrin physiology. J. Agric. Food Chem. 45,967-975. Dayan, F. E., Duke, S. 0., Weete, J. D., and Hancock, H. G. (1997b). Selectivity and mode of action of carfentrazone-ethyl, a novel phenyl triazolinone herbicide. Pestic. Sci. 51,65-73. Dayan, F. E., Reddy, K. N., and Duke, S. O. (1999). Structure-activity relationships of diphenyl ethers and other oxygen-bridged protoporphyrinogen oxidase inhibitors. In "Peroxidizing Herbicides" (P. Boger and K. Wakabayashi, eds.), pp. 144-162. Springer-Verlag, Berlin. Dayan, FE., Weete, J. D., Duke, S. 0., and Hancock, H. G. (I997c). Soybean (Glycine max) cultivar differences in response to sulfentrazone. Weed Sci. 45,634-641. Dayan, F. E., Weete, J. D., and Hancock, H. G. (1996). Physiological basis for differential sensitivity to sulfentrazone by sicklepod (Senna obtusifolia) and coffee senna (Cassia occidentalis). Weed Sci. 44, 12-17. Devine, M. D., Duke, S. 0., and Fedtke, C. (1993). Oxygen toxicity and herbicidal action. In "Physiology of Herbicide Action," pp. 177-189. Prentice Hall, Englewood Cliffs, NJ. Deybach, J. c., de Vemeuil, H., and Nordmann, Y. (1981). The inherited enzymatic defect in porphyria variegata. Hum. Genet. 58, 425-428. Dickmann, R., Melgarejo, J., Loubiere, P., and Montagnon, M. (1997). Oxadiargyl, a novel herbicide for rice and sugarcane. In "Brighton Crop Protection Conference," pp. 51-57. Duke, S. 0., and Kenyon, W. H. (1986). Photosynthesis is not involved in the mechanism of action of acifluorfen in cucumber (Cucumis sativus L.). Plant Physiol. 81, 882-888. Duke, S. 0., and Rebeiz, C. A. (1994). Porphyric pesticides: Chemistry, toxicology, and pharmaceutical applications. Am. Chem. Soc. Symp. Series 559. Duke, S. 0., Lee, H. J., Duke, M. v., Reddy, K. N., Sherman, T. D., Becerril, J. M., Nandihalli, U. B., Matsumoto, H., Jacobs N. J., and Jacobs, J. M. (1997). Mechanisms of resistance to protoporphyrinogen oxidase-inhibiting herbicides. In "Herbicide Resistance in Crops and Weeds" (R. DePrado, L. Garda-Torres, and J. Jorrin, eds.), pp. 155-160. Kluwer Academic, Amsterdam.
Duke, S. 0., Lydon, J., Becerril, J., Sherman, T. D., Lehnen, L. P., and Matsumoto, M. (1991). Protoporphyrinogen oxidase-inhibiting herbicides. Weed Sci. 39, 465-473. Frear, D. S., and Swanson, H. R. (1973). Metabolism of substituted diphenyl ether herbicides in plants. 1. Enzymic cleavage of fluorodifen in peas (Pisum sativum). Pestic. Biochem. Physiol. 3,473-482. Frear, D. S., Swanson, H. R., and Mansager, E. R. (1983). Acifluorfen metabolism in soybean: Diphenyl ether bond cleavage and the formation of homoglutathione, cysteine and glucose conjugates. Pestic. Biochem. Physiol. 20,299-310. Gilham, D. J., and Dodge, A. D. (1987). The mode of action of nitro-diphenyl ether herbicides. In "Progress in Pesticide Biochemistry and Physiology" (D. H. Hutson and T. R. Roberts, eds.), Vol. 7, pp. 147-167. Wiley, New York. Grey, T. L., Walker, R. H., Wehtje, G. R., and Hancock, H. G. (1997). Sulfentrazone adsorption and mobility as affected by soil and pH. Weed Sci. 45, 733-738. Hailing, B. P., Yuhas, D. A., Fingar, V. F., and Winkelmann, J. W. (1994). Protoporphyrinogen oxidase inhibitors for tumor therapy. Am. Chem. Soc. Symp. Ser. 559, 280-290. Hamper, B. c., Leschinsky, K. L., Massey, S. S., Bell, C. L., Brannigan, L. H., and Prosch, S. D. (1995). Synthesis and herbicidal activity of 3-aryl-5(haloalkyl)-4-isoxazolecarboxamides and their derivatives. J. Agric. Food Chem. 43,219-228. Hunt, L. M., Chamberlain, W. F., Gilbert, B. N., Hopkins, D. E., and Gingrich, A. R. (1977). Absorption, excretion, and metabolism of nitrofen by a sheep. J. Agric. Food Chem. 25, 1062-1065. Ichinose, K., Che, F-S., Kimura, Y., Matsunobu, A., Sato, F., and Yoshida, S. (1995). Selection and characterization of protoporphyrinogen oxidase inhibiting herbicide (S23142) resistant photomixotrophic cultured cells of Nicotiana tabacum. J. Plant Physiol. 146, 693-698. Jacobs, J. M., Jacobs, N. J., and Duke, S. O. (1996). Protoporphyrinogen destruction by plant extracts and correlation with tolerance to protoporphyrinogen oxidase inhibiting herbicides. Pestic. Biochem. Physiol. 55, 77-83. Jacobs, J. M., Jacobs, N. J., Sherman, T. D., and Duke, S. O. (1991). Effect of diphenyl ether herbicides on oxidation of protoporphyrinogen to protoporphyrin in organellar and plasma membrane-enriched fractions of barley. Plant Physiol. 97, 197-203. Johnson, W. 0., Kollman, G. E., Swithenbank, c., and Yih, R. Y. (1978). RH6201 (Blazer): A new broad spectrum herbicide for postemergence use in soybean. J. Agric. Food Chem. 26, 285-286. Karp, G. M., Condon, M. E., Arthen, F J., Birk, J. H., Marc, P. A., Hunt, D. A., Lavanish, J. M., and Schwindeman, J. A. (1995). Aryloxyindolin-2(3H)ones. Am. Chem. Soc. Symp. Ser. 584, 136-148. Kawamura, S., Kato, T., Matsuo, M., Katsuda, Y., and Yasuda, M. (1996). Species difference in protoporphyrin IX accumulation produced by an N -phenylimide herbicide in embryos between rats and rabbits. Toxicol. Appl. Pharmacol. 141,520-525. Kenyon, W. H., Duke, S. 0., and Vaughn, K. C. (1985). Sequences of effects of acifluorfen on physiological and ultrastructural parameters in cucumber cotyledon discs. Pestic. Biochem. Physiol. 24, 240-250. Krijt, J., Stranska, P., Maruna, P., Vokurka, M., and Sanitnik, J. (1997). Herbicide-induced experimental variegate porphyria in mice: Tissue porphyrinogen accumulation and response to porphyrogenic drugs. Can. J. Physiol. Pharmacol. 75, 1181-1187. Krijt, J., Vokurda, M., Sanitriik, J., and Janousek, V. (1994). Effect of protoporphyrinogen oxidase inhibition on mammalian porphyrin metabolism. Am. Chem. Soc. Symp. Series. 559,247-254. Kunert, K. J., Sandmann, G., and Boger, P. (1987). Modes of action of diphenyl ethers. Rev. Weed Sci. 3,35-55. Lee, H. J., Duke, M. v., Birk, J. H., Yamamoto, M., and Duke, S. O. (1995). Biochemical and physiological effects ofbenzheterocycles and related compounds. 1. Agric. Food Chem. 43, 2722-2727. Lee, H. J., Duke, M. v., and Duke, S. O. (1993). Cellular localization of protoporphyrinogen-oxidizing activities of etiolated barley (Hordeum vulgare L.) leaves. Plant Physiol. 102,881-889.
References Leung, L. Y., Lyga, J. w., and Robinson, R. A. (1991). Metabolism and distribution of the experimental triazolinone herbicide sulfentrazone in the rat, goat and hen. 1. Agric. Food Chem. 39, 1509-1514. Lydon, J., and Duke, S. O. (1998). Inhibitors of glutamine biosynthesis. In "Plant Amino Acids: Biochemistry and Biotechnology" (B. K. Singh, ed.), pp. 445-463. Dekker, New York. Lyga, J. w., Patera, R. M., Theodoridis, G., Hailing, B. P., Hotzman, E w., and Plummer, M. J. (1991). Synthesis and quantitative structureactivity relationships of herbicidal N -(2-fluoro-5-methoxyphenyl)-3,4,5,6tetrahydrophthalimides. J. Agric. Food Chem. 39, 1667-1673. Matringe, M., and Scalla, R. (1988). Effects of acifluorfen-methyl on cucumber cotyledons: Protoporphyrin accumulation. Pestic. Biochem. Physiol. 32, 164-168. Matringe, M., Camadro, J.-M., Labbe, P., and Scalla, R. (l989a). Protoporphyrinogen oxidase as a molecular target for diphenyl ether herbicides. Biochem. J. 260, 231-235. Matringe, M., Camadro, J.-M., Labbe, P., and Scalla, R. (l989b). Protoporphyrinogen oxidase inhibition by three peroxidizing herbicides: Oxadiazon, LS 82-556 and M&B 39279. FEBS Lelt. 245,35-38. Matringe, M., Mornet, R., and Scalla, R. (1992). Characterization of [3H]acifluorfen binding to purified pea etioplasts, and evidence that protoporphyrinogen oxidase specifically binds acifluorfen. Eur. J. Biochem. 209, 861-868. Matsumoto, H., Lee, J. J., and Ishizuka, K. (1994). Variation in crop response to protoporphyrinogen oxidase inhibitors. Am. Chem. Soc. Symp. Ser. 559, 120-132. Matsunaka, S. (1969). Acceptor of light energy in photoactivation of diphenyl ether herbicides. J. Agric. Food Chem. 17, 171-175. Matsunaka, S. (1976). Diphenyl ether herbicides. In "Herbicides: Chemistry, Degradation and Mode of Action" (P. C. Keamey and D. D. Kaufman, eds.), Vol. 2, pp. 709-739. Dekker, New York. Mito, N., Sato, R., Miyakado, M., Oshio, H., and Tanaka, S. (1991). In vitro mode of action of N -phenylimide photobleaching herbicides. Pestic. Biochem. Physiol. 40, 128-135. Nandihalli, U. B., and Duke, S. O. (1993). The porphyrin pathway as a herbicide target site. Am. Chem. Soc. Symp. Ser. 524, 62-78. Nandihalli, U. B., Duke, M. v., and Duke, S. O. (1992). Quantitative structureactivity relationships of protoporphyrinogen oxidase-inhibiting diphenyl ether herbicides. Pestic. Biochem. Physiol. 43, 193-211. Pornprom, T., Matsumoto, H., Usui, K., and Ishizuka, K. (1994). Characterization of oxyfluorfen tolerance in selected soybean line. Pestic. Biochem. Physiol. 50, 107-114. Prosch, S. D., Ciha, A. J., Grogna, R., Hamper, B. c., Feucht, D., and Dreist, M. (1997). JV 485: A new herbicide for pre-emergence broad spectrum weed control in winter wheat. In "Brighton Crop Protection Conference," pp. 4550. Reddy, K. N., Dayan, E E., and Duke, S. O. (1998). QSAR analysis of protoporphyrinogen oxidase inhibitors. In "Comparative QSAR" (J. Devillers, ed.), pp. 197-234. Taylor & Francis, London. Reddy, K. N., Nandihalli, U. B., Lee, H. J., Duke, M. v., and Duke, S. O. (1995). Predicting activity of protoporphyrinogen oxidase inhibitors by computeraided molecular modeling. Am. Chem. Soc. Symp. Series. 589, 221-224. Ritter, R. L., and Coble, H. D. (1981a). Influence of temperature and relative humidity on the activity of acifluorfen. Weed Sci. 29, 480-485. Ritter, R. L., and Coble H. D. (l98Ib). Penetration, translocation, and metabolism of acifluorfen in soybean (Glycine max), common ragweed
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(Ambrosia artemissiJolia), and cocklebur (Xanthium pensylvanicum). Weed Sci. 29, 474-480. Sato, R., Nagano, E., Oshio, H., and Kamoshita, K. (1987). Diphenylether-like physiological and biochemical actions of S-23142, a novel N-phenylimide herbicide. Pestic. Biochem. Physiol. 28, 194-200. Scalla, R., and Matringe, M. (1994). Inhibitors of protoporphyrinogen oxidase as herbicides: Diphenyl ethers and related photobleaching molecules. Rev. Weed Sci. 6, 103-132. Sherman, T. D., Duke, M. v., Clark, R. D., Sanders, E. E, Matsumoto, H., and Duke, S. O. (1991). Pyrazole phenyl ether herbicides inhibit protoporphyrinogen oxidase. Pestic. Biochem. Physiol. 40,236-245. Sumida, M., Niwata, S., Fukami, H., Tanaka, T., Wakabayashi, K., and Boger, P. (1995). Synthesis of novel diphenyl ether herbicides. J. Agric. Food Chem. 43, 1929-1934. Theodoridis, G. (1997). Structure-activity relationships of herbicidal aryltriazolinones. Pestic. Sci. 50, 283-290. Theodoridis, G., Bahr, J. T., Davidson, B. L., Hart, S. E., Hotzman, E W., Baum, J. S., Hotzman, E W., Poss, K. M., and Tutt, S. E (1995). Alkyl 3-[2,4-disubstituted-4,5-dihydro-3-methyl-5-oxo-1 H -1,2,4triazol-I-yl)phenyl]propenoate derivatives: Synthesis and structure-activity relationships. Am. Chem. Soc. Symp. Ser. 584,90-99. Theodoridis, G., Baum, J. S., Hotzman, E W., Manfredi, M. c., Maravetz, L. L., Lyga, J. w., Tymonko, J. M., Wilson, K. R., Poss, K. M., and Wyle, M. J. (1992). Synthesis and herbicidal properties of aryltriazolinones. A new class of pre- and postemergence herbicides. Am. Chem. Soc. Symp. Ser. 504, 135-146. Vidrine, P. R., Griffin, J. L., Jordan, D. L., and Reynolds, D. B. (1996). Broadleaf weed control in soybean (Glycine max) with sulfentrazone. Weed Technol. 10, 762-765. Vidrine, P. R., Jordan, D. L., and Girlinghouse, J. M. (1994). Efficacy ofF-6285 in soybeans. Proc. South. Weed Sci. Soc. 47, 62. Vidrine, P. R., Reynolds, D. B., and Griffin, J. L. (1993). Weed control in soybean (Glycine max) with lactofen plus chlorimuron. Weed Technol. 7, 311-316. Walker, R. H., Richburg, J. S., and Jones, R. E. (1992). F6285 efficacy as affected by rate and method of application. Proc. South. Weed Sci. Soc. 45, 51. Ward, E. R., and Volrath, S. (1995). "Manipulation of Protoporphyrinogen Oxidase Enzyme Activity in Eukaryotic Organisms." International Patent Application WO 95/34659. Watanabe, N., Che, E S., Iwano, M., Nakano, T., Takayama, S., Yoshida, S., and Isogai, A. (1998). Molecular characterization ofphotomixotrophic tobacco cells resistant to photoporphyrinogen oxidase-inhibiting herbicides. Plant Physiol. 118,451-458. Wehtje, G. R., Walker, R. H., Grey, T. L., and Hancock, H. G. (1997). Response of purple (Cyperus rotundus) and yellow nutsedge (Cyperus esculentus) to selective placement of sulfentrazone. Weed Sci. 45, 382-387. Wepplo, P., Birk, J. H., Lavanish, J. M., Mandredi, M., and Nielsen, D. R. (1995). 5-Aryloxybenzisoxazole esters. Am. Chem. Soc. Symp. Ser. 584, 149-160. Wills, G. D., and McWhorter, C. G. (1981). Effect of environment on the translocation and toxicity of acifluorfen to showy crotalaria (Crotalaria spectabilis). Weed Sci. 29, 397-401.
CHAPTER
69 Chloracetanilides William F. Heydens Monsanto Company
Ian C. Lamb Pioneer Hi-Bred International, Inc.
Alan G. E. Wilson Pharmacia Corporation
69.1 INTRODUCTION
69.2 ALACHLOR 69.2.1 IDENTITY, PROPERTIES, AND USES
This chapter describes the toxicology of several chloracetanilide herbicides, a subclass of the acetamides [general structure RI-C(O)-N(R2R3)] that have as a common structural feature a CIH2C group as the RI substitution. Information is provided for alachlor, acetochlor, butachlor, metolachlor, and propachlor (see Fig. 69.1 for chemical structures); all of these herbicides, with the exception of butachlor, are sold in the United States, used primarily on corn, and collectively have the largest share in this market. The herbicidal mode of action for chloracetaniiides is not totally understood. It is known that this class of herbicides inhibits the biosynthesis of lipids, a1cohols, fatty acids, proteins, isoprenoids, and flavonoids. By inhibiting synthesis of various terpenoid precursors (e.g., kaurene), these herbicides appear to interfere with the production of gibberellin. Terpenes and waxes are formed via different biosynthetic pathways both using coenzyme A intermediates and substrates; interference with the synthesis of both substances may indicate a common mechanism of inhibition through actions on coenzyme A. Furthermore, it has been shown that chloracetaniiides are detoxified in plants by conjugation with glutathione. This has also led to the suggestion that these compounds cause their herbicidal effect via conjugation of acctyl coenzyme A and other sulfhydryl-containing enzymes, with consequent inhibition of some critical function needed for the germination or survival of seedlings.
Handbook of Pesticide Toxicology Volume 2. Agents
69.2.1.1 Chemical Name Alachlor is N -methoxymethyl-2',6'-diethyl-2-chloroacetanilide.
69.2.1.2 Structures See Fig. 69.1.
69.2.1.3 Synonyms The common name alachlor is in general use. The major trade name for alachlor products in the United States is Lasso®. The CAS registry number for alachlor is 15972-60-8.
69.2.1.4 Physical and Chemical Properties Alachlor has the empirical formula of Cl4H20N02Cl and a molecular weight of 269.8. It is an odorless solid at room temperature with a melting point of approximately 38°C and has a vapor pressure of 1.6 x 10-5 mm Hg at 25°C. The solubility of alachlor in water is 242 ppm at 25°C. Alachlor is also soluble in ether, acetone, benzene, chloroform, ethanol, and ethyl acetate; it is slightly soluble in heptane.
69.2.1.5 History, Formulations, and Uses Alachlor was registered and introduced in 1967 for the preplant or preemergence control of a broad spectrum of grass, sedge, and broadleaf weeds. It is used in corn, soybeans, dry beans, cotton, sorghum, sunflowers, peanuts, and other crops.
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Copyright © 200 1 by Academic Press. All rights of reproduction in any fonn reserved.
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CHAPTER 69 Chloracetanilides 69.2.2.3 Repeated Dose Studies
Alachlor
Acetochlor
Butachlor
Metolachlor
Propachlor
Several subchronic and chronic toxicology studies of alachlor have been conducted, the results of which have recently been reported by the V.S. Environmental Protection Agency EPA (1998a) and Heydens (1998). The major studies are summarized next. Administration of alachlor to beagle dogs for 6 months at dose levels of 5, 25, 50, and 75 mg/kg/day produced changes indicative of liver toxicity; the effect at the lowest dose was limited to a slight elevation in the liver weights of male dogs only. In a subsequent I-year dog study, a no observable effect level (NOEL) was established at 1 mg/kg/day based on evidence of slight anemia at 3 mg/kg/day in two animals. In the first of two chronic studies conducted in Long-Evans rats, alachlor was administered in the diet at doses of 14, 42, and 126 mg/kg/day for approximately 2 years. Hepatotoxicity was evident at all dose levels. Eye examinations revealed the presence of an ocular lesion, progressive uveal degeneration syndrome (VDS). This syndrome was noted at the two highest doses tested, and may also have occurred in two rats at 14 mg/kg/day. This ocular lesion was considered to be unique to the Long-Evans rat because the response has not been observed in other strains of rats, mice, or dogs. Furthermore, the effect has not been observed in humans involved in the manufacture of alachlor (see Section 69.2.3). The second long-term feeding study in rats was conducted at dose levels of 0.5, 2.5, and 15 mg/kg/day for approximately 25 months. Liver and ocular effects were not observed at any dose level; the NOEL was 2.5 mg/kg/day based on nasal hyperplasia and submucosal gland hyperplasia at the highest dose tested. In conclusion, the lowest NOEL for all subchronic and chronic effects was determined to be 1 mgikg/day in the I-year dog study. 69.2.2.4 Pharmacokinetic Studies
Figure 69.1
Structures of Acetanilides.
69.2.2 TOXICITY TO LABORATORY ANIMALS 69.2.2.1 Irritation and Sensitization Eye and skin irritation studies conducted in rabbits showed alachlor to be nonirritating to the eye and slightly irritating to the skin. Alachlor produced skin sensitization in guinea pigs (Ahrens, 1994). 69.2.2.2 Acute Studies Acute toxicity data have been reported by Ahrens (1994). The oral LDso in rats ranges from 930 to 1350 mg/kg, while the dermal LDso is reported to be 13,300 mg/kg. The 4-h inhalation LCso in rats was shown to be greater than 5.1 mg/l, the highest concentration tested (Monsanto, 1997a).
The absorption, metabolism, and excretion of alachlor has been extensively studied in rats, mice, and monkeys (EPA, 1998a; Heydens et aI., 1998). Alachlor is well absorbed in rats following oral administration. The metabolism of alachlor in rats is complex due to extensive biliary excretion, intestinal microbial metabolism, and enterohepatic circulation of metabolites. In excess of 30 metabolites of alachlor have been identified in rat excreta, with approximately equal quantities appearing in urine and feces. Nearly 90% of the administered dose was eliminated in lO days. Qualitatively, alachlor metabolism in the mouse is similar to the rat; however, there are significant quantitative differences between the two species. In contrast, alachlor is metabolized in monkeys to a limited number of glutathione and glucuronide conjugates, which are excreted primarily via the kidney. Excretion in monkeys is more rapid than in rodents, with 2:90% or more of an administered dose being excreted in the urine within 48 h. The large differences in metabolic profile patterns and urinary excretion rates between rats and monkeys is thought to be due to a physiological phenomenon commonly referred to
69.2 Alachlor as the molecular weight threshold for biliary excretion. Being of intermediate molecular weight (i.e., 300-500 g/mol), alachlor metabolites have been shown to undergo biliary excretion and enterohepatic recirculation in rodents, but are not good candidates for biliary excretion in primates (Millburn, 1975; Williams, 1971). The dermal penetration of alachlor has been investigated in monkeys and shown to be relatively low. One study showed that penetration rates (i.e., percentage of applied dose that is absorbed) for the emu1sifiable concentrate (EC) formulation over a 12-h period were 7.7 and 9.1 % for the undiluted formulation and diluted spray solution, respectively; values of 2.7 and 5.0% were reported for the microencapsulated product (Kronenberg et al., 1988). Another study done with a spray dilution of the EC formulation reported that the penetration rate ranged from 15 to 21 % after a continuous 24-h exposure (Wester et al., 1992).
69.2.2.5 Genotoxicity Studies Numerous genetic toxicology studies have been conducted that assessed a variety of in vitro and in vivo endpoints. These include studies generated for regulatory purposes using established testing guidelines conducted under good laboratory practices (GLPs) and studies published in the scientific literature, some of which have a limited amount of validation data and/or employ questionable assay conditions. The results of all these studies provide no evidence that neoplastic responses observed in the rat (discussed later) arise through a genotoxic mode of action, and the weight of evidence indicates that alachlor does not have general genotoxic potential in mammals. The most relevant studies for evaluating genotoxic potential are well-validated, standard assays required/recommended by regulatory authorities worldwide. Alachlor has been tested in several of these test systems (Kier et al., 1996). Ames/Salmonella assays conducted on alachlor as well as urine and bile samples from alachlor-treated rats showed no mutagenic activity. Alachlor was negative in a CHOIHGPRT mammalian cell gene mutation assay when tested up to cytotoxic levels. In an in vivo cytogenetics assay, alachlor was not clastogenic to rat bone marrow cells when given orally at dose levels up to 1000 mg/kg. Alachlor was negative in rat and mouse micronucleus assays conducted at doses of 600 mg/kg (ip) and 1000 mg/kg (po), respectively. In two in vivo/in vitro UDS assays, variable responses were seen at 1000 mg/kg, but alachlor was clearly negative at all other doses below that level. The 1000 mg/kg dose is near the oral LDso and has been shown to produce severe hepatotoxicity under the conditions employed in the UDS assay. Therefore, the biological relevance of the results at this high dose is doubtful. Two specialized studies have also been conducted to assess possible interactions with nasal tissue DNA in vivo (Heydens et al., 1998). The first study was conducted to determine if alachlor bound to DNA in rat nasal tissue. Following administration of 14C-Iabeled alachlor at a dose of 125 mg/kg, DNA and protein were purified and harvested from nasal turbinate tissue. There was an extremely low level of net radioactivity
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«5 dprnlmg DNA) associated with nasal tissue DNA, the nature of the bound radioactivity was not determined, and it was difficult to eliminate the possibility that the apparent binding represented protein contamination. Nevertheless, the results do represent an upper bound on the possible level of DNA binding. The possible biological significance of this DNA-associated radioactivity was evaluated by comparing the covalent binding index (CBI) to alachlor's oncogenic potential expressed as the TDso (Gold et al., 1984). Using a relationship developed for hepatocarcinogens (Lutz, 1986), it was concluded that the radioactivity was much too low to be consistent with a genotoxic mode of action for the induction of nasal tumors in rats. Thc second study investigated the ability of alachlor to produce DNA damage following administration in the diet at a dose of 126 mg/kg/day for 7 days. No evidence of DNA strand breakage was observed in the nasal cells. These studies support the conclusion that alachlor produces nasal tumors in rats via a mechanism that does not involve the initial induction of DNA damage. A number of studies evaluating the genotoxic potential of alachlor have been reported in the literature. General conclusions from the major studies can be summarized as follows: Alachlor has been tested for mutagenicity in 14 bacterial test systems. While one spot test and one plate incorporation test were reported positive, the other 12 tests showed a uniform and consistent pattern of negative results. Both positive and negative results have been reported for a number of in vitro studies conducted to assess chromosome effects. Some of the studies reporting positive effects were conducted using alachlor samples that were produced by inexpensive, alternative manufacturing processes that are not used by Monsanto. One of these processes was known to involve the use of the alkylating agent, chloromethyl methyl ether, which is a known mutagen and carcinogen. The test materials used in these studies (Georgian et al., 1983; Lin et al., 1987) were substantially more toxic (10- to > lOO-fold) to the mammalian cells tested than alachlor produced by a high-quality manufacturing process. Therefore, results from these studies are not applicable to quality-produced, name brand alachlor products and should not be included in an assessment of alachlor's genotoxic potential. Conflicting results have been reported for two other in vitro mammalian chromosome effect studies using alachlor of high purity (Erexson et al., 1993; Meisner et al., 1992). Use of nonstandard procedures (e.g., extended treatment period) and an unusual frequency of aberrant human lymphocytes in the control group of the study reporting positive effects were undoubtedly important factors. DNA strand breakage has also been reported in vitro at concentrations eliciting cytotoxicity and which, using pharmacokinetic analyses, would be associated with lethality in vivo (Bonfanti et al., 1992). In vivo DNA strand breakage studies in rats and mice showed no evidence of DNA damage (Taningher et al., 1993). These studies, along with the other negative in vivo work described previously, clearly demonstrate that the in vitro DNA damage findings reported by some investigators are not reflective of the universally negative in vivo mammalian effects.
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CHAPTER 69
Chloracetanilides
69.2.2.6 Carcinogenicity Studies
The oncogenic potential of alachlor has been assessed in two bioassays conducted with the Long-Evans strain of rat and in two studies done with CD-l mice (EPA, 1998a; Heydens, 1998). Alachlor produced significant increases in glandular stomach and thyroid follicular tumors in rats at the highest dose tested, a level exceeding the maximum tolerated dose (MTD); nasal epithelial (olfactory) tumors were also observed at lower doses. Based on these findings, the EPA had previously classified alachlor as a Group B2 carcinogen. However, the EPA's Cancer Peer Review Committee (CPRC) reconsidered the weight of evidence for alachlor, taking into account new mechanistic information (Section 69.7) in accordance with its Proposed Guidelines for Carcinogen Risk Assessment (EPA, 1996). Alachlor was reclassified as "likely" at high doses but "not likely" at low doses to be a human carcinogen. The term "low doses" denotes anticipated human exposures resulting from pesticide use. The EPA, in the Reregistration Eligibility Decision (RED) document for alachlor, agreed that a nonlinear approach (margin of exposure, or MOE) should be used for the purpose of risk assessment (EPA, 1998a). In the first rat bioassay, alachlor was administered in the diet at dose levels of 14, 42, and 126 mg/kg/day. Surviving male rats were sacrificed after 27 months while females were sacrificed after 25 months on study. In the second study, which was conducted to follow-up on non-neoplastic effects observed in the first bioassay, rats received alachlor in their diets at dose levels equivalent to 0.5, 2.5, 15, and 126 mg/kg/day. The highest dose level of 126 mg/kg/day exceeded the maximum tolerated dose (MTD) as evidenced by excessive body weight loss (>30% below controls), hepatocellular necrosis, and decreased survival. Neoplastic responses attributable to alachlor administration were observed in the nasal turbinate mucosa and glandular stomach mucosa of both sexes and in the thyroid follicular epithelium of male rats. Significant increases in stomach and thyroid tumors were restricted to the highest dose tested. One benign nasal tumor was noted at 2.5 mg/kg/day. Although single nasal tumors of this type are occasionally observed in control animals, the tumor at 2.5 mg/kg/day was considered to be treatment related by the EPA and the definitive NOEL for oncogenicity is, therefore, 0.5 mg/kg/day. Subsequent modeof-action investigations have shown that the nasal, stomach, and thyroid tumors are produced via non-genotoxic, thresholdsensitive mechanisms. The studies supporting this conclusion are described in Section 69.7. The first oncogenicity study in mice was conducted at dose levels of 26, 78, and 260 mg/kg/day for approximately 19 months. The only notable finding was the occurrence of benign lung (bronchoalveolar) tumors in high-dose females. Although the incidence was statistically different from concurrent controls, it was within the range expected for untreated animals. Due to this equivocal finding and the overall poor survival of animals in this study, a second mouse bioassay was conducted at dose levels of 20, 78, and 331 mg/kg/day for 18 months. There was no dose-response relationship for the incidence of lung tumors and no other indication that the tumors were related to
administration of the test material. Based on all the available information, it was concluded that alachlor was not oncogenic in the mouse. 69.2.2.7 Development and Reproduction Studies
The reproductive and developmental toxicity database has been reported by Heydens (1998) and recently evaluated by the EPA (1998a). The EPA's assessment included special consideration of possible effects on infants and children as required by the Food Quality Protection Act (FQPA) of 1996. This provision of the FQPA requires the use of an additional safety factor for the protection of infants and children when warranted by the severity of effects observed in toxicology studies. The EPA considered the alachlor database to be complete, and the NOELs for developmental effects were equal to or greater than those for maternal effects in both developmental toxicity studies. Therefore, the EPA concluded there is no unique sensitivity from prenatal exposure. Likewise, in the reproduction study, the reproductive NOEL is greater than the systemic NOEL. Thus, no special sensitivity for infants or children was indicated. Brief details of the studies supporting these conclusions are given next. Alachlor was fed to male and female rats at doses of 3, 10, and 30 mg/kg/day throughout premating, mating, gestation, and lactation periods for three successive generations. Nephritis and an apparent decrease in ovarian weights were noted in high-dose adults. The significance of the latter finding is doubtful because there was no microscopic change in the tissue and no effects on reproductive parameters. The systemic and reproductive toxicity NOELs were 10 and 30 (or more) mg/kg/day, respectively. In a developmental toxicity study with rats, alachlor was administered by gavage at doses of 50, 150, and 400 mg/kg/day on gestation days 6-19. Maternal and fetal toxicity were noted at the highest dose tested as evidenced by maternal deaths and decreased body weight gains, a slight decrease in fetal body weight, and a slight increase in postimplantation loss. The NOEL for developmental toxicity in the rat was 150 mg/kg/day. Oral administration of alachlor to rabbits at doses of 50,100, and 150 mg/kg/day on gestation days 7-19 produced maternal toxicity at the highest dose tested but no effects on the fetus. Therefore, the NOEL for developmental toxicity in the rabbit was greater than or equal to 150 mg/kg/day. 69.2.3 HUMAN EXPERIENCE
A large number of workers have had long-term occupational exposure to alachlor at the Monsanto manufacturing facility for this herbicide, which has been in continuous operation since 1968. An exposure analysis indicated that exposure of these workers exceeded that of farmers and herbicide applicators (Acquavella et at., 1994). These manufacturing workers therefore presented an opportunity to assess alachlor's potential to produce adverse health affects in humans. Three epidemiological studies were conducted on these individuals, the focus of which was the ocular and oncogenic effects observed in the chronic rat studies.
69.3 Acetochlor In an evaluation of ocular health, the eyes of 135 workers judged to have the highest alachlor exposure were examined for the presence of a specific eye abnormality termed pigmentary dispersion syndrome (PDS) (Ireland et aI., 1994). This lesion is analogous to the initiating lesion that occurred in one specific strain of rat from the chronic alachlor studies. Eye examinations were also given to an unexposed control group. PDS was not found in any of the exposed workers, and other eye abnormalities occurred at similar rates in exposed and unexposed individuals. These results indicate that humans exposed to alachlor are not at an increased risk of developing ocular disease. Mortality and cancer incidence studies for the period 19701990 were originally reported in 1994 (Leet et aI., 1996) and have been updated with additional data through 1993 (Acquavella et aI., 1996). The mortality cohort comprised 1199 workers employed for at least 1 year between 1961 and 1993. The cancer incidence cohort was a subset of 1169 of the mortality cohort whose members lived in Iowa for some time during the period 1969-1993. Using an in-depth knowledge of the plant and process, all job titles and descriptions used in personnel records were allocated into occupational exposure categories. Each such category was assigned a qualitative exposure rank (high, medium, low, or negligible) for alachlor. A total of 1036 workers who had potential alachlor exposure met the criteria for inclusion in the mortality analysis. Mortality from all causes among workers judged to have high alachlor exposure was lower than expected (standardized mortality ratio, SMR = 0.7), and mortality from cancer was similar to that for the state of Iowa (SMR = 0.9). Likewise, there was no increased cancer mortality among those workers with 5 or more years of exposure and 15 or more years since first exposure, the group in which occupationally related cancers would be mostly to occur. There were no deaths from nasal, stomach, or thyroid cancers. There were 1025 workers who met the criteria for cancer incidence analysis, and 701 (68%) of them belonged to the high-exposure category. The cancer incidence for workers with high-exposure potential was similar to that of the Iowa population (standardized incidence ratio, SIR = 1.2), especially for those exposed for 5 or more years and with at least 15 years since exposure began (SIR = 1.0). There were no cases of thyroid, stomach, or nasal cancers. These results indicate that alachlor exposure had no effect on cancer incidence rates in the workers. Thus, there was no indication of increased mortality rates from cancer or any other causes among alachlor manufacturing workers with up to 25 years of follow-up. It is especially noteworthy that cancer rates were not elevated in the highest exposed alachlor workers and that there were no cases of thyroid, stomach, or nasal cancers, which were the oncogenic responses observed in the rat bioassays. The production workers were exposed to alachlor during the year at a level that exceeded that of agricultural exposure; in fact, exposure to manufacturing personnel was estimated to exceed that of pesticide applicators by a factor of 10,000 or more (Acquavella et aI., 1994). Dietary exposure is even lower than
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that of applicators. Therefore, the absence of ocular effects, elevated mortality, and increased cancer incidence in production workers serves as an important indicator of the low potential for adverse effects among the general population, which is exposed to extremely low levels of alachlor.
69.3 ACETOCHLOR 69.3.1 IDENTITY, PROPERTIES, AND USES 69.3.1.1 Chemical Name is 2-chloro-N -ethoxymethyl-N -(2'-ethyl-6'Acetochlor methylphenyl)acetamide. 69.3.1.2 Structure See Fig. 69.1. 69.3.1.3 Syuonyms The common name acetochlor is in general use. The CAS registry number for acetochlor is 34256-82-1. 69.3.1.4 Physical and Chemical Properties Acetochlor has the empirical formula of C14H20N02Cl and a molecular weight of 269.77. It is a light amber to violetcolored oily liquid at room temperature with a specific gravity of 1.110 glml at 30°C. Acetochlor has a boiling point of 162°C at 7 mm Hg and a vapor pressure of 3.4 x 10- 8 at 25°C. Solubility in water is 233 ppm at 25°C. Acetochlor is also soluble in organic solvents including alcohol, acetone, toluene, and carbon tetrachloride. 69.3.1.5 History, Formulations, and Uses This herbicide was registered in the United States in 1994 by the Acetochlor Registration Partnership (ARP), which is now a partnership between Monsanto Company and Dow Agro Sciences. The registration of acetochlor is owned by the ARP, but both companies compete in the marketplace with different formulations. Acetochlor is a selective herbicide that controls a broad spectrum of annual grasses, sedge, and broadleaf weeds primarily in corn. Monsanto manufactures a number of formulations under the trade name Harness®. Formulations produced by Zeneca are sold under the trade name Surpass®. Prior to formation of the ARP, two companies had separately pursued registration with independently generated toxicology databases. Therefore, there are two or more studies for each type of toxicology test performed for regulatory purposes.
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69.3.2 TOXICITY TO LABORATORY ANIMALS 69.3.2.1 Irritation and Sensitization Acetochlor has been shown to be practically nonirritating to the eyes and skin of rabbits. A dermal sensitization study in guinea pigs was positive (Monsanto, 1997b). 69.3.2.2 Acute Studies Thc acute oral and dermal toxicity of acetochlor is low. The oral LDso value is 2148 mg/kg in the rat (Monsanto, 1997b). Dermal LDso values were shown to be greater than 2000 mg/kg in both rats and rabbits. 69.3.2.3 Repeat Dose Studies Several subchronic and chronic toxicology studies have been conducted with acetochlor. The results have been reported by the EPA (1994) and are summarized next. A 21-day dermal study in rats at dose levels ranging from 0.1 to 100 mg/kg/day resulted in mild to minimal skin irritation but no systemic effects. In a second study conducted with rabbits at doses of 100 to 1200 mg/kg/day, the NOEL for systemic toxicity was 400 mg/kg/day. The main effects in two 90-day feeding studies with rats were reductions in body weight and food consumption. The lowest observed effect levels (LOELs) in both studies were 100 mg/kg/day; the NOELs were 10 and 40 mg/kg/day. Oral administration to dogs for 12 months at 4, 12, and 40 mg/kg/day resulted in decreased food consumption, body weight loss, testicular atrophy, and an increase in relative liver weights at the highest dose tested. The NOEL was 12 mg/kg/day. In a second l2-month dog study, the NOEL was 2 mg/kg/day based on changes in serum chemistry values as well as renal and testicular effects in males at 10 mg/kg/day. Dietary administration of acetochlor to CD-l mice for 23 months produced excessive mortality, body weight loss, anemia, interstitial nephritis, and changes indicative of liver damage at the highest dose tested (5000 ppm); this level clearly exceeded the MTD. Increased liver, kidney, and adrenal weights were observed at lower dose levels (500 and 1500 ppm). In a second mouse study conducted at dietary levels of 10, 100, and 1000 ppm for 18 months, the NOEL for systemic toxicity was 10 ppm (1.1 mg/kg/day) in males based on renal changes and 100 ppm (~13 mg/kg/day) in females. The administration of acetochlor to rats at doses of 500-5000 ppm (26-297 mg/kg/day) for 27 months exceeded the MTD at the highest dose tested. This was evidenced by increased mortality, excessive body weight loss, hepatocellular necrosis, and other effects. A NOEL for chronic effects was not established. In two subsequent 24-month studies conducted at lower dose levels, the lowest NOEL was determined to be 7.4 mg/kg/day based on decreased body weight gain and indications of liver toxicity.
69.3.2.4 Pharmacokinetic Studies Rats were found to rapidly metabolize acetochlor to numerous polar metabolites, which were then quickly excreted in the urine and feces; more than 95% of the recovered dose was excreted within 72 h. Blood was the only tissue in which significant retention of radioactivity was observed (2-3% of the dose), and the activity was shown to be associated with hemoglobin. The major metabolite, as well as some minor degradates, were identified and shown to be a result of the mercapturic acid pathway formed by initial glutathione conjugation. The glucuronide conjugate is the major metabolite in bile, and enterohepatic recirculation is known to occur as discussed previously for alachlor. In the mouse, little or no enterohepatic recirculation occurs and the glucuronide is the major urinary metabolite. The dermal penetration of acetochlor was measured in rhesus monkeys following a 24-h exposure. The penetration of the concentrated material was 4.9%, while that of a 1:70 spray dilution was 17.3% (Wester et al., 1996). 69.3.2.5 Genotoxicity Studies Acetochlor gave uniformly negative results in a wide range of in vivo rodent genetic toxicity studies conducted in rats and mice at and below toxic levels. This inactivity is consistent with its lack of in vitro gene mutation and DNA-damaging activity. Several negative in vivo clastogenicity assays indicate that acetochlor's in vitro clastogenicity is not relevant to its activity in vivo or to its ability to produce tumors in rodents (discussed later). This conclusion is further supported by the fact that acetochlor is not genotoxic to the olfactory nasal epithelium of rats, the primary site of its rodent carcinogenicity. Detailed discussions of the genetic toxicity data for acetochlor have been published (Ashby et al., 1996, 1997); a summary of those studies is provided next. An extensive set of studies have led to the conclusion that acetochlor is not mutagenic to Salmonella typhimurium. The experiments included observations made with strains TA98, TA 100, TA1535, TA1537, and TA1538 using S9 mixes derived from rats pretreated with either aroclor or a combination of phenobarbitonelbeta naphthofiavone. Acetochlor was inactive in an in vitro assay for unscheduled DNA synthesis (UDS) using isolated primary rat hepatocytes. In two CHOIHGPRT gene mutation assays conducted at the same concentrations, acetochlor marginally increased the mutation frequency at toxic dose levels and in the presence of S9 mix in one investigation, and produced a clearly negative response in the second assay. Acetochlor is clastogenic to human lymphocytes in vitro at cytotoxic dose levels. The nonclastogenicity of the deschloro analog of acetochlor and the clastogenicity of the desoxy (N -butyl) analog indicate that the chloroacetyl substituent on acetochlor is the clastogenic moiety. Although relatively inert, this substituent can react with sulfhydryl (-SH) groups such as that present on reduced glutathione (Ashby et al., 1996). This reactivity most likely accounts for the clastogenicity of acetochlor in
69.3 Acetochlor isolated lymphocytes that have extremely low levels of protective glutathione. Acetochlor gave a weak positive response in the mouse lymphoma TK± mutation assay when tested in the presence of S9 mix at excessively toxic levels (::s 10% relative survival of cells). It is concluded that acetochlor is not directly mutagenic to DNA, but that it is clastogenic in vitro by virtue of the sulfur reactivity of its chloroacetyl substituent. In vivo, normal levels of glutathione protect against this activity. Several in vivo assays have been conducted and provide important information regarding the relevance of the in vitro assay results. Negative results have been obtained in six separate in vivo assays assessing chromosome aberration activity in somatic and germ cells. Acetochlor was negative in an in vivo rat bone marrow cytogenetic assay, in two mouse bone marrow micronucleus assays, and in two rat and a mouse dominant lethal assays. In the first rat dominant lethal assay (single oral dose), a reversible toxic effect, which was observed only at supraMTD dose levels, led to a reduction in litter sizes at the 3-week sampling period (2000 mg/kg). This effect was not observed in either the mouse or the second rat dominant lethal assay or in multi generation studies conducted using the dietary route up to the MTD of acetochlor. Clearly, the clastogenic activity observed in vitro was not expressed in vivo in the primary rodent cytogenetic assays. Acetochlor gave a negative response in the rat liver unscheduled DNA synthesis (UDS) assay when tested at dose levels up to, and including, the MTD. At a supra-MTD dose level (2000 mg/kg) , a weak positive UDS response was observed. However, this dose depressed hepatic glutathione levels by up to 80% and was associated with severe liver necrosis, substantial release of hepatic enzymes, and lethality among the treated animals (up to 30%). Acetochlor gave negative results in assays for DNA damage (comet assays) conducted using nasal tissue (respiratory and olfactory) derived from rats treated with the supra-MTD dose of 1750 ppm of acetochlor in the diet for either 7 days or 18 weeks. These negative data are particularly relevant, given that the primary target for acetochlor carcinogenesis in the rat is the olfactory nasal epithelium.
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DGXI Labelling Committee did consider all relevant information in 1997, and concluded that acetochlor should not be classified as a carcinogen. The oncogenic potential of acetochlor in rats was assessed in one study at dietary levels of 500, 1500, and 5000 ppm (equivalent to 26, 81, and 297 mg/kg/day) and in a followup study at doses of 40, 200, and 1000 ppm (equivalent to 2, 10.5, and 54 mg/kg/day). The third rat bioassay was conducted at dietary levels of 8, 175, and 1750 ppm (equivalent to 0.8, 7.9, and 80 mg/kg/day). The dose of 54 mg/kg/day represented an MTD based on a 10% depression in body weight gain. A marginal increase in liver tumors was seen but only at a dose (297 mg/kg/day) that greatly exceeded the MTD. Nasal epithelial (olfactory) adenomas and small increases in thyroid follicular tumors occurred at and above the MTD. The NOEL for oncogenic effects was 26 mg/kg/day. Acetochlor's oncogenic potential in mice was evaluated in one study conducted at dietary doses of 500, 1500, and 5000 ppm (equivalent to 85, 254, and 973 mg/kg/day) for 23 months, and in a second study at dietary doses of 10, 100, and 1000 ppm (equivalent to 1.2, 12, and 126 mg/kg/day) for 18 months. In the first study, the dose of 973 mg/kg/day greatly exceeded the MTD as evidenced by decreased body weight gain (males, 70%; females, 24%), increased mortality, and organ toxicity. Under these excessively toxic conditions, the incidence of liver tumors was increased. The only consistent finding in the mouse across the two studies was a marginal increase in the incidence of lung tumors, predominantly adenomas in female mice. These tumors are common in the mouse strain (CD-I) employed in these studies and occur spontaneously at variable incidences. The incidences observed in treated animals from the acetochlor studies were within the historical control range. An apparent increase in the incidence of uterine histiocytic sarcomas was noted in all treatment groups in the first mouse study. However, there was clearly no dose-response relationship across the lO-fold range of doses tested, and there was no increase in the second study. Based on these and other factors, the higher incidence of histiocytic sarcomas was not clearly related to acetochlor administration and is most likely the result of normal variation.
69.3.2.6 Carcinogenicity Studies Acetochlor has been assessed for oncogenic potential in a total of five dietary studies (three rat and two mouse). These studies have been discussed by Ashby et al. (1996) and are summarized next. Although tumors were reported at various sites, the only potentially toxicologically significant effects are nasal and thyroid tumors in the rat. In 1986, acetochlor was classified as a Group B2 carcinogen by the EPA. However, this evaluation was conducted prior to the availability of recent negative in vivo genotoxicity studies and mechanistic information. The negative in vivo genotoxicity results, along with the absence of tumors below the MTD, provide evidence that the oncogenic responses arise through non-genotoxic, threshold-sensitive mechanisms. This conclusion is further supported by the results of the mechanistic investigations (see Section 69.7). The European Union's
69.3.2.7 Development and Reproduction Studies Developmental toxicity has been assessed in two rat and two rabbit studies in which acetochlor was administered by gavage (EPA, 1994). Acetochlor did not produce a teratogenic response in any of these studies. In each of the two rat studies, maternal and fetal toxicity were noted at the highest dose tested (400 and 600 mg/kg/day). The NOEL for both maternal and developmental toxicity was shown to be 150 mg/kg/day in one study and 200 mg/kg/day in the other. In a study with rabbits, acetochlor did not produce developmental toxicity at doses up to 190 mg/kg/day, the highest dose tested; the NOEL for maternal toxicity in this study was 50 mg/kg/day. The second rabbit study was conducted at levels up to 300 mg/kg/day without producing any developmental toxicity. Because the
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NOELs for developmental effects were equal to or greater than the NOELs for maternal effects in all three studies, it was concluded that there is no unique sensitivity from prenatal exposure. The reproductive toxicity of acetochlor has been evaluated in two 2-generation rat reproduction studies (EPA, 1994). The first study resulted in decreased body weight gain and reduced viable litter size at dietary concentrations of 1500 and 5000 ppm. The highest dose of 5000 ppm ("-'400 mg/kg/day) exceeded the MTD. The NOEL for reproductive toxicity was 500 ppm (30.4 and 44.9 mg/kg/day for males and females, respectively). In the second study, systemic toxicity was observed in parental animals at 160 mg/kg/day, which was the highest dose tested. Acetochlor had no effects on reproductive performance, but probably because of a slightly reduced body weight gain in offspring late in lactation at 160 mg/kg/day, the reproductive NOEL was considered to be 21 mg/kg/day.
69.4 BUTACHLOR 69.4.1 IDENTITY, PROPERTIES, AND USES 69.4.1.1 Chemical Name Butachlor is N -(butoxymethy1)-2-chloro-2',6' -diethy lacetanilide. 69.4.1.2 Structure See Fig. 69.1. 69.4.1.3 Synonyms The common name butachlor is in general use. The CAS registry number for butachlor is 23184-66-9.
tization was observed in a study with guinea pigs (Monsanto, 1991). 69.4.2.2 Acute Studies Butachlor is slightly to practically nontoxic in standard animal tests. The oral LDso in rats is 2000 mg/kg, and the dermal LDso is 13,300 mg/kg (Monsanto, 1991). 69.4.2.3 Repeat Dose Studies New Zealand white rabbits were exposed to butachlor dermally for 21 days at dose levels up to 2500 mg/kg/day; the only sign of toxicity was dermal irritation, and the systemic NOEL was the highest dose tested. Feeding studies of 90 days' duration have been performed with the Fischer 344, Sprague-Dawley, and Wistar strains of rats at dietary concentrations ranging from 300 to 40,000 ppm. Toxicity was manifest in one or more strains as decreased survival, body weight depression, anemia, and effects in the liver, kidney, and bladder. The lowest NOEL was observed in the Fischerrat at 300 ppm ("-'18 mg/kg/day). A 90-day feeding study in CD-l mice at dietary concentrations ranging from 1000 to 6000 ppm resulted in liver and kidney toxicity; the NOEL was determined to be less than 1000 ppm due to increased liver weights in male mice at the lowest dose tested. An 8-week oral capsule study was performed in beagle dogs at dose levels ranging from 10 to 100 mg/kg/day. The subchronic NOEL in dogs was 10 mg/kg/day based on indications of liver toxicity (Wilson and Takei, 1999). The chronic toxicity of butachlor has also been evaluated in dogs, mice, and rats. As in the subchronic studies, the liver, kidney, and bladder were the primary target organs in one or more species. The chronic NOELs in dogs and mice were 5 and 8 mg/kg/day, respectively. Chronic NOELs of 4 and 5 mg/kg/day were established in Fischer 344 and SpragueDawley rats, respectively (Wilson and Takei, 1999).
69.4.1.4 Physical and Chemical Properties Butachlor has the chemical formula C17H26N02Cl and a molecular weight of 311.89. It is a liquid at room temperature with a vapor pressure of 1.8 x 10- 6 mm Hg at 25°C. The solubility of butachlor in water is 20 ppm at 25°C. Butachlor is miscible with alcohol, ether, acetone, and benzene. 69.4.1.5 History, Formulations, and Uses Butachlor was developed for the preemergent control of grass and broadleaf weeds in rice and barley. Butachlor is available as emulsifiable concentrate and granular formulations sold under the trade name Machete®. 69.4.2 TOXICITY TO LABORATORY ANIMALS 69.4.2.1 Irritation and Sensitization In studies with rabbits, butachlor was practically nonirritating to the skin and moderately irritating to the eye; dermal sensi-
69.4.2.4 Pharmacokinetic Studies Similar to alachlor and acetochlor, investigations into the metabolism and pharmacokinetics of butach10r have revealed species differences in the way that this molecule is biotransformed and eliminated from the body. Butachlor metabolism in rats is complex due to extensive biliary excretion, intestinal microbial metabolism, and enterohepatic circulation of metabolites. Metabolism in rats follows three major pathways: (1) initial conjugation with glutathione followed by mercapturic acid pathway metabolism; (2) cytochrome P-450-mediated hydroxylation of the aromatic ring, its ethyl groups, and the N -butoxymethylene group; and (3) cleavage of the amide bonds via aryl amidase to form 2,6-diethylaniline, which is further oxidized to 4-amino-3,5-diethylphenol. Approximately 85% of an orally administered dose is eliminated in 48 h; 60% of the excreted material is found in feces and 40% in urine. The results of dermal penetration studies with rhesus monkeys indicate that butachlor is poorly absorbed through the skin. In studies employing a 6-h topical exposure period, only 0.02%
69.5 Metolachlor of the dose was systemically absorbed during exposure to a granular formulation, and 5% of the dose was absorbed when an EC formulation was applied (Wilson and Takei, 1999).
69.4.2.5 Genotoxicity Studies The genotoxic potential of butachlor has been evaluated in numerous assay systems, using a variety of species, metabolic activation conditions, and endpoints. Results from an extensive battery of well-validated tests conducted under GLPs have shown that butachlor is not genotoxic. Butachlor produced no response in the Escherichia coli wp2 reverse mutation assay. A weak positive response was observed in the TA100 strain of Salmonella in one Ames assay; however, this finding was not reproduced in subsequent assays. When tested in cultured Chinese hamster ovary (CHO) cells, there was no mutagenic response in the HGPRT forward gene mutation assay. A CHO in vitro cytogenetics assay was also negative for clastogenicity. A bone marrow cytogenetics assay and a mouse micronucleus assay conducted at ip dose levels up to 750 and 1000 mg/kg, respectively, were both negative. Exposure of CD-l male mice to butachlor for 7 weeks at dietary concentrations up to 5000 ppm produced significant body weight depression but no evidence of dominant lethal effects. An in vivo/in vitro DNA assay in F-344 rats at oral dose levels ranging from 50 to 1000 mg/kg produced no increase in unscheduled DNA synthesis (Wilson and Takei, 1999).
69.4.2.6 Carcinogenicity Studies Butachlor has been evaluated for oncogenic potential in two strains of rats (Sprague-Daw ley and Fischer 344) and in CD-l mice. Butachlor administration induced nasal, stomach, and thyroid tumors only in the Sprague-Dawley rat, but not in the Fischer 344 rat or the CD-l mouse. In the Sprague-Dawley rat, the tumors were similar to those seen with alachlor and occurred only above toxic dose levels (i.e., ::::MTD). Mechanistic studies have been conducted that support the conclusion that these tumors are not relevant in assessing the oncogenic risk to humans. The results of these investigations are presented in Section 69.7. Butachlor was administered to Sprague-Dawley rats in the diet at concentrations of 100, 1000, and 3000 ppm for 26 months. Increases in neoplastic lesions were observed in the olfactory epithelium of the nasal turbinate, glandular stomach mucosa, and thyroid follicular epithelium. The stomach tumors were observed only at the 3000-ppm level. This dose level exceeded the MTD as evidenced by reduced survival and severe body weight depression (> 20%). Increased incidences of nasal and thyroid tumors occurred only at levels of 1000 ppm and above. No treatment-related tumors were noted in Fischer 344 rats fed butachlor at concentrations of 10, 100, and 1000 ppm for 24 months. Butachlor was administered in the diet to mice at concentrations of 50, 500, and 2000 ppm. This study provided no convincing evidence of oncogenic potential (Wilson and Takei, 1999).
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69.4.2.7 Development and Reproduction Studies The potential of butachlor to produce developmental toxicity has been evaluated in both rats and rabbits. The ability of butachlor to impair normal reproduction following continuous oral exposure over two generations was evaluated in rats. The results of these studies showed that butachlor was not a teratogen or reproductive toxin. The rat and rabbit developmental toxicity studies were conducted at dose levels ranging from 49 to 490 mg/kg/day and 50 to 250 mg/kg/day, respectively. In rats, maternal toxicity was observed at the highest dose tested, but there was no effect on the developing fetus. In the rabbit study, a slight increase in postimplantation loss and decreased fetal weights were observed at maternally toxic dose levels (150 and 250 mg/kg/day). The NOEL for maternal and fetal effects was 50 mg/kg/day. Butachlor administration at dietary concentrations of 100-3000 ppm over two successive generations did not adversely affect reproductive performance or pup survival (Wilson and Takei, 1999).
69.5 METOLACHLOR 69.5.1 IDENTITY, PROPERTIES, AND USES 69.5.1.1 Chemical Name Metolachlor is 2-chloro-2'-ethyl-6'-methyl-N -(2-methoxy-lmethylethyl)acetanilide.
69.5.1.2 Structure See Fig. 69.1.
69.5.1.3 Synonyms The common name metolachlor is in general use. A code designation is CGA-24,705. The CAS registry number for metolachlor is 51218-45-2.
69.5.1.4 Physical and Chemical Properties Metolachlor has the chemical formula C15H22N02Cl and a molecular weight of 283.8. It is a liquid that is white to tan in color. Metolachlor has a vapor pressure of 1.3 x 10- 6 and a boiling point of 100°C at 0.001 mm Hg. The solubility of me tolachlor in water is 530 ppm at 20°e. Metolachlor is miscible with most organic solvents.
69.5.1.5 History, Formulations, and Uses Metolachlor was registered with the EPA in 1976. It is a selective herbicide for the control of annual grass weeds, yellow nutsedge, and some broadleaf species. Metolachlor is used in corn, peanuts, and soybeans. S-metolachlor, which contains a higher percentage of the more active of two isomers, was registered in 1997. Its formulations are soid under trade names such as Dual® Magnum®.
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69.5.2 TOXICITY TO LABORATORY ANIMALS
DNA damage/repair assays in rat liver cells and human fibroblasts. An in vivo/in vitro unscheduled DNA synthesis assay also produced no adverse effects.
69.5.2.1 Irritation and Sensitization Metolachlor did not produce eye or skin irritation in rabbits; the material was positive in a dermal sensitization study with guinea pigs (EPA, 1995). 69.5.2.2 Acute Studies Oral (rats) and dermal (rabbits) LD50 values for metolachlor are 2780 and more than 10,000 mg/kg, respectively (EPA, 1995). Inhalation LC50 values (rats) of more than 1.75 mg/l (EPA, 1995) and more than 4.3 mg/l (Ahrens, 1994) have been published. 69.5.2.3 Repeat Dose Studies The results of subchronic and chronic studies have been summarized in the Reregistration Eligibility Decision (RED) document issued by the EPA (1995), and are briefly described next. A 21-day dermal study was conducted with New Zealand white rabbits at dose levels of 10, 100, and 1000 mg/kg/day. Effects observed in high-dose animals were increased bilirubin, liver weights (males), and kidney weights (females); the systemic NOEL was 100 mg/kg/day. A 3-month feeding study conducted in beagle dogs produced no effects at dose levels of 500 and 1000 ppm. In a 6-month feeding study with dogs, the NOEL was 300 ppm (approximately 7.5 mg/kg/day) based on decreased food consumption and body weight gain at 1000 ppm (25 mg/kg/day), the highest dose tested. Beagle dogs were fed metolachlor at dose levels of 100, 300, and 1000 ppm for 1 year. The NOEL for female dogs was 300 ppm (9.7 mg/kg/day) based on decreased body weight gain. In another chronic toxicity study, metolachlor was fed to Sprague-Daw ley rats at dietary levels of 30, 300, and 3000 ppm (approximately 1.5, 15, and 150 mg/kg/day) for 2 years. At the high dose level, decreased body weight gain and increased liver weights (males) were observed. The NOEL for systemic toxicity was 300 ppm. 69.5.2.4 Pharmacokinetic Studies Data reviewed by the EPA (1995) indicate that metolachlor is readily absorbed following oral exposure and excreted in the urine and feces (approximately 50% in each component) over 3 days. Dermal absorption was assumed to be 62.8% in the human based on the results from an in vitro study with rat skin.
69.5.2.6 Carcinogenicity Studies The oncogenic potential of metolachlor has been assessed in two bioassays with CD-l mice and in two studies using Sprague-Dawley rats (EPA, 1995). The EPA's CPRC classified metolachlor as a Group C (possible human) carcinogen. This classification was based on an increased incidence of liver tumors observed in female rats, an effect that was reproduced in the second rat study. The CPRC also recommended that a margin of exposure (MOE) methodology be used for the estimation of human risk (EPA, 1995) rather than a cancer potency factor calculation. Dose levels in the 2-year mouse studies ranged from 30 to 3000 ppm in the diet. No treatment-related carcinogenic effects were noted in these studies. In a feeding study with rats, metolachlor was administered at dietary concentrations of 30, 300, and 3000 ppm. The incidence of benign liver tumors was significantly increased in high-dose females (150 mg/kg/day). This effect was reproduced in a second rat bioassay. Nasal tumors (one adenocarcinoma, one fibrosarcoma) were observed in high-dose males (vs. zero in controls); however, the EPA stated it was not clear that an obvious toxic effect was exerted on the nasal tissue (EPA, 1997). 69.5.2.7 Development and Reproduction Studies Two developmental toxicity studies have been conducted in Sprague-Dawley rats and another was performed using New Zealand white rabbits (EPA, 1995). The dose levels in the two rat studies ranged from 60 to 1000 mg/kg/day. The NOELs for maternal and developmental toxicity were 300 mg/kg/day based on effects at 1000 mg/kg/day. Maternal toxicity was manifested by mortality, convulsions, and reduced body weight gain and food consumption. The effects observed in offspring were reduced mean body weight and an increase in resorptions. Rabbits were evaluated for developmental effects at doses of 36, 120, and 360 mg/kg/day. Maternal toxicity was observed at the highest dose tested, but there was no evidence of developmental toxicity at any dose level. A two-generation rat reproduction study was conducted at doses of 30, 300, and 1000 ppm in the diet (EPA, 1995). The reproductive NOEL was 300 ppm (23.5 and 26.0 mg/kg/day for males and females, respectively) based on reduced pup weights in the Fla and F2a litters. No toxicity was observed in parental animals.
69.5.2.5 Genotoxicity Studies Metolachlor produced negative results in the genotoxicity assays conducted with the material (EPA, 1995). No gene mutations were detected in the Ames/Salmonella assay or the L5178ITK± mouse lymphoma test. No chromosome aberrations were observed in a hamster micronucleus assay and a dominant lethal assay in mice. Metolachlor was negative in
69.6 PROPACHLOR 69.6.1 IDENTITY, PROPERTIES, AND USES 69.6.1.1 Chemical Name Propachlor is 2-chloro- N -isopropylacetanilide.
69.6 Propachlor 69.6.1.2 Structure See Fig. 69.1. 69.6.1.3 Synonyms The common name propachlor is in general use. The major trade name for propachlor products in the United States is Rarnrod®. The CAS registry number for propachlor is 191816-7. 69.6.1.4 Physical and Chemical Properties Propachlor has the empirical formula of CIIHl4ClNO and a molecular weight of 211.7. It is a tan solid with a melting point of noc and has a vapor pressure of 7.9 x lO- s mm Hg at 25°C. The solubility of propachlor in water is 613 ppm at 25°C. Propachlor is soluble in most organic solvents except aliphatic hydrocarbons (Monsanto, 1995; WHO, 1993). 69.6.1.5 History, Formulations, and Uses Propachlor was introduced by Monsanto in 1965. It is a preemergence herbicide for annual grass and broadleaf weed control in corn and sorghum. Propachlor is available as a granular and flowable formulation; it is also available as a prepack formulation with atrazine. 69.6.2 TOXICITY TO LABORATORY ANIMALS 69.6.2.1 Irritation and Sensitization Eye and skin irritation studies conducted in rabbits showed propachlor to be severely irritating to the eye and slightly irritating to the skin. Propachlor produced skin sensitization in guinea pigs (EPA, 1998b). 69.6.2.2 Acute Studies Acute toxicity data have been reported by the WHO (1993). The oral LDso in rats ranges from 550 to 1700 mg/kg, while the dermal LDso is reported to be greater than 20,000 mg/kg in the rabbit. The inhalation (4-h) LCso in rats is greater than 1.2 mg/l, the maximum attainable concentration achieved in the study (EPA, 1998b).
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glands) and centrilobular/midzonal region of the liver (hepatocellular hypertrophy and eosinophilic foci). In the second study conducted in the mouse (EPA, 1998b) at dietary levels up to 6000 ppm, the NOEL was 14.6 and 19.3 mg/kg/day in male and female mice, respectively, based on changes observed in the glandular region of the stomach (erosion and ulceration of the mucosa and herniated mucosal glands) and centrilobular/midzonal region of the liver (hepatocellular hypertrophy and necrosis and eosinophilic foci). Administration of propachlor to beagle dogs for 1 year at dietary levels of 0, 25, 250, and 1000 ppm resulted in reductions in body weight gain and food consumption at the highest concentration (~33 mg/kg/day). These effects may have resulted from poor diet palatability. The NOEL was approximately 6 mg/kg/day (EPA, 1998b). No adverse effects were observed in a 3-week dermal toxicity study of propachlor conducted in rats; the no observed adverse effect level (NOAEL) was 500 mg/kg/day (Rush, 1998). 69.6.2.4 Pharmacokinetic Studies The absorption, distribution, metabolism, and excretion of propachlor has been studied extensively in the rat (WHO, 1993). Propachlor is well absorbed following oral administration; following a single oral dose, approximately 70% is recovered in urine 48-56 hours after administration. The biotransformation of propachlor is complex due to extensive biliary excretion, intestinal microbial metabolism, and enterohepatic recirculation of metabolites. Propachlor is initially metabolized via the mercapturic acid pathway; the molecule is conjugated to glutathione and excreted in the bile along with the catabolites cysteinyl-glycine, cysteine, and N -acetylcysteine-mercapturic acid. The biliary mercapturic acid metabolites undergo deconjugation via intestinal/microbial carbon-sulfur (C-S) lyase activity and can be reabsorbed. The reabsorbed metabolites are subsequently glucuronidated and eliminated in the urine or bile. Glucuronides eliminated in the bile can subsequently undergo further enterohepatic recirculation. Elimination of propachlor is rapid; more than 90% of a single dose is excreted within 48 h, primarily in the urine. Dermal penetration studies of propachlor have not been conducted. However, such data are available for one liquid formulated product; the dermal penetration rates were 20% and 51 % for undiluted and diluted spray formulations, respectively (van de Sandt, 2000). 69.6.2.5 Genotoxicity Studies
69.6.2.3 Repeat Dose Studies Two chronic toxicity studies of propachlor have been conducted in both the rat and the mouse. No treatment-related effects were observed in the first study for either species (WHO, 1993), in which the highest dietary level was 500 ppm, and it was concluded that the dose level selection for these studies was inadequate. In the second study conducted in the rat (EPA, 1998b) at dietary levels up to 2500 (males)/5000 (females) ppm, the NOEL was 300 ppm (~6 mg/kg/day) based on changes observed in the pyloric region of the stomach (erosion, ulceration, and hyperplasia of the mucosa and herniated mucosal
Negative results were observed in in vitro prokaryote (Ames/ Salmonella) and mammalian gene mutation (CHOIHGPRT) assays of propachlor (WHO, 1993). Propachlor had no effect on DNA repair in rat hepatocytes following the conduct of in vitro and in vivo/in vitro UDS assays (EPA, 1998b). A possible weak clastogenic response was observed in an in vitro mammalian (CHO) assay in the presence of metabolic activation; no evidence of clastogenic activity was observed in the absence of a metabolic activation system (EPA, 1998b). In vivo, there was no evidence of a clastogenic response in a rat bone marrow cytogenetics assay (EPA, 1998b). Propachlor had no effects on germ
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cells in a dominant lethal study (EPA, 1998b) conducted in rats at dietary concentrations up to 2500 ppm (111.8 mg/kg/day). Overall, the weight of evidence indicates that propachlor is not genotoxic or clastogenic in mammals. 69.6.2.6 Carcinogenicity Studies Two carcinogenicity studies of propachlor have been conducted in both the rat and the mouse (WHO, 1993). No treatmentrelated carcinogenic effects were observed in the first study for either species in which the highest dietary level was 500 ppm. In the second study conducted in the rat (EPA, 1998b) at dietary levels up to 2500 (males)/5000 (females) ppm, the only evidence of a possible oncogenic effect was a single carcinoma observed in the pyloric (nonglandular) region of the stomach in one male only, at the highest dietary level (2500 ppm). No carcinogenic effects were seen in the female stomach at twice this dietary concentration (5000 ppm). Erosion and proliferation of the pyloric mucosa were observed microscopically at this dietary concentration (2500/5000 ppm); these findings were dose dependent, showed a clear threshold, and are consistent with a non-genotoxic mode of action. In the second study conducted in the mouse (EPA, 1998b) at dietary concentrations up to 6000 ppm, a statistically significant increase in hepatic adenomas was observed in males at the highest dietary level only. Administration of propachlor to male CD-1 mice in their diet for a period of 3 months produced a statistically significant increase in hepatic cell proliferation at 6000 ppm (Hotz and Wilson, 1998); the increased cell proliferation was shown to be dose dependent with a clear threshold at 1000 ppm (NOEL). The increase in hepatic cell proliferation is believed to represent a regenerative response to the underlying severe non-neoplastic changes observed in this organ (see Section 69.6.2.3). These data provide support for a non-genotoxic, threshold-sensitive mechanism (cell proliferation) being responsible for the increase in the predominantly benign hepatic tumors observed in male mice at a dietary concentration of 6000 ppm. 69.6.2.7 Development and Reproduction Studies Two reproduction studies have been conducted in the rat with propachlor; in the first study (EPA, 1998b), parental toxicity was observed at the high dose level only (30 mg/kg/day), by reductions in food consumption and body weight and microscopic changes to the liver (eosinophilia and hypertrophy). No treatment-related reproductive effects were observed in the study and the NOELs for systemic and reproductive toxicity were 3 and 30 (or more) mg/kg/day, respectively. In the second reproduction study (EPA, 1998b), the parental MTD was clearly exceeded at the highest dietary concentration of 2500 (males)/5000 ppm (females); offspring survival and weights were also adversely affected and this dietary level was discontinued after the first litter was weaned. Parental toxicity was observed at a dietary level of 1000 ppm by reductions in weight gain and a microscopic change in the liver (hepatocyte hypertrophy). Slight effects on offspring weight were
observed late in the lactation period at a dietary concentration of 1000 ppm. The NOEL for systemic toxicity to parents and offspring was 100 ppm (~7 mg/kg/day), and the NOEL for reproductive effects was 1000 ppm (~75 mg/kg/day). In a developmental toxicity study in the rabbit (EPA, 1998b), propachlor was administered by gavage at doses of 5.8, 58.3, and 116.7 mg/kg/day on gestation days 7-19. The maternal MTD was exceeded at the highest dose, as indicated by mortality, clinical signs, body weight loss, and reduced food consumption. Slight effects on postimplantation loss, the number of viable fetuses, and fetal weight were observed at the highest dose. None of the effects was statistically significant, and all values were within the relevant laboratory historical control ranges; however, they were considered to be possible effects of treatment. The NOEL for maternal and developmental toxicity was 58.3 mg/kg/day. In a developmental toxicity study in the rat (EPA, 1998b), propachlor was administered by gavage at doses of20, 60, and 200 mg/kg/day on gestation days 6-19. The NOEL was 200 mg/kg/day. Propachlor had no effect on germ cells in male rats at a dietary concentration up to 2500 ppm (111.8 mg/kg/day). Overall, the weight of evidence indicates that propachlor does not produce developmental or reproductive effects.
69.6.3 HUMAN EXPERIENCE Effects reported in humans occupationally exposed to prop achlor have been limited to local skin changes (Von Schubert, 1979). Positive responses have been observed following controlled skin patch testing (Iden and Schroeter, 1977). This is consistent with observations made in experimental animals.
69.7 MODE-OF-ACTION EVALUATIONS: ONCOGENICITY As discussed previously, chronic administration of chloracetanide herbicides has resulted in the production of nasal, stomach, liver, and thyroid tumors in rats. Most tumors occurred at excessively toxic dose levels, at or above the MTD. Such oncogenic responses are of questionable significance because of the extreme doses required to produce them. A weight-of-evidence analysis of mutagenicity results indicates that chloracetanides have no significant genotoxic potential in mammalian systems. This suggests that the oncogenic responses in rats arise through non-genotoxic, threshold-sensitive mechanisms. To better understand the relevance of the rat nasal, thyroid, and stomach tumors to humans, extensive mechanistic investigations were undertaken with alachlor, acetochlor, and butachlor.
69.7 Mode-of-Action Evaluations: Oncogenicity
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Significant species differences in metabolism were demonstrated that provided a mechanistic basis for the rat-specific production of nasal tumors. In rats, alachlor is initially me69.7.1.1 Alachlor tabolized primarily in the liver via the P-450 pathway and Two metabolic pathways have been proposed to explain the de- by glutathione conjugation (Feng and Patanella, 1988, 1989). velopment of chloracetanilide-induced nasal tumors in rats. The The glutathione conjugates and their metabolites undergo enfirst scheme involves the generation of formaldehyde, while the terohepatic circulation with further metabolism in liver and second pathway leads to quinone mine metabolite formation. nasal tissue to form the putative carcinogenic metabolite, a The formaldehyde theory has not been supported by the availdiethyl quinoneimine (DEIQ). Higher rates of intestinal miable data and is only briefly mentioned here. This is followed crobial metabolism, enterohepatic circulation, and target tissue by a discussion of the quinone mine pathway, which is widely metabolism result in a greater conversion of alachlor to DEIQ accepted as being involved in nasal tumor induction. in rat nasal mucosa as compared to other species (Feng et aI., Alachlor and other chloracetanllfde herbicides undergo O-de1990). methylation with the release of formaldehyde (Brown et aI., Quinone imines such as DEIQ are electrophilic, deplete glu1988; Jacobsen et aI., 1991). This led to the suggestion that tathione (GSH), and can exert toxicity by binding to cellular formaldehyde may be involved in the nasal carcinogenicity of proteins. It has been shown that a DEIQ protein adduct is the alachlor and other chtoracetanilide herbicides. However, almajor alachlor-derived protein adduct in rat nasal mucosa (Wilthough formaldehyde is known to produce rat nasal tumors son et aI., 1995a); however, no evidence of this adduct was upon inhalation exposure, the nature and distribution of these observed in the nasal tissue of mice or monkeys (Hey dens et tumors are quite different from those seen with the chloracaI., 1998). The binding of DEIQ to nasal protein is thought to etanilide herbicides. For example, formaldehyde-induced nasal disturb cell structure and function, which leads to cytotoxiclesions are essentially confined to the anterior nose, in regions ity, prolonged regenerative cell proliferation, and the eventual lined by transitional or respiratory epithelium (Morgan et aI., development of nasal tumors. Increased cell proliferation was 1986), whereas chloracetanilide-induced lesions are essentially shown in rats but not mice, and a clear threshold was demonconfined to the posterior region, lined by olfactory epithelium strated; the proliferation was sustained during treatment and (Morgan et aI., 1997). In addition, formaldehyde nasal tumors reversible after dosing was terminated (Heydens et aI., 1998). are characterized by marked irritancy and involve carcinomas of These findings indicate that increased cell turnover is a prereqthe squamous epithelium, the squamous cell metaplasia being a uisite for tumor development and that its induction is threshold consequence of the severe damage produced in the nasal epithesensitive. lium due to the irritant properties of the molecule. In contrast, Critical differences in metabolic capability result in much there is no evidence of irritancy with the chloracetanilides, and higher formation of DEIQ in the nasal mucosa of rats than the tumors are predominately benign adenomas of the olfacother species. For example, the ability of rat nasal tissue to tory epithelium. In addition, as mentioned previously, primary convert the secondary sulfide metabolite of alachlor to 2,6O-deatkylation of alachlor occurs in the liver and there is litDEA-phenol, the proximate metabolite of DEIQ, is more than tle evidence of unchanged alachlor reaching the nasal turbinate 30 times greater than that of monkeys (Li et aI., 1992) and area. Thus, any formaldehyde released during the metabolism 751 times higher than that of human nasal tissue (Wilson et of alachlor in the rat liver would be expected to rapidly unaI., 1995b). Further, species differences are even greater when dergo further metabolism and not be available in the systemic the relative enzymatic rate of alachlor conjugation to GSH circulation to reach the offactory area. Therefore, the avail(the initial metabolic step) is included in the overall rate of able evidence does not support formaldehyde as the metabolite DEA-phenol formation from alachlor; the overall ability of rats responsible for the nasal carcinogenicity observed with chloto convert alachlor to DEA-phenol is 3000- and 22,000-fold racetanilide herbicides. greater than that of humans when the initial GSH conjugation Investigations to understand the mechanism by which chlooccurs in the liver or nose, respectively (Wilson et aI., 1995b). racetanilides produce nasal tumors in rats were initially underThese data indicate that the potential for the formation of the retaken with alachlor. Early whole-body autoradiography (WBA) active DEIQ metabolite in human nasal tissue is negligible. The work showed that alachlor-derived radioactivity specifically loresults further indicate that the rat is not an appropriate model calized in the nasal mucosa of rats but not mice or monkeys. In for assessing alachlor's oncogenic risk to humans. another WBA study, the tertiary amide methylsulfide metabolite of alachlor, a metabolite arising only through enterohepatic cir69.7.1.2 Butachlor culation, was orally administered to rats. Specific localization in nasal mucosa was again observed. The intensity of the labeling The metabolism of butachlor closely parallels that of alachlor. was greater than that observed when parent alachlor was ad- The only difference between these two molecules is the length ministered. These findings showed the importance of metabolic of the N -alkoxymethyl side chain. An important, common step processes in the localization of alachlor metabolites in rat nasal in the metabolism of alachlor and butachlor is P-450-mediated tissue and provided evidence implicating metabolism in the pro- N-dealkylation of this side chain. Once this occurs, the prodduction of nasal tumors. uct of the two parent molecules is identical, and the subsequent 69.7.1 RAT NASAL TUMORS
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metabolism to DEIQ and toxicologic response would be the same. Indeed, it has been shown that oral administration of butachlor results in the same rat-specific nasal localization and induction of cell proliferation observed with alachlor. Similar differences in nasal metabolism across species as observed with alachlor are also apparent for butachlor; with the potential for DEIQ formation substantially higher in the rat. 69.7.1.3 Acetochlor The metabolism of acetochlor has also been extensively studied. These studies have confirmed that the overall metabolism of acetochlor shares critical commonality with alachlor and butachlor. In rats, metabolism in the liver and gastrointestinal tract produces sulfur-containing metabolites that are delivered to the nose and undergo further metabolism. WBA studies corresponding to those described previously for alachlor have been conducted for acetochlor. Acetochlor-derived radioactivity localized in the nasal turbinates of rats but not in mice. Analysis of protein adducts found in nasal tissue from rats treated with acetochlor or its methylsulfide metabolite confirmed that the major adducts involved EMIQ, the acetochlor quinone imine metabolite analogous to DEIQ from alachlor. As with alachlor, these adducts were not found in the nasal tissues of mice or monkeys. Similarly, when acetochlor was fed to rats at doses that produced nasal tumors in the chronic studies, prolonged cell proliferation was observed. Acetochlor administration did not induce cell proliferation at the same dose level in mice. Acetochlor administration to rats at 200 ppm, a nononcogenic dose level, did not cause increased cell proliferation. In vitro metabolism studies similar to those done with alachlor have measured the enzymatic activities involved in the conversion of acetochlor and key metabolites to EMIQ. The data showed that the potential for EMIQ formation is significantly lower in the mouse and monkey than the rat. Overall, the relative rates for conversion of acetochlor to quinone imines in rat, mouse, and monkey were very similar to those seen with alachlor, suggesting that the differences between rat and human nasal tissue for acetochlor would also be comparable to that seen with alachlor. 69.7.2 RAT STOMACH TUMORS Butachlor and alachlor are close structural analogs that produce the same stomach tumors in rats. An extensive mechanistic research program was undertaken to understand the mechanism by which these tumors are induced. A stomach tumor initiation-promotion study demonstrated that butachlor was not active as an initiator, but it did promote the formation of tumors after treatment with N-methyl-N'-nitro-Nnitrosoguanidine (MNNG), a known initiator (Branch et aI., 1995). A subsequent tumor promotion study with alachlor showed that it, too, produced stomach tumors by the same promotional activity. The results of these studies provided direct
experimental evidence indicating that the stomach tumors are produced by a non-genotoxic mode of action. Mechanistic studies with butachlor have shown that chloracetanilide-induced gastric neoplasia involves toxicity (atrophy) to the fundic mucosa as an initial event following high dose exposure (Hard et aI., 1995; Thake et aI., 1995). This atrophy then results in compensatory cell proliferation in the fundic mucosa. The accompanying profound loss of parietal cells leads to an extensive gastric hypochlorhydria and a subsequent increase in pH of the gastric contents. This increase causes excessive gastrin production, resulting in a substantial elevation of serum gastrin levels. The tropic effect of long-term stimulation of enterochromaffin-like (ECL) cells and fundic stem cells by gastrin further drives a sustained regenerative cell proliferation response that ultimately results in the induction of the gastric neoplasms observed in the chronic studies. Additional work demonstrated that high-dose alachlor exposure also induces the same mucosal atrophy, hypochlorhydria, and hypergastrinemia that characterize the unique oncogenic mechanism demonstrated with butachlor. Mucosal atrophy did not occur at a lower, nononcogenic dose of alachlor. Prolonged exposure to toxic doses of alachlor and butachlor were required to produce stomach tumors in rats. Such exposure would not occur in humans; results from a study with rhesus monkeys showed that mucosal atrophy, the initial preneoplastic event, did not occur at doses ranging from 100 to 400 mg/kg/day. These doses are comparable to, and higher than, the oncogenic dose (150 mg/kg/day) in rats. Based on all the data, it was concluded that these chloracetanilide-induced stomach tumors are not relevant to humans (Heydens et aI., 1998). 69.7.3 RAT THYROID TUMORS Various studies have shown that the prolonged alteration of thyroid homeostasis can lead to the development of thyroid follicular tumors in rats (Hill et aI., 1989; Thomas and Williams, 1991; Zbinden, 1989). The oncogenic response is mediated via increased levels of circulating thyroid stimulating hormone (TSH), which result in hyperplasia and, ultimately, neoplasia. One mechanism causing such a thyroid imbalance involves the induction of liver enzymes (McClain, 1989). This induction increases the rate of thyroid hormone excretion and is responsible for the compensatory elevation in TSH observed. Separate studies were conducted to determine if alachlor, acetochlor, and butachlor produce thyroid tumors by this mechanism (Ashby et al., 1996; Wilson and Takei, 1999; Wilson et aI., 1996). The results of these studies showed that the administration of each chloracetanilide at the dose level producing thyroid neoplasia in the chronic studies caused significant increases in liver weight and T4-UDPGT (thyroxine-uridine diphosphate glucuronosyl transferase) activity, serum TSH levels, and thyroid weight at several time points. These changes were observed as early as 7-14 days after dosing began and continued throughout 2 or more months of dosing. Reversibility studies with alachlor and butachlor showed that serum TSH
References
and hepatic UDPGT activity returned to normal after dosing was discontinued. Rats are especially sensitive to altered thyroid function by this mechanism because of their susceptibility to liver enzyme induction and lack of thyroid binding globulin in plasma, making thyroid hormone more susceptible to metabolic activity. Furthermore, because this mechanism is believed to be a threshold-sensitive phenomenon, it is not expected to be relevant for humans under actual exposure scenarios with chloracetanilide herbicides.
69.8 COMMON MECHANISM OF TOXICITY The Food Quality Protection Act of 1996 requires the EPA to perform a combined risk assessment for chemicals that produce adverse effects by a common mechanism of toxicity. The extensive database of mechanistic information, developed (see Section 69.7) to support (1) a nongenotoxic threshold mechanism of action and (2) lack of relevance to humans for the nasal turbinate, stomach, and/or thyroid oncogenic effects produced in rats for alachlor, acetochlor, and butachlor, provides the basis for considering whether these chemicals can be grouped based on a common mechanism of toxicity. The data support grouping alachlor, acetochlor, and butachlor with respect to a common mechanism of toxicity for nasal turbinate and thyroid tumors, and alachlor and butachlor for stomach tumors.
REFERENCES Acquavella, J. F., Ireland, B., Leet, T., Anne, M., Farrell, T., and Martens, M. (1994). Epidemiological studies of morbidity and mortality among alachlor manufacturing workers. In "Proceedings of the XII Joint CIGR, IAAMRH, IUFRO International Symposium: Health, Safety and Ergonomic Aspects in Use of Chemicals in Agriculture and Forestry;' pp. 184-194. Acquavella, J. E, Riordan, S. G., Anne, M., Lynch, C. F., Collins, J. J., Ireland, B. K, and Heydens, W. F. (1996). Evaluation of mortality and cancer incidence among alachlor manufacturing workers. Environ. Health Perspect. 104,728-733. Ahrens, W. H. (1994). "Herbicide Handbook of the Weed Science Society of America," 7th ed. Weed Sci. Soc. Am., Champaign, IL. Ashby, J., Kier, L., Wilson, A. G. E., Green, T., Lefevre, P. A., Tinwell, H., Willis, G. A., Heydens, W. F., and Clapp, M. J. L. (1996). Evaluation of the potential carcinogenicity and genetic toxicity to humans of the herbicide acetochlor. Hum. Exp. Toxicol. 15,702-735. Ashby, J., Tinwell, H., Lefevre, P. A., Williams, J., Kier, L., Adler, I.-D., and Clapp, M. J. L. (1997). Evaluation of the mutagenicity of ace tochIor to male rat germ cells. Mutat. Res. 393, 263-281. Bonfanti, M., Taverna, P., Chiappetta, L., Villa, P., D'Incalci, M., Bagnati, R., and Fanelli, R. (1992). DNA damage induced by alachlor after in vitro activation by rat hepatocytes. Toxicology 72, 207-219. Branch, D. K., Shibata, M., Thake, D. C., and Wilson, A. G. E. (1995). "Gastric Tumor InitiationlPromotion Study of Butachlor in Sprague-Dawley Rats." Presented at Annual Conference of International Federation of Societies of Toxicologic Pathologists, Tours, France. Brown, M. A., Kimmel, E. c., and Casida, J. E. (1988). DNA adduct formation by alachlor metabolites. Life Sci. 43, 2087-2094. Erexson, G. L., Bryant, M. F., Doer, C. L., Kwanyuen, P., and KIigerman, A. D. (1993). Cytogenetic analyses of human peripheral blood Iymphocytes exposed to alachlor in vitro. Environ. Mo!. Mutagen. 21, 19.
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Feng, P. C. C., and Patanella, J. E. (1988). Identification of mercapturic acid pathway metabolites of alachlor formed by liver and kidney homogenates of rats, mice, and monkeys. Pestic. Biochem. Physiol. 31, 84-90. Feng, P. C. c., and Patanella, J. E. (1989). In vitro oxidation of alachlor by liver microsomal enzymes from rats, mice, and monkeys. Pestic. Biochem. Physiol. 33, 16-25. Feng, P., Wilson, A., McClanahan, R., Patanella, J., and Wratten, S. (1990). Metabolism of alachlor by rat and mouse liver and nasal turbinate tissues. Drug Metab. Dispos. 18,373-377. Georgian, L., Moraru, I., Draghicescu, T., Dinu, I., and Ghizelea, G. (1983). Cytogenetic effects of alachlor and mancozeb. Mutat. Res. 116,341-348. Gold, L. S., Sawyer, C. B., Magaw, R., Backman, G. M., deVediana, M., Levinson, R., Hooper, N. K, Havender, W. R., Bernstein, L., Peto, R., Pike, M. c., and Ames, B. N. (1984). A carcinogenic potency data base of the standardized results from animal bioassays. Environ. Health Perspect. 58,9-319. Hard, G. c., Iatropoulos, M. J., Thake, D. C., Wheeler, D., Tatematsu, M., Hagiwara, A., Williams, G. M., and Wi1son, A. G. E. (1995). Identity and pathogenesis of stomach tumors in Sprague-Dawley rats associated with the dietary administration ofbutachlor. Exp. Toxicol. Pathol. 47, 95-105. Heydens, W. F. (1998). Summary of toxicology studies with alachlor. J. Pestic. Sci. 24, 75-82. Heydens, W. F., Wilson, A. G. E., Kier, L. D., Lau, H., Thake, D. c., and Martens, M. A. (1998). An evaluation of the carcinogenic potential of the herbicide alachlor to humans. Hum. Exp. Toxico!. 18,363-391. Hill, R. N., Erdreich, L. S., Paynter, O. E., Roberts, P. A., Rosenthal, S. L., and Wilkinson, C. F. (1989). Thyroid follicular cell carcinogenesis. Fundam. App!. Toxicol. 12, 629-698. Hotz, K J., and Wilson, A. G. E. (1998). "Effect of Propachlor on Cell Proliferation in the Liver of Male Mice." Monsanto Unpublished Report. Iden, D. L., and Schroeter, A. L. (1977). Allergic contact dermatitis to herbicides. Arch. Dermatol. 113, 983. Ireland, B., Acquavella, J., Farrell, T., Anne, M., and Fuhremann, T. (1994). Evaluation of ocular health among alachlor manufacturing workers. J. Occup. Med. 36,738-742. Jacobsen, N. E., Sanders, M., Toia, R. F., and Casida, J. E. (1991). Alachlor and its analogues as metabolic progenitors of formaldehyde: Fate of N-methoxymethyl and other N-alkoxylalkyl substltuents. J. Agric. Food Chem. 39, 1342-1350. Kier, L. D., Heydens, W. F., Lau, H., Thake, D. C., and Wilson, A. G. E. (1996). Genotoxicity studies of alachlor. Toxicologist 30, 231. Kronenberg, J. M., Fuhremann, T. W., and Johnson, D. E. (1988). Percutaneous absorption and excretion of alachlor in rhesus monkeys. Fundam. App!. Toxico!. 10,664-671. Leet, T., Acquavella, J., Lynch, c., Anne, M., Weiss, N., Vaughan, T., and Checkoway, H. (1996). Cancer incidence among alachlor manufacturing workers. Am. 1. Ind. Med. 30,300-306. Li, A. A., Asbury, K J., Hopkins, W. E., Feng, P. c., and Wilson, A. G. E. (1992). Metabolism of alachlor by rat and monkey liver and nasal turbinate tissue. Drug Metab. Dispos. 20,616-618. Lin, M. F., Wu, C. L., and Wang, T. C. (1987). Pesticide clastogenicity in Chinese hamster ovary cells. Mutat. Res. 88, 241-250. Lutz, W. K. (1986). Quantitative evaluation of DNA binding data for risk estimation and for classification of direct and indirect carcinogens. J. Cancer Res. Clin. Onco!. 112,85-91. McC1ain, R. (1989). The significance of hepatic microsomal enzyme induction and altered thyroid function in rats: Implications for thyroid gland neoplasia. Toxicol. Patho!. 17,294-306. Meisner, L. F., Belluck, D. A., and Roloff, B. D. (1992). Cytogenetic effects of alachlor and/or atrazine in vivo and in vitro. Environ. Mo!. Mutagen. 19, 77-82. Millburn, P. (1975). Excretion of xenobiotic compounds in bile. In "The Hepatobiology System: Fundamental and Pathological Mechanisms" (W. Taylor, ed.), p. 109. Plenum, New York. Monsanto Company (1991). "Material Safety Data Sheet: Butachlor Technical." Monsanto Company, SI. Louis, MO. Monsanto Company (1995). "Material Safety Data Sheet: Propachlor Technical." Monsanto Company, SI. Louis, MO.
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CHAPTER 69
Chloracetanilides
Monsanto Company (1997a). "Material Safety Data Sheet: Alachlor Technical." Monsanto Company, St. Louis, MO. Monsanto Company (1997b). "Material Safety Data Sheet: Acetochlor Technical." Monsanto Company, St. Louis, MO. Morgan, K. T., Gross, E. A., Joyner, D. R., Ishmael, J., and Thake, D. (1997). Proliferative nasal lesions induced in rats by alachlor, acetochlor, and butachlor originate in specific regions of the olfactory mucosa. Toxicologist 36,12. Morgan, K. T., Jiang, X. Z., Starr, T. B., and Kerns. W. D. (1986). More precise localization of nasal tumors associated with chronic exposure of F-344 rats to formaldehyde gas. Toxieo!. App!. Pharmaeo!' 82, 264-271. Rush, R. E. (1998). "Propachlor: A 21-Day Dermal Toxicity Study in Rats." Monsanto Unpublished Report. Taningher, M., Terranove, M. P., Airoldi, L., Chiappetta, L., and Parodi, S. (1993). Lack of alachlor induced DNA damage as assayed in rodent liver by the alkaline elution test. Toxicology 85, 117-122. Thake, D. c., Iatropoulos, M. J., Hard, G. c., Hotz, K. J., Wang, C.-X., Williams, G. M., and Wilson, A. G. E. (1995). A study of the mechanism of butachlor-associated gastric neoplasms in Sprague-Dawley rats. Exp. Toxico!. Patho!. 47, 107-116. Thomas, G., and Williams, E. (1991). Evidence for and possible mechanisms of non-genotoxic carcinogenesis in the rodent thyroid. Mutat. Res. 248, 357370. U.S. Environmental Protection Agency (EPA) (1994). Decision Document, Conditional Registration of the New Chemical Acetochlor. Office of Pesticide Programs, Washington, DC. U.S. Environmental Protection Agency (EPA) (1995). Metolachlor Reregistration Eligibility Decision Document, Office of Pesticide Programs, Washington, DC. U.S. Environmental Protection Agency (EPA) (1996). Proposed guidelines for carcinogen risk assessment: Notice. Fed. Reg. 17960-18011. U.S. Environmental Protection Agency (EPA) (1997). "Common Mechanism of Toxicity." Presentation to the Scientific Advisory Panel, Washington, DC. U.S. Environmental Protection Agency (EPA) (1998a). Alachlor Reregistration Eligibility Decision Document, Office of Pesticide Programs, Washington, DC.
U.S. Environmental Protection Agency (EPA) (1998b). Propachlor Reregistration Eligibility Decision Document, Office of Pesticide Programs, Washington, DC. U.S. Environmental Protection Agency (EPA) (2001). The Grouping of a Series of Chloroacetanilide Pesticides Based on a Common Mechanism of Toxicity, Office of Pesticide Programs, Washington, DC. van de Sandt, J. J. M. (2000). "In Vitro Percutaneous Absorption Study with Propachlor in Ramrod® SC through Viable Human Skin Membranes." Monsanto Unpublished Report. Von Schubert, H. (1979). Allergic contact dermatitis caused by propachlor. Dermato!' Monatssehr. 165, 495-498 (In German). Wester, R. C., Melendres, J., and Maibach, H. I. (1992). In vivo percutaneous absorption and skin decontamination of alachlor in rhesus monkey. 1. Toxico!. Environ. Health 36, 1-12. Wester, R. c., Melendres, J. L., and Maibach, H. I. (1996). In vivo percutaneous absorption of acetochlor in the rhesus monkey: Dose-response and exposure risk assessment. Food Chem. Toxieo!. 34, 979-983. Williams, R. (1971). Species variations in drug biotransformations. In "Fundamentals of Drug Metabolism and Drug Disposition" (B. Lau, H. Mandel, and E. Way, eds.), p. 187. Williams & Wilkins, Baltimore. Wilson, A. G. E., and Takei, A. S. (1999). Summary of toxicology studies with butachlor. 1. Pestie. Sei. 25, 75-83. Wilson, A. G. E., Lau, H., Asbury, K. J., Thake, D. c., and Heydens, W. F. (l995a). "Mechanistic Basis for the Rat Specific Nasal Tumors Observed with Alachlor." Abstract, International Congress of Toxicology-VII, Seattie. Wilson, A. G. E., Lau, H., Asbury, K. J., and Heydens, W. F. (I 995b). Metabolism of alachlor by human nasal tissue. Fundam. App!. Toxieo!. 15, 1398. Wilson, A. G. E., Thake, D. c., Heydens, W. F., Brewster, D. w., and Hotz, K. J. (1996). Mode of action of thyroid tumor formation in the male Long-Evans rat administered high doses of alachlor. Fundam. App!. Toxieo!. 33, 16-23. World Health Organization (WHO) (1993). "Propachlor." Environmental Health Criteria Document 147 (prepared by L. Ivanova-Chemishanka). Zbinden, G. (1989). Hyperplastic and neoplastic responses of the thyroid gland in toxicological studies. Arch. Toxieol. 12, 98-106.
CHAPTER
70 Paraquat Edward A. Lock Syngenta Central Toxicology Laboratory
Martin F. Wilks Syngenta Crop Protection AG
70.1 IDENTITY, PROPERTIES, AND USE
mixtures with urea herbicides include Dexuron®, Gramocil®, Gramonol®, Gramuron®, and Tota-Col®.
70.1.1 CHEMICAL NAME Paraquat is 1,I'-dimethyl-4,4'-bipyridinium ion (IUPAC, CAS RN [4685-14-7}), also known as the l,l'-dimethyl-4,4'-bipyridyldiylium ion.
70.1.4 PHYSICAL AND CHEMICAL PROPERTIES The molecular formula of the cation is C12H14N2 with a molecular weight of 186.3. The dichloride salt has the formula C12H14Cl2N2 and a molecular weight of 257.2. Paraquat dichloride forms colorless, hygroscopic crystals which decompose at 300°C. It is practically nonvolatile with a vapor pressure of <0.1 mPa. It is very soluble in water (700 g/l at 20°C) and practically insoluble in most other organic solvents. It is stable in neutral and acidic media but readily hydrolyzed in alkaline media. Paraquat is photochemically decomposed by ultraviolet radiation in aqueous solution.
70.1.2 STRUCTURE
CHJ-Q-G"'+-CH J [20-J Paraquat dichloridc
Figure 70.1
70.1.5 HISTORY, FORMULATIONS, AND USES 70.1.3 SYNONYMS The common name paraquat is in general use (BSI, E-ISO, ANSI, WSSA, JMAF), except in Germany. Paraquat is usually formulated as the dichloride salt (also known as methyl viologen) (CAS MR [1910-42-5]). The bis(methyl sulphate) salt (CAS NR [2074-50-2]) is no longer commercialized. Code designations for the material are PP148 (for the dichloride salt) and PP910 (for the bis(methyl sulphate) salt). Trade names for paraquat dichloride formulations include Crisquat®, Cyclone®, Dextrone X®, Esgram®, Efoxon®, Goldquat® 276, Gramoxone®, Herbaxon®, Katalon®, Osaquat Super®, Pilarxone®, R-Bix®, Speeder®, Starfire®, Sweep®, Total®, and Weedless®. Mixtures of paraquat with diquat are sold under trade names including Actor®, Dukatalon®, Opal®, Pathclear® (also includes simazine and aminotriazole), Preeglox®, Preglone®, Seccatuto®, Spray Seed®, and Weedol®. Trade names of Handbook of Pesticide Toxicology Volume 2. Agents
Paraquat was first described in 1882 by Weidel and Russo. In 1933, Michaelis and Hill discovered its redox properties and called the compound methyl viologen. The herbicidal properties of paraquat were first described by Brian et at. (1958) and it became commercially available in 1962. Paraquat is mainly formulated as an aqueous solution with surface-active agents. In some countries, a low-strength granular formulation (also containing diquat) is available. Paraquat is a fast-acting, nonselective contact herbicide, absorbed by the foliage with some translocation in the xylem. It is used for broadspectrum control of broad-leaved weeds and grasses in fruit orchards and plantations, and for inter-row weed control in many crops. It is also used for general weed control on noncrop land, as a defoliant on cotton and hops, for destruction of potato haulms, as a desiccant, and for control of aquatic weeds. Paraquat is rapidly deactivated upon contact with the soil and does not leach.
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Copyright © 200 1 by Academic Press All rights of reproctw;;tion in any form reserved.
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CHAPTER 70
Paraquat
70.2 TOXICITY TO LABORATORY ANIMALS
exposure signs of toxicity are few but may include respiratory effects.
70.2.1 SIGNS OF TOXICITY
70.2.2 ACUTE TOXICITY
Following a lethal dose of paraquat to rats mortality is first seen on days 2-5 after dosing but deaths can also occur around days 10-12 (Clark et aI., 1966; Sharp et aI., 1972; Smith and Rose, 1977), indicating there is considerable interindividual animal response to the chemical. The major cause of death after a median lethal dose is due to lung damage. The animals develop acute pulmonary edema with signs of labored respiration and ultimately die of respiratory failure (Clark et aI., 1966; Kimbrough and Gaines, 1970; Murray and Gibson, 1972; Sharp et aI., 1972). Rabbits, however, do not show signs of respiratory distress. They stop eating and drinking and tend to die without overt toxicity, following oral dosing (Butler and K1einerman, 1971; Clark et aI., 1966). Rats and mice given doses above the maximum lethal dose (MLD), by intraperitoneal (ip) or subcutaneous (sc) administration show signs of hyperexcitability, ataxia, and convulsions and usually die within a few hours of dosing, indicative of an effect on the central nervous system (Bagetta et aI., 1992; Clark et aI., 1966). Following chronic
The acute oral toxicity of paraquat to the rat is shown in Table 70.1. The ask for of pure paraquat dichloride expressed as the cation was about 150 mg/kg to female rats and ranged from 100 to 143 mg/kg in a number of different strains from a number of different laboratories. No sex difference in toxicity was seen and the toxicity was similar for the two different salts of paraquat (Table 70.1). Fasting rats prior to oral administration of paraquat made little difference to the toxicity. The 7 day MLD with 95% confidence limits were 143 (123-166), 130 (106-159), and 126 (102-156) mg paraquation/kg, respectively, for rats fasted for 0, 4, and 8 hr (Murray and Gibson, 1971). Mice are less sensitive than the rats to orally administered paraquat, while guinea pigs, cats, monkeys, and rabbits are more susceptible (Table 70.2). Paraquat was more toxic when given by the ip or intravenous (iv) routes with a MLD of approximately 20 mg paraquat ion/kg (Table 70.3), indicating that following oral dosing the compound is poorly absorbed from the gastrointestinal tract (see
Table 70.1 Acute Toxicity of Paraquat to the Rat (Data Expressed as mg Paraquat ion/kg) Route of
Median lethal dose
dichloride
Sex
Strain
administration
(time studied)
Reference
Pure salts
F
Nsa
po
112 (104-122),
Clark et ai., 1966
Paraquat
ISO (139-162), 141 b (140-142) (14 days) Pure salt
F
NS
po
ISO (110-173)
Mehani, 1972
(21 days) Formulation
M
Sprague-
po
Dawley Formulation
M
Sherman
143 (123-166)
Murray and Gibson, 1971
(7 days) po
100 (87-117)
Kimbrough and Gaines, 1970
(15 days) Formulation
F
Sherman
po
llO (90-134)
Kimbrough and Gaines, 1970
(15 days) Formulation
M
Sprague-
po
Formulation
M
Wistar
ll5 (90-150)
Sharp et ai., 1972
(30 days)
Dawley po
95 (79-ll4)
Sharp et ai., 1972
(30 days) Pure salt
F
NS
ip
19 (16-21)
Clark et ai., 1966
16b (14-19)
Pure salt
F
NS
ip
16 (10-26)
Mehani, 1972
(21 days) Formulation
M
SpragueDawley
aNot stated.
bDimethosulphate salt.
iv
21 (19-25)
Sharp et ai., 1972
70.2 Toxicity to Laboratory Animals
1561
Table 70.2 Acute Toxicity of Paraquat to Laboratory Animals (Data Expressed as mg Paraquat Ion/kg) Route of
Median lethal
Species
Sex
administration
dose
Reference
Mouse
Fa
ip
30c 30c (26.5-35.1)
Bus et ai., 1976a
F
Ecker et ai., 1975b;
Mouse
F
po
196c
Bus et aI., 1976b
Guinea pig
F
ip
3
Clark et aI., 1966 Clark et aI., 1966
Guinea pig
Mb
po
30 (22-41)
Guinea pig
M&F
po
22c (15-33)
Murray and Gibson, 1972
Cat
F
po
CIark et aI., 1966
Monkey (Macaca
M&F
po
35 (27-46) 50c
Rabbit
M
po
100C
Kuo and Nanikawa, 1990
Rabbit
M
po
50 (45-58)
Mehani, 1972
Rabbit
M
ip
25 (15-30)
Mehani, 1972
Rabbit
M
ip
McElligott, 1972
Dog
M
sc
18 (11-31) 1.8c (1-6.1)
Dog
F
sc
3.5" (2.4-10.1)
Nagata et ai., 1992a
Murray and Gibson, 1972
jascicu[aris)
Nagata et aI., 1992a
a Female.
bMale. cReference refers to paraquat, not clear if salt or ion.
later). The guinea pig and dog (Nagata et aI., 1992a) are also more sensitive to systemic administration with a MLD of 23 mg/kg (Table 70.2), reflecting poor or incomplete absorption of paraquat from the gastrointestinal tract after oral administration. Rabbits given a single iv dose of paraquat at 40 or 80 mg/kg died within 24 h; while they survived a single dose of 10 mg/kg iv, no lung lesions were seen at these doses (Ilett et aI., 1974). The vehicle used to administer paraquat can influence lethality in mice. For example, paraquat was more toxic when given by the ip or sc route in water than in isotonic saline, suggesting that the solvent may influence the absorption from the site of injection and hence the amount delivered to the lung (Drew and Gram, 1979). The dermal toxicity of paraquat has been studied in rabbits (Table 70.3). The precise technique of application of paraquat to the skin, whether the site of application is open to the air or covered and whether the rabbits are prevented from grooming, affects the findings (Clark et aI., 1966; McElligott, 1972). Rabbits fitted with restraining collars to reduce grooming the site of ap-
plication, followed by decontamination of the skin and removal of the collars, showed glossitis, anorexia, weakness, and loss of weight with some skin erythema followed by hyperkeratosis and desquamation at the higher doses, indicating that some oral ingestion had still occurred. This technique resulted in a MLD following a single application of 236 mg paraquat ionlkg. If, however, the restraining collars were not removed, then the erythema and desquamation was mild and the extent of glossitis and hence body weight loss was less. Under these conditions the MLD was found to be >480 mg ionlkg, the maximum dose possible to apply in a satisfactory manner (McElligott, 1972). Thus, when compared to the systemic MLD of 18 mg ionlkg (Table 70.2). It indicates that little of the applied dose has been absorbed through intact skin. Dermal exposure of rats to paraquat gave an MLD of 80-90 mg paraquat ionlkg (Table 70.3). However these authors (Kimbrough and Gaines, 1970) gave no information on the state of the skin after application, whether the site was occluded or free for the rats to groom. The absorption of paraquat across the skin has been reviewed by Smith (1988),
Table 70.3 Acute Toxicity of Paraquat to Laboratory Animals Following Dermal Application or Inhalation Exposure (Data Expressed as mg Paraquat Ion/kg)
Species
Sex
Route of
Median lethal
administration
dose
Reference
Rat
M
Dermal
80 (60-96)
Kimbrough and Gaines, 1970
Rat
F
Dermal
90 (74-110)
Kimbrough and Gaines, 1970
Rabbit
M
Dermal
236 (collars removed)
Clark et aI., 1966;
>480 (with collars)
McElligott, 1972
6 J.l-g/llh
Gage, 1968a
Rat
M&F
Inhalation
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CHAPTER 70
Paraquat
who concluded that paraquat is poorly absorbed by intact skin and raised technical concerns about the validity of the earlier dermal studies reported by Kimbrough and Gaines (1970). Paraquat is not volatile, but following inhalation exposure to an aerosol is irritant to the respiratory tract. At lethal concentrations under these conditions, death is usually delayed for several days and is due to respiratory failure. Following single exposures the MLD is a function of both the amount and duration of exposure, which in the rat is approximately 6 !-Lg/l/hr (Table 70.3). Guinea pigs and male mice are of similar sensitivity to the rat, while female rats and rabbits are less sensitive. Dogs can tolerate a concentration-time product of 25 !-Lg/l/hr without ill effects (Gage, 1968a). The toxicity is also a function of particle size, and 3 !-Lm was the most lethal to the rat. Large particles do not reach the alveolar region and are less toxic. Under normal conditions of manufacture, handling, and use, inhalation exposure is not considered to be a hazard. Studies with rabbits have shown that the lung is susceptible to paraquat injury following intrabronchial deposition (Zavala and Rhodes, 1978) and inhalation exposure (Seidenfeld et aI., 1978), although as mentioned above it is refractory following oral or intraperitoneal administration. Local instillation of paraquat in the lungs of rats will also produce local injury and fibrosis (Kimbrough and Gaines, 1970; Wyatt et aI., 1981). 70.2.3 IRRITATION AND SENSITIZATION
Paraquat is a skin and eye irritant but is not a skin sensitizer (Bainova, 1969). As discussed earlier skin irritation has been reported in rabbits when the area of application is occluded (Clark et aI., 1966; McElligott, 1972), resulting in local erythema followed by hyperkeratosis and desquamation. Instillation of a 0.29% aqueous solution of paraquat into the rabbit eye produced no effect. However, more concentrated solutions produced inflammation of the conjunctiva and nictitating membrane. This response developed gradually over 12 h and lasted for 48-96 h (Clark et aI., 1966). Instillation of higher concentrations of paraquat (3-48 mg contained in 0.2 ml of water) into the rabbit eye produced a dose-related increase in ocular injury with doses of 48 mg (about 16 mg/kg) and above producing fatalities (Sinow and Wei, 1973). These findings indicate that absorption of paraquat from the eye is similar to that following systemic administration. 70.2.4 SUBCHRONIC TOXICITY
Dally administration of paraquat in the diet of rats at an inclusion rate of 100 ppm (about 5 mg ion/kg/day) was tolerated for several months. However, if increased to 250 ppm (about 12.5 mg ion/kg/day) the rats became ill and died within 27 to 57 days. Females appeared to be more susceptible than males, the primary target organ for toxicity being the lungs (Clark et aI., 1966). A number of other studies have shown that moderate daily doses of paraquat can be tolerated. Rats fed 125 ppm (about 6.25 mg/kg/day) for 2 years showed no toxic effects
and dogs tolerated 50 ppm (about 0.9 mg/kg/day) for 2 years (Howe and Wright, 1965). Rats given paraquat in their drinking water at 1.3 or 2.6 mg/l for 2 years showed some mortality and histological changes in the lung at the highest dose, while only minimal changes to the lung were seen at the lower dose (Bainova and Vulcheva, 1977). The MLD for paraquat fed in the diet for 90 days has been determined. Groups of female rats were fed 300, 400, 500, 600, and 700 ppm paraquat and their food consumption was recorded at intervals to enable the dose in mg/kg/day to be calculated. After 90 days the surviving rats were held for 2 weeks to allow time for any delayed deaths. The MLD was 21 mg/kg/day, giving a subchronic toxicity factor of 5.2 (ratio of acute to subchronic MLD's), indicating that paraquat has a moderate cumulative toxicity in this species (Kimbrough and Gaines, 1970). Rabbits given paraquat ip at 10 or 20 mg/kg at 48 h intervals showed marked signs of toxicity with a high mortality following three to five doses. There was little evidence of lung damage and it is likely that the animals died from multiorgan failure (Butler and Kleinerman, 1971). Rabbits can, however, tolerate 3 mg/kg day ip for up to 14 days, but when increased to 6 mg/kg/day significant mortality was seen (Hassan et aI., 1989). Daily oral dosing at 11 mg/kg/day to male rabbits for 30 days produced few signs of toxicity, with only one animal showing lung damage (Dikshith et aI., 1979). Subchronic exposure following dermal application has been examined in rabbits. The mortality observed with repeated daily applications beneath an occlusive dressing gave a MLD of 6.24 (4.6-8.5) mg ion/kg/day of paraquat over 20 days (McElligott, 1972). At the higher doses the skin was reddened and sloughing with local edema, while at the lowest dose some scab formation was seen after about 7 days of application. Systemic effects at postmortem included renal tubular necrosis, focal hepatocellular necrosis, and pulmonary congestion. Studies were also conducted where the skin was not occluded; rabbits were fitted with collars and these were either removed after decontamination of the skin or at the end of the observation period. The MLD for 20 days exposure was between 7.25 and 14.5 mg paraquat ion/kg/day for the animals where the collars were removed after decontamination and at least 24 mg ion/kg/day for those where the collars were left on all the time. The rabbits showed marked signs of salivation, which was associated with glossitis and ulceration of the tongue. The animals refused to eat and death occurred in a state of cachexia; this effect was less marked at the lower doses when the collars were kept in place all the time (McElligott, 1972). When the "Gramoxone" formulation of paraquat was diluted to spray strength and applied to the skin of rabbits for 20 days (2.4 mg ion/kg/day) no clinical signs of toxicity or pathological changes were seen (McElligott, 1972). Daily subcutaneous dosing of paraquat to dogs for 4 weeks resulted in some animals being terminated at the top dose of 0.495 mg/kg/day, while at the other doses of 0.165 and 0.055 mg/kg/day all animals appeared well (Nagata et aI., 1992b). Histopathology of the lungs showed proliferation of alveolar lining cells and some fibrosis at the top dose and pul-
70.2 Toxicity to Laboratory Animals monary changes (thickening of the alveolar wall and pleura) at all doses. The 28 day MLD from this study was about 0.5 mg/kg/day. Repeated exposure of rats to a respirable aerosol of paraquat (approx. 90% < 2.5 J.lm) by inhalation at 0.4 J.lg/l for 6 hr per day for 15 days, over 3 weeks, led to intermittent respiratory problems after about four exposures. At postmortem after 15 exposures the animals showed marginal paraquat-related pathology to the lungs. Exposure to 0.1 J.lg/l for 15 daily, 6 hr periods showed no signs of toxicity or pathology in the lungs (Gage, 1968a). Rats exposed to 0.003 J.lg/l for 6 hr a day for 5 days per week for 2 months put on body weight, remained in good condition, and showed no histopathological evidence of lung damage. Bainova et al. (1972) exposed rats to a respirable paraquat aerosol at 1.1 or 0.05 mg/L for 6 hr/day for 4.5 months and found evidence of lung damage at the higher dose, with little effect at the lower dose. Seidenfeld et al. (1978) exposed rabbits by inhalation to paraquat by an ultrasonic nebulizer (mean particle size 4 J.lm) at a concentration of 0.1 mg/ml of aqueous solution for 2 hr/day for 5 days per week for 3 months and found no lung damage. However rabbits exposed to 2 mg/ml of aqueous solution for 2 hr/day could only tolerate three exposures and developed a reduced arterial oxygen tension and specific compliance which was associated with marked lung injury. Overall these studies show that following acute and chronic exposure the primary target organ for toxicity is the lung, with deaths from lung damage frequently taking many days to occur following a single dose. The rabbit is unusual in that it does not readily develop a lung lesion following oral or parenteral exposure, but if instilled into the lung or exposed via inhalation, lung injury ensues. Renal functional impairment with some renal tubular necrosis is the other major organ affected. Dose levels of paraquat that do not cause lung damage in laboratory animals following acute and chronic exposure have been clearly established. 70.2.5 MUTAGENIC AND CARCINOGENIC POTENTIAL
Paraquat is not carcinogenic in either rats or mice. The activity seen in some short-term assays for mutagenesis is associated with cytotoxicity and is believed to arise as a consequence of the redox cycling ability of paraquat, leading to superoxide anion formation. Paraquat has minimal to no genotoxic activity when evaluated in a wide range of in vitro and in vivo test systems. Many groups have reported the absence of an effect while others have reported weakly positive effects (Dabney, 1995; IPCS, 1984; Ribas et al., 1995 and references therein). These later effects were usually associated with high cytotoxicity or mortality and are believed to arise as a consequence of the redox cycling ability of paraquat. It is known that DNA damage frequently occurs when cells are exposed to oxidative stress (Brawn and Fridovich, 1981; Repine et aI., 1981).
1563
Paraquat-mediated effects on DNA have been reported in bacteria (Moody and Hassan, 1982; Yonei et aI., 1986), Chinese hamster cells (Nicotera et aI., 1985; Sofuni et aI., 1985; Tanaka and Amano, 1989), isolated alveolar macrophages and epithelial type 11 cells (Dusinska et aI., 1998), and in a few cases cells from treated mice (He and Yasumoto, 1994; Rios et aI., 1995). These responses are all considered to be secondary to superoxide anion generation. Studies with cultured mammalian cells have shown that paraquat inhibits DNA synthesis leading to the arrest of the cells in S-phase (Tomita, 1996; Yamagami et al., 1994). This effect occurs prior to the onset of cytotoxicity and is thought to be part of a cascade of events initiated by the production of oxygen free radicals by the redox cycling of paraquat. These findings have been extended to rat lung cells exposed to paraquat in vivo which also showed S-phase arrest at early times after dosing. Prior treatment of the rats with a diet enriched in sodium tungstate, an inhibitor of xanthine oxidase to reduce the production of free radicals, prevented the S-phase arrest produced by paraquat (Matsubara et aI., 1996) and reduced mortality (Kitazawa et aI., 1991). Once inside a cell, paraquat can redox cycle, producing oxygen free radicals that can cause cell cycle arrest and inhibit DNA synthesis. These findings are consistent with early studies showing that paraquat reduces DNA synthesis at early times after dosing (Smith and Rose, 1977; Van Osten and Gibson, 1975). Paraquat has been evaluated for its carcinogenic potential in both rats and mice and it was concluded that at all doses up to the maximum tolerated dose, paraquat did not result in a compound related increase in tumour incidence (Bainova and Vu1cheva, 1977; FAOIWHO, 1986). 70.2.6 EFFECTS ON REPRODUCTION, EMBRYOTOXICITY AND TERATOGENICITY
Paraquat has no effect on fertility, is not teratogenic, and only produces fetotoxicity at doses that are maternally toxic. The main finding in multigeneration studies was lung damage. Paraquat does not readily cross the placenta and enter the embryo of mice when given either orally or by ip administration (Bus et aI., 1975). In contrast, paraquat appears to readily cross the placenta of rats, being detected in fetuses within 30 min of an iv injection to 20 day pregnant rats (Ingebrigtsen et aI., 1984). A three-generation reproduction study in rats maintained on dietary levels of paraquat of 30 or 100 ppm showed no effect on food intake, fertility, fecundity, neonatal morbidity, mortality. No teratogenesis or other changes in gross or histological morphology were seen, except for a slight increase in the incidence of renal hydropic degeneration in the 3-4 week old young receiving 100 ppm (about 10 mg/kg/day). Pregnant and young animals did not appear to be more susceptible than adults (FAOIWHO, 1973). A two-generation reproduction study in mice maintained on dietary levels of 45,90, or 125 ppm showed no effects on age to parturition, number born, or abnormalities
1564
CHAPTER 70
Paraquat
in the pups in the first generation following 45 or 90 ppm. However, at 125 ppm an increase in mortality was seen in the dams and pups during the first few weeks oflife (Dial and Dial, 1987). The second generation mice were more resistant to the effects of paraquat, the only effect being an increase in the age of the mothers at second parturition on the highest dose of paraquat (Dial and Dial, 1987). Subsequent studies to explore the basis for the high mortality in the first generation dams, and pups exposed to 125 ppm paraquat in the diet showed that they almost certainly died from lung damage. This only occurred in pups exposed prenatally via the placenta, not in pups exposed postnatally (Dial and Dial, 1989). Bus and Gibson (1975) also reported that paraquat given to mice in their drinking water at either 50 or 100 ppm from day 8 of gestation and to the young until 42 days of age increased pup mortality at 100 ppm but not 50 ppm. The lungs of mice killed 42 days after 100 ppm snowed extensive alveolar consolidation and collapse, supporting the view that the deaths at this dose were probably due to lung damage. No dominant lethal effects were seen in mice exposed to paraquat at oral doses up to 4 mg/kg/day for 5 days (Anderson et aI., 1976). High doses of paraquat injected ip into pregnant rats or mice on various days of gestation can produce significant maternal toxicity (Bus et aI., 1975; Khera et aI., 1970). Examination of the fetuses of mice exposed to 1.67 or 3.35 mg/kg ip or 20 mg/kg per os po daily on days 8-16 of gestation induced no teratogenic effects, although a slight increase in nonossification of the sternbrae was seen (Bus et aI., 1975).
and in some areas hemorrhage into the air spaces occurred. At this time there was extensive infiltration of inflammatory cells into the alveolar interstitium, air spaces, and perivascular areas, although the alveolar endothelial capillaries were mainly spared. The animals died as a consequence of severe anoxia usually within the first few days after dosing and this has been confirmed by others (Clark et aI., 1966; Sharp et aI., 1972; Smith and Rose, 1977). This phase has been called the destructive phase (Smith and Heath, 1976). Similar early pathological changes have been reported by Kimbrough and Gaines (1970), Brooks (1971), Modee et aI. (1972), Wasan and McElligott (1972), Smith et aI. (1974), Sykes et al. (1977), and Smith and Heath (1976). Some rats that survive for up to 10-12 days after dosing develop an extensive hypercellular lesion in the lung which is dominated by proliferation of fibroblasts. This phase of the lesion is called the proliferative phase and is characterized by attempts by the epithelium to regenerate and restore normal architecture of the alveolar epithelium (Kimbrough and Gaines, 1970; Smith and Heath, 1974a; Vijeyaratnam and Corrin, 1971). The findings in these animals are typically extensive intraalveolar and interalveolar fibrosis, which in association with residual edema reduces gaseous exchange results in death from anoxia. It appears that the initial damage to the alveolar epithelium, produced by paraquat, is the primary event in the development of the lung injury, with the proliferative fibrosis being a consequence of the extensive damage produced. For a more detailed review on pulmonary injury see Smith and Heath (1976).
70.2.7 PATHOLOGY OF THE LUNG
70.2.8 ABSORPTION
The toxic effects of paraquat were first described by Clark et aI. (1966) who reported that the histological effects of paraquat in rats, mice, and dogs are similar. The lung, liver, kidney, and thymus were affected, the lung being the major target. The effect of paraquat in the cynomologus monkey is similar to that in rats (Murray and Gibson, 1972). In contrast, as mentioned previously, rabbits do not develop lung lesions following acute oral or ip administration (Butler and Kleinerman, 1971; lIett et aI., 1974; Mehani, 1972; Zavala and Rhodes, 1978). There is one report of daily administration of paraquat in the drinking water to rabbits over several days leading to lung damage that resembles that seen in rats (Restuccia et aI., 1974). Inhalation exposure to paraquat produces lung damage in the rabbit (Seidenfeld et aI., 1978). The hamster responds in a similar way, being refractory to a single sc dose of paraquat, but lung fibrosis is produced by repeated sc injections (Butler, 1975). The most extensive studies on the pathogenesis of lung damage produced by paraquat have been conducted in rats. The time course of development of the injury in rats given a single MLD ip was reported by Vijeyaratnam and Corrin (1971) and Smith and Heath (1974a). Damage to the type I and 11 alveolar epithelial cells was seen within a day of dosing. This damage was more marked by days 2-4 with large areas of the alveolar epithelium being completely lost. Alveolar edema developed
The first studies on the absorption and excretion of paraquat from the gastrointestinal tract were conducted by Daniel and Gage (1966) in rats. Following a single oral dose of 4, 6, or 50 mg/kg 4 C-methyl] paraquat dichloride, most of the radioactivity was excreted within 48 h. Occasionally some appeared in the feces 3 and 4 days after dosing at the higher doses, with small amounts also in the urine. Between 6 and 14% of the dose was excreted in the urine over 48 h when given as the dichloride salt, and 16-23% when given as the dimethylthiosulphate salt, the remainder being in the feces. In contrast, when paraquat as either salt was given sc, the bulk of the radioactivity appeared in the urine within 24 h of dosing, showing that paraquat is poorly absorbed across the gastrointestinal tract of the rat. Subsequent studies have extended and essentially confirmed these findings (Chui et aI., 1988; Lock and Ishmael, 1979; Molnar and Hayes, 1971; Murray and Gibson, 1974). The concentration of paraquat in the plasma following an oral dose to the rat is determined largely by the amount of paraquat present in the small intestine (Smith et aI., 1974). Studies in the dog using tracer doses (129 ).Lg/kg) of 4 Cmethyl]paraquat support this. Peak plasma concentrations following oral dosing were observed at 75-90 min (Fig. 70.2), with about 46-66% of the dose absorbed, as judged by the amount excreted in the urine at 6 h (Davies et aI., 1977). Thus,
e
e
70.2 Toxicity to Laboratory Animals 1.6
i-:-~ ~Ratl
~ 1.2
.:. -:;; :::J
0.8~ [
..
~
0.4
.. ~
a::
o ~-+--+---I---I--+-o 2 3 4 5
0
6
TIIll! (h)
Figure 70.2 Plasma levels of paraquat in the rat and dog following a single nontoxic oral dose. The dog was given a total dose of 1.03 mg of paraquat, while the rats were dosed at 0.038 mglkg. Data adapted from (Davies et aI., 1977; Chui et aI., 1988).
1565
Following oral administration of paraquat to rats, the peak plasma concentration is seen between 30 and 60 min (Figs. 70.2 and 70.3) following either a tracer dose (Chui et aI., 1988) or a toxic dose (Murray and Gibson, 1974). This profile is similar to that seen in the dog (Fig. 70.2) (Davies et aI., 1977). The peak plasma concentration in the monkey and guinea pig occurs within the first hour (Fig. 70.2) and 30 min respectively following a toxic oral dose (Murray and Gibson, 1974). Overall, these studies indicate that paraquat is rapidly but incompletely absorbed from the gastrointestinal tract of laboratory animals and humans (see later), with peak plasma concentrations occurring within 30-90 min. Paraquat is poorly absorbed across human skin in vitro, human skin being less permeable to paraquat than the skin of rats, rabbits, or guinea pigs (Walker et aI., 1983). Application of a low dose of 4 C] paraquat (150 nmol/kg) in acetone to rat skin resulted in a peak blood level about 1 hr after dosing and a total of 3.5% of the dose absorbed (Chui et aI., 1988). It should be pointed out that an occlusive dressing was applied in these studies which has previously been shown to greatly enhance the percutaneous absorption of paraquat in animals (McElligott, 1972). Overall, these studies, plus those of Hoffer et al. (1989) on rabbits, indicate that paraquat is poorly absorbed across the intact skin of laboratory animals.
e
the dog absorbs a greater percentage of an orally administered dose of paraquat than the rat, which is consistent with the greater susceptibility of the dog to paraquat by this route of administration. Pretreatment of dogs with a drug that will block gastric emptying delayed the peak plasma concentration by 3 to 6 h, indicating that the stomach is not the major site of absorption (Bennett et aI., 1976). These data in both rats and dogs indicate that the absorption of paraquat from the gastrointestinal tract occurs somewhere beyond the stomach. It is assumed this is similar for humans but there is limited evidence to support this. Based on the cationic nature of paraquat, it would not be expected to readily cross cellular membranes, and it seems unlikely that simple diffusion would explain the rapid but incomplete absorption seen in the rat and dog. Studies in vitro with isolated mucosa from a number of different regions of the rat gastrointestinal tract (Steffen and Konder, 1979) have confirmed that the jejunum and ileum have the greatest capacity to transport paraquat from the lumen into the bloodstream and also showed that a component of the transport is facilitated (Heylings, 1991).
70.2.9 DISTRIBUTION
In the rat, after a lethal oral dose, the plasma paraquat concentration remained relatively constant after the initial peak for up to 32 h (Murray and Gibson, 1974; Rose et aI., 1976a). During this time the concentration in the lung rose progressively to several times that found in the plasma. In no other organ, apart from the kidney, the major organ for the excretion of paraquat, was a time-dependent accumulation of paraquat detected (Murray and Gibson, 1974; Rose et aI., 1976a). These findings, plus the earlier observation of Sharp et al. (1972) who
5
I-+-Rat
4
-.-Monkey
I
~
.:!.
-:;;
3
:::J
tT
I!
.. ..
~
2
E III
a::
0 0
6
12
18
Time (h)
Figure 70.3 Plasma levels of paraquat in the rat and monkey following a single toxic oral dose. The rats were given 126 mglkg paraquat while the monkeys received 50 mglkg. Data adapted from (Murray and Gibson, 1974).
1566
CHAPTER 70
Paraquat
administered paraquat iv and showed that paraquat was retained in the lung with a half-life of 50 h, provided the key evidence showing that those organs that had the highest concentration of paraquat were those that were susceptible to injury, namely the lung and kidney. Many other groups have subsequently examined the pharmacokinetics and elimination of paraquat in the rat (Chui et al., 1988; Dey et al., 1990; Maling et al., 1978), dog (Giri et al., 1982; Hawksworth et al., 1981; Pond et al., 1993), rabbit (Ilett et al., 1974; Yonemitsu, 1986; Yu et al., 1994), and mouse (Drew and Gram, 1979). The distribution of paraquat in the body is best described by a three-compartment model, with input to and removal from the central plasma compartment. Simulations of plasma concentrations in the peripheral compartments show there is a compartment with rapid uptake and removal of paraquat, which was assumed to be the highly vascular tissues such as the kidney, and a slow uptake compartment reaching a maximum about 4-5 h after iv dosing, which may be the lung (Hawksworth et al., 1981). Using lung slices, Rose et al. (1974a) first described the time-dependent accumulation of paraquat into lung tissue. This process was shown to be energy-dependent in that it could be inhibited by the addition of the metabolic inhibitors cyanide
plus iodoacetate to the incubation medium. The accumulation of paraquat into rat lung was shown to obey saturation kinetics with an apparent Km of 70 !1M and a Vmax of 300 nmol/h/g wet weight of lung slice (Table 70.4) (Rose et aI., 1974a). Other aspects of the accumulation of paraquat into the lung will be discussed in more detail later. Hawksworth et al. (1981) also showed that early onset of renal failure markedly affected the concentration of paraquat in the peripheral compartments, suggesting that any reduction in renal excretion of paraquat may allow more of the chemical to be transported into the lung. The distribution in the rabbit, which is refractory to lung damage following a single systemic dose, showed the organs with the highest concentration of paraquat were the lung and kidney at 6 and 24 h after dosing, but the concentration in rabbit lung appeared to decline more rapidly than from rat lung (lIett et aI., 1974). Whole body autoradiography studies have provided valuable information on the tissue distribution of paraquat; early studies by Litchfield et al. (1973) in mice given iv 4 C methyl]paraquat showed retention in the lung. A more detailed study using [3H-methyl]-paraquat and thin tissue sections revealed localization of radioactivity at all time intervals after dosing in the
e
Table 70.4
Kinetic Constants for the Accumulation of Paraquat and Putrescine by Rat and Human Lung Slices or Isolated Alveolar Type 11 Cells
Species/ tissue Rat-lung slice
Human-lung slice
Paraquat accumulation
Putrescine accumulation
Km (j.!M)
Vrnax a
70 210 119
300 710 636
40
300
244
370
Cultured rat-type
Km
(j.lM)
8 7 31 12-18 13.5 13.1
Vrnax b
480 330 870 720 723
Reference Rose et a!., 1974a Ross and Krieger, 1981 Karl and Friedman, 1983 Smith and Wyatt, 1981 Nemery et a!., 1987 O'Sullivan et a!., 1991 Hardwick et a!., 1990 Smith et aI., 1982 Rose et aI., 1974a
7 2-11 7
376 99-249 414
Hoet et aI., 1994 Brooke-Taylor et a!., 1983 Hoffer et a!., 1993
5
18
Lewis, 1989
8-14
58
Richards et aI., 1987 Van der Wal et aI., 1990 Oreffo et aI., 1991
11 cells
29 64 Suspensions of rat -type 11 cells
88
15 29
Cultured humantype 11 cells a Vrnax b Vrnax
in lung slices expressed as nmol/h/g wet weight of slice. Alveolar type 11 cells expressed as pmol/h/j.lM DNA.
128
2.5
34
Chen et a!., 1992
6-8
12-14
Hoet et aI., 1994
70.2 Toxicity to Laboratory Animals lung, choroid plexus, muscle, and melanin in addition to excretory pathways such as the proximal tubules of the kidney, urine, liver, gall bladder, and intestinal contents of the mouse (Waddell and Marlowe, 1980). Radioactivity in the lungs appeared to be higher in certain areas and higher cellular resolution autoradiography revealed that the radioactivity was confined to alveolar type 11 cells, which are one of the major target cells for paraquat toxicity. In these studies it was essential to keep the tissue frozen at all times to prevent diffusion of paraquat which is highly polar. An association of paraquat with melanin has been demonstrated and this is probably due to an ionic interaction (Larsson et aI., 1977, 1978; Lindquist et aI., 1988). Immunohistochemical approaches utilizing specific antibodies to paraquat have shown immunoreactive material localized primarily in bronchiolar epithelial cells and walls of blood vessels in the lungs of rats, 3 h to 10 days after an iv dose. Other studies have localized immunoreactive material in the intestine, liver, kidneys, and brain to capillary walls and glial cells but not neurones, after paraquat administration (Nagao et aI., 1990,1991, 1993).
70.2.10 METABOLISM Paraquat is very poorly metabolized with the bulk of the administered dose being excreted unchanged in the urine and faeces. Daniel and Gage (1966) compared the colorimetric assay for paraquat with that found by radiochemical detection on the urine and feces of rats dosed with paraquat and demonstrated that there was very close agreement. Chromatography of the urine and lung tissue from rats treated with paraquat also showed no evidence of biotransformation (Hughes et aI., 1973; Murray and Gibson, 1974; Rose et aI., 1974a). No radioactivity was excreted in expired air following paraquat administration to rats, indicating that it did not undergo metabolism to C02 (Murray and Gibson, 1974). Incubation of paraquat with rat caecal contents for up to 24 h showed up to a 50% loss, indicating microbial metabolism. The loss was not seen when the contents of the caecum were heat treated (Daniel and Gage, 1966). However, in vivo studies in rats, guinea pigs, and dogs showed little
evidence of biotransformation, indicating that the in vitro studies had overpredicted the likely metabolism (Summers, 1980). The overriding weight of evidence is that metabolism does not contribute to the toxicity of paraquat.
70.2.11 EXCRETION Elimination of paraquat from the body is almost exclusively via the kidneys. The renal clearance of paraquat is greater than that of creatinine in the rat (Chan et aI., 1997; Lock, 1979), dog (Hawksworth et aI., 1981), sheep (Webb, 1983), monkey (Purser and Rose, 1979), and humans (Bismuth et aI., 1982); see later for a more detailed discussion on humans. Thus paraquat is actively secreted by the kidney. Renal tubular secretion was completely inhibited by N' -methylnicotinamide, suggesting that paraquat is secreted via a cationic transport system (Hawksworth et aI., 1981). The transport mechanisms for organic cations in renal proximal tubular cells is not fully understood. Recently two membrane proteins, organic cation transporter 1 (Grundemann et aI., 1994) and organic cation transporter 2 (Okuda et aI., 1996), have been isolated from rat kidney. The organic cation transporter 1 located on the basolateral membrane will transport tetraethylammonium, and this can be inhibited by other organic cations such as quinine. The organic cation transporter 2, which is predominantly expressed in the kidney, stimulates the uptake of tetraethylammonium and this can be markedly inhibited by cimetidine. Studies using freshly isolated renal proximal tubules and renal cell lines have shown that paraquat is transported across the basolateral membrane (from the bloodstream into the renal tubular epithelial cell) using an organic cation transport system (Chan et aI., 1996a, b, 1997, 1998; Groves et aI., 1995). The transport of paraquat can be blocked by the addition of the divalent cation quinine, cimetidine, and to a lesser extent tetraethylammonium (Chan et aI., 1996b), suggesting that paraquat may be transported by both transport systems (Fig. 70.4). Exit across the apical membrane into the tubular lumen is also an active process; current evidence suggests that there are
+2
PQ
PQ+2 pH 7.4
1567
pH 7.2
Basolateral
Apical
membrane
membrane
Figure 70.4 Mechanism of paraquat transport across renal tubular cells. A schematic representation of the proposed transport systems for paraquat across renal tubular cells. The transporters are OCT 1 at the basolateral membrane and P-glycoprotein and the cation/H+ exchange system at the brush border membrane. Adapted from (Chan et al., 1998). Reproduced with permission from © 1998.
1568
CHAPTER 70
Paraquat
two cation transport systems, an electroneutral organic cation IH+ exchange (Sokol et aI., 1988) and P-glycoprotein (Dutt et aI., 1992). Studies with rabbit brush-border membrane vesicles have shown that paraquat is a substrate for the cationIH+ exchange transporter and further that it can inhibit the transport of other monovalent cations such as tetraethylammonium (Wright and Wunz, 1995). In the rat in vivo, the fractional excretion of paraquat decreased from 2.1 at a plasma concentration of about 0.4 nmol/ml to 1.2 at a plasma concentration of 21 nmol/ml, demonstrating that the excretion of paraquat is greater than the glomerular filtration rate and that the process is saturable (Chan et aI., 1997). Thus, at low plasma concentrations paraquat will be readily cleared from the body; however, at higher plasma concentrations the system will become saturated and less paraquat will be cleared. At toxic doses it is well established that paraquat can cause renal functional impairment. In rats, given 126 mg ion/kg po (Lock, 1979), and mice given 50 mg ion/kg iv (Ecker et aI., 1975b) renal impairment was observed 17-24 h after dosing. In the cynomologus monkey given 85 mg ion/kg po the decline in renal clearance was seen 12 h after dosing, the first time examined (Purser and Rose, 1979). In dogs given 20 mg ion/kg iv (Hawksworth et aI., 1981) renal impairment was observed as early as 2.5 h after dosing. An early report on the renal handling of paraquat by the dog suggested that paraquat was reabsorbed by the proximal tubules. This study was conducted at high plasma concentrations (54-810 nmol/ml) where the transport system will have been saturated and function impairment almost certainly will have occurred (Ferguson, 1973). The weight of evidence strongly supports the view that paraquat is actively secreted by the kidney of laboratory animals and humans (see later). The implication of impairment of renal excretion is that more paraquat is available in the plasma to accumulate into the lung. Whole body autoradiography has shown that paraquat was present in the gall bladder of mice, indicating some biliary excretion (Waddell and Marlowe, 1980). The extent of biliary excretion of paraquat was <5% when dosed to bile cannulated rats, rabbits, or guinea pigs and measured over a 3 h period (Hughes et aI., 1973). The bulk ofthe dose appeared unchanged in the urine. These authors suggest that the molecular weight of paraquat at 186 was below the minimal molecular weight of about 500 for chemicals that are excreted in bile. Radioactivity from paraquat was also detected in the bile of dogs given a single iv dose, indicating some biliary excretion in this species (Giri et aI., 1982).
(1976a). The apparent kinetic constants for the uptake process were very similar for all species examined except the rabbit. Slices of rabbit had a very high affinity, but low capacity, to accumulate paraquat which is consistent with the in vivo findings that show that following oral or parenteral administration of paraquat the rabbit does not develop a lung lesion. For the rat the derived Km was 70 J.LM with a Vrnax of 300 nmol/h/g wet weight of lung (Table 70.4). The kinetic constants for rat and human lung were very similar, suggesting that the rat lung was a good surrogate for studying paraquat uptake into human lung (Rose et al., 1976a). The kinetics of accumulation of paraquat into human lung slices has been confirmed by others, the Vrnax being similar at 370 nmol/h/g wet weight while the Km was lower at 244 J.LM (Hoet et aI., 1994). Considerable interindividual variation is seen in paraquat accumulation into human lung slices (Brooke-Taylor et aI., 1983) which may either reflect individual variability or more likely the state of the tissue and delay between removal of the tissue and analysis of paraquat transport. Table 70.4 summarizes the available data on the transport kinetics for paraquat in rat and human lung tissue. These observations, coupled with the finding that paraquat is not metabolized by the lung nor covalently bound to any degree (Forman et aI., 1982; Ilett et aI., 1974; Sullivan and Montgomery, 1983), suggests that this accumulation is mediated through binding to and subsequent translocation into lung cells by a carrier-mediated system. The finding that paraquat was actively transported into lung slices lead to a search for chemicals that might inhibit this process (Dunbar et aI., 1988; Lock et aI., 1976; Maling et aI., 1978; Ross and Krieger, 1981; Smith et aI., 1981) and hence provide protection against paraquat-induced lung toxicity. A number of chemicals were identified that could block paraquat uptake into lung slices but none of these were effective in the whole animal (see later under treatment of poisoning). Studies were also undertaken to try and identify the endogenous chemicals for this transport system. A wide range of chemicals was examined and a number of naturally occurring amines were identified as the most effective inhibitors of paraquat accumulation into slices of rat lung, and which themselves act as substrates. These amines include the diamine putrescine, the oligoamines spermidine and spermine (Gordonsmith et aI., 1983; Smith and Wyatt, 1981; Smith et aI., 1982), and the disulphide cystamine (Lewis et aI., 1989). The physiological role for this transport system is not known, but it has been suggested that polyamines, which are known to regulate cell growth, may play a role in the differentiation of alveolar type 11 cells to type I cells (Smith, 1982). It has also been proposed that cystamine represents a source of taurine, which may 70.2.12 ACCUMULATION OF PARAQUAT have an antioxidant role in the lung (Lewis et aI., 1989; Wright INTO THE LUNG et aI., 1986). Cystamine has also been implicated in playing a role in regulating cellular NADPH levels in response to oxThe original discovery of an energy-dependent accumulation of idative stress (Brigelius, 1985). The structural requirements of paraquat into rat lung tissue (Rose et aI., 1974a) lead to studies substrates for this system have been examined and at least two to look for this transport system in the lung of other species, charged nitrogen atoms separated by a distance of at least four including human. The accumulation of paraquat by slices of methylene groups (about 6.6 0 A) is essential for uptake (Gorlung from a number of species was reported by Rose et al. donsmith et aI., 1983; O'Sullivan et aI., 1991; Ross and Krieger,
70.2 Toxicity to Laboratory Animals 1981). It is probable that paraquat, which meets these criteria, is recognized as a substrate and thereby accumulated (Smith, 1987). Paraquat accumulation into rat lung slices is reduced in the presence of putrescine in a dose-related manner (Karl and Friedman, 1983; Smith and Wyatt, 1981). Subsequent studies showed that putrescine was accumulated into slices of rat lung by saturable energy-dependent process with an apparent Km of 7 J.!M and a Vmax of 330 nmol/h/g wet weight of lung. The Km is about lO-fold lower than that for paraquat, indicating that the endogenous substrate has a higher affinity for the uptake process than paraquat (Table 70.4). These studies stimulated work to try and identify the specific cell types into which both paraquat and putrescine are accumulated. Slices of rat lung from rats treated with paraquat, which had been shown to cause selective damage to alveolar type I and type 11 cells, had a decreased ability to accumulate both paraquat and putrescine, suggesting that the transport system resides at least in part in these cell types (Smith et aI., 1976; Smith and Wyatt, 1981). This finding is consistent with the autoradiographic studies reported by Waddell and Marlowe (1980), who showed the distribution of paraquat in mouse lung following iv administration to be consistent with localization in alveolar type 11 cells. Studies with rat lung slices in vitro have shown localization of [3H]-paraquat to alveolar type 11 cells (Wyatt et aI., 1988). Similar studies with rat lung slices using eH]-putrescine, eHJspermine have also shown localization to alveolar type 11 cells and in addition provided evidence for accumulation of radiolabel in bronchiolar Clara cells and possibly alveolar type I cells (Wyatt et aI., 1988). Similar localization of eH]-putrescine was reported by Nemery et al. (1987) and the localization confirmed by electron microscopy to the type I and type 11 alveolar epithelial cells and Clara cells (Dinsdale et aI., 1991). In contrast, in rabbit lung slices [3H]-putrescine was localized to alveolar type 11 cells and macrophages but not in Clara cells (Saunders et aI., 1988). More recent studies with slices of human lung have established that [3H]-putrescine also accumulates into type I and type 11 alveolar epithelial cells (Hoet et aI., 1993). Paraquat accumulation has also been demonstrated in isolated alveolar type 11 cells from rat and rabbit lung (Chen et aI., 1992; Forman et aI., 1982; Horton et aI., 1986) and in isolated Clara cells from rabbit lung (Horton et aI., 1986), suggesting that paraquat transport resides in both cell types. Paraquat is toxic to isolated mouse Clara cells and the addition of putrescine affords some protection (Masek and Richard, 1990). No accumulation of paraquat was, however, detected in isolated rabbit lung macrophages, although Saunders et al. (1988) have reported putrescine accumulation by rabbit lung macrophages. The basis for this difference is currently not clear but it is now well established that polyamine transport systems are present in a number of transformed and nontransformed blood cells (see Smith et aI., 1990). The kinetics of transport of paraquat into isolated type 11 alveolar epithelial cells has been reported by Chen et al. (1992). Using freshly isolated cell suspensions, they found a Km of 88 J.!M with a Vmax of 20 pmollhlJ.!M DNA. They also exam-
1569
ined putrescine transport in these alveolar type 11 suspensions and found a Km of 2.5 J.!M with a Vmax of 33 pmollhlJ.!M DNA. This finding is in broad agreement with that for rat lung slices where the Vmax is very similar for both substrates while the Km for putrescine is higher than that for paraquat (Table 70.4). The accumulation of both spermidine and putrescine has been characterized in rat alveolar type 11 cells in culture (Kameji et aI., 1989; Oreffo et aI., 1991; Richards et aI., 1987). The uptake of spermidine into isolated cells was inhibited by putrescine, spermine, and paraquat as described for slices of rat lung. The accumulation of putrescine has also been studied in human alveolar type 11 cells in culture. The uptake of putrescine and the competitive inhibition by paraquat was essentially the same as that seen in human lung slices (Hoet et aI., 1994). Some difficulties have been experienced by several groups in determining the kinetics of transport of paraquat into isolated alveolar type 11 cells in culture. This may reflect changes to the cell membrane during the isolation procedure, such that the findings in these cells may not accurately reflect that occurring in vivo. A summary of the kinetic constants for the accumulation of both paraquat and putrescine by lung slices and isolated alveolar type 11 cells for rats and humans is shown in Table 70.4. These data show that paraquat and putrescine are accumulated by lung slices and alveolar type 11 cells from both rats and humans and that putrescine has a higher affinity for this system than paraquat. 70.2.13 EFFLUX OF PARAQUAT FROM
THE LUNG The amount of paraquat that accumulates into the lung is determined by both the rate of accumulation and the rate of efflux from the cells in which it concentrates. The loss of paraquat from rat lung following in vivo administration is slow. There appears to be a rapid phase of elimination over the first 20-30 min following iv administration of paraquat which is then followed by slower loss that obeys first-order kinetics with a half-life of about 50 h (Sharp et aI., 1972). Similar studies by Smith et al. (1978) and Dey et al. (1990) showed a rapid phase of elimination that was similar to that reported by Sharp et al. (1972) while the second phase showed a half-life for paraquat loss from the lung of approximately 20 h, which was independent of the plasma concentration. Studies in vitro using lung slices from rats dosed in vivo with paraquat also showed a biphasic elimination, with a rapid loss within 30 min presumably reflecting loss from the extracellular space followed by a slower phase with a half-life of 17 h similar to that seen in vivo (Smith et aI., 1981). Thus, the basis for the selective toxicity of paraquat to the lung resides in paraquat's ability to become concentrated in alveolar type I and 11 cells and Clara cells. The concentration of paraquat retained in the lung is a combination of that retained during the time of the peak plasma concentration, plus that accumulated via the carrier-mediated process. Paraquat, once accumulated into lung cells, is not then readily lost.
1570
CHAPTER 70 Paraquat
70.2.14 BIOCHEMICAL MECHANISMS OF
can also mediate redox cycling of paraquat to produce superoxide anion (Sakai et aI., 1995), indicating that two intracellular enzyme systems are probably involved. Mammalian cells have many enzyme systems which provide them with protection against free radical attack and it is assumed that once these defenses have been overwhelmed that cell death occurs. Superoxide dismutase (SOD) is a family of metalloenzymes that can dismutate superoxide anion to hydrogen peroxide and oxygen:
PARAQUAT TOXICITY Paraquat can be reduced to form a free radical which is stable in aqueous solution in the absence of oxygen (Michaelis and Hill, 1933):
In the presence of oxygen, in biological systems, the radical will rapidly reoxidize to the cation with the concomitant production of superoxide anion O2' (Farrington et aI., 1973): PQ+-
+ 02 ~
PQ2+
O2' + O2' ~ H202
+ O2'
Thus, once paraquat enters a cell it will undergo alternate reduction followed by reoxidation, a process known as redox cycling. Gage (1968b) first reported that the paraquat cation could be reduced by rat liver NADPH-dependent microsomal flavoprotein reductase to form the radical, with the concomitant oxidation of NADPH. Redox cycling of paraquat has also been reported in microsomal preparations of lung, liver, and kidney (Baldwin et aI., 1975) and in lung microsomal and slice systems (Adam et aI., 1990). Studies using antibodies against NADPH-cytochrome c reductase have shown that paraquat radical formation can be blocked, demonstrating a role for this enzyme in the reduction process (Bus et aI., 1974; Horton et aI., 1986). Further support for a key role for NADPH-cytochrome c reductase comes from the studies of Kelner and Bagnell (1989) using a lymphoblastoid cell line with a specific deficiency in this enzyme which they reported was very resistant to paraquat toxicity. Thus, provided there is sufficient NADPH as an electron donor and 02 as an electron acceptor, paraquat will redox cycle inside a cell, generating superoxide anion and consuming NADPH. This reaction is believed to be a key step in the mechanism of paraquat toxicity. However, the biochemical consequences of this reaction which leads to lung cell death are complex and still not fully understood. Recent studies with endothelial cells in culture have indicated that xanthine oxidase
+ 02
The importance of this enzyme in cellular toxicity comes from studies where cellular SOD activity has been genetically modified either by spontaneous mutation or by the transfection of SOD genes. Bilinski and Litwinska (1987) isolated a mutant yeast deficient in SOD activity, which had a greater sensitivity to paraquat than its isogenic wild type. In contrast, Hela cells which possess a higher content of both manganese and copper/zinc SOD had an increased resistance to paraquat (Krall et aI., 1988). Transfection of human copperlzinc SOD into various cell lines also lead to resistance to paraquat toxicity (ElroyStein et aI., 1986; Krall et aI., 1988). Recent studies have shown that mice lacking copperlzinc SOD show a marked increase in sensitivity to paraquat (Fig. 70.5) Sod-/- mice showed a median survival time of about 1.5 days after 10 mg/kg ip, while the Sod+ / - and Sod+ / + mice appeared normal at the end of 7 days of observation (Ho et al., 1998). These studies provide strong evidence for a role for superoxide anion radical in the mechanism of cellular toxicity and for the role of copperlzinc SOD in protecting the lungs against paraquat toxicity. However, superoxide anion itself is unlikely to be the ultimate toxic species as it has limited reactivity in biological systems (Halliwell and Gutteridge, 1984). Dismutation of superoxide anion leads to hydrogen peroxide formation which can undergo detoxification by catalase and glutathione peroxidase. Studies with genetically engineered cells have shown that the balance between these two enzymes plays an important role in cellular toxicity of paraquat. Increasing intracellular concentra-
100 80 "ii > .~
'-.&
60
.&
•
:= III
0~
~
+1+n=11
~
+1-n=18
\
40
- - .& - - -1-
n = 14
.&.
20
--
0 0
2
3
4
5
6
7
8
Time (days)
Figure 70.5 Increased susceptibility of mice lacking CulZn superoxide dismutase to paraquat. The survival times of age-matched, male Sodl +/+, Sodl +/-, and Sodl-/- mice was determined following ip administration of paraquat at 10 mg/kg. From (Ho et aI., 1998). Reproduced with permission from © 1998.
70.2 Toxicity to Laboratory Animals
tions of SOD to high levels can alter the balance of metabolism of hydrogen peroxide from two electron addition via catalase and glutathione peroxidase to produce water to allow an increase in one electron metabolism to form hydroxyl radical: H202 -+e -+ OH· -+e- -+ H2 0
t 2e-
t
catalasejGSH peroxidase
Increasing intracellular SOD content to a very high level ultimately leads to an increase in toxicity to paraquat in a number of transfected cells or Escherichia coli (Bloch and Ausubel, 1986; Elroy-Stein et aI., 1986; Scott and Eaton, 1996; Scott et aI., 1987). In contrast, cells having an increase in both SOD and catalase exhibited a greater resistance to paraquat than with just SOD alone (Krall et aI., 1988). Generation of hydroxy 1radical has been proposed as the critical event in the toxicology of paraquat. This reaction requires the presence of iron and is generated by the Fenton reaction. In this reaction ferrous ions react with hydrogen peroxide to generate hydroxyl radicals: Fe2+ + H202 -+ Fe3 + + OH- + OH· Under physiological conditions free iron predominately exists in the ferric form (Fe3+) as a chelate with ADP, ATP, and citrate. The reduction of ferric iron may be achieved directly by the paraquat radical (Sutton et aI., 1987; Winterbourn and Sutton, 1984) or indirectly by superoxide anion generated from the redox cycling of paraquat (McCord and Da, 1987). A role for transition metals such as iron in the toxicity is supported by studies showing that paraquat toxicity is reduced by removal of iron and enhanced by its addition (Kohen and Chevion, 1985; Sion et al., 1989; Van der Wal et al., 1990). The role of the iron chelator desferrioxamine in affording some protection against paraquat toxicity will be discussed in the section on antidotes. Many other studies too numerous to mention have been conducted both in vitro and in vivo to explore the effect of altered antioxidant status on the toxicology of paraquat. Examples include the role of GSH and GSH reductase (Bus et aI., 1976a; Hardwick et aI., 1990; Keeling et aI., 1982), the role of selenium deficiency, vitamin E, and glutathione peroxidase (Block, 1979; Bus et aI., 1976b; Cagen and Gibson, 1977; Kelner et aI., 1995; Omaye et aI., 1978), and the role of metallothionein (Lazo et aI., 1995; Satoh et aI., 1992). Metallothionein appears to have play a role as a free radical scavenger in addition to its well established role as a heavy metal chelator. Metallothionein has been reported to quench both superoxide anion and hydroxyl radicals, with a significantly higher reactivity toward hydroxyl radicals (Thornalley and Vasak, 1985). Genetically engineered animals have been used as tools to elucidate the function of the various antioxidant defense mechanisms against paraquat-induced oxidant injury. In addition to the discussion above regarding mice deficient in copperlzinc SOD, Sato et al.
1571
(1996) found mice deficient in metallothionein I and 11 genes to be more susceptible to paraquat toxicity. Glutathione peroxidase deficient mice show an increased susceptibility to paraquat toxicity with a mean survival time of 5 h compared to the wild type of 69 h following an ip dose of 50 mg/kg (Cheng et aI., 1998). Mice overexpressing glutathione peroxidase are more tolerant to paraquat toxicity; wild type mice given a large 125 mg/kg ip dose of paraquat died within 5 h while the mice overexpressing the enzyme lived for about 54 h (Cheng et aI., 1998). Figure 70.6 shows a schematic representation of the key requirements to enable paraquat to enter a cell and the subsequent redox cycling steps believed to lead to cytotoxicity. Three hypotheses have been proposed to account for the ensuing cytotoxicity, one involving lipid peroxidation, another the oxidation of NADPH, and the third mitochondrial toxicity; none of these hypotheses are mutually exclusive. 70.2.15 LIPID PEROXIDATION HYPOTHESIS
Bus and co-workers (1974, 1976a) proposed the sequential generation of superoxide anion and hydroxyl radical and the initiation of lipid peroxidation as the mechanism of cellular toxicity of paraquat. However, there is little direct evidence which demonstrates lipid peroxidation occurs in the lung of animals dosed with paraquat before there is morphological evidence of cell damage. Paraquat-induced lipid peroxidation has been demonstrated in vitro in broken cell systems and isolated cells from the lung and liver (Aldrich et aI., 1983; Bus et aI., 1976a; Kornbrust and Mavis, 1980; Saito et aI., 1985; Sandy et aI., 1986; Sata et aI., 1983; Trush et aI., 1981) and in vivo (Burk et aI., 1980; Bus et aI., 1976b; Reddy et aI., 1977). However, others have questioned its significance in the toxicity. For example Steffen et al. (1980) only found a small increase in the exhalation of ethane (a marker of lipid peroxidation) in rats suffering from respiratory distress following exposure to paraquat and oxygen. Similarly, others have been unable to find evidence of lipid peroxidation in the lungs of mice given large doses of paraquat (Shu et al., 1979; Younes et aI., 1985) or it is only detected as a late event in the toxicity (Ogata and Manabe, 1990). So the question remains as to whether lipid peroxidation is a cause, or a consequence, of the toxicity. These contrasting findings in vivo may also reflect the difficulty in detecting a small but critical increase in lipid peroxidation in the alveolar type I and 11 cells and Clara cells that are only a small population of the total cells in the lung. 70.2.16 OXIDATION OF NADPH HYPOTHESIS
Intracellular redox cycling of paraquat results in the oxidation of NADPH leading to cellular depletion such that those cells that selectively accumulate paraquat can longer function normally. Fisher et al. (1975) first suggested that the redox potential of lung cells may be altered by the redox cycling of paraquat. A marked stimulation of the activity of the pentose
1572
CHAPTER 70 Paraquat
Hoxose
II ...
qA_."
~
PO
NADPH
+~;='\+ CH3N~NCHl
NADP+
pQ2+'>..
/
IT]
02 Putrescine
then
+
+ NH 3- (CH 2 >4- NH3
-
0.622 nm----l
(':::xt2\
GSSG
GSH
~=/ i
~
NAOPH
02
+ H Fe + 0;
Fe H
Alveolar ~ epithelial .; cell " membrane
PQ+'
>-<
f---- 0.702 nm ---1
f--
IT]
- 2H+
0; --H 20
2
·· .... .. ·· .... · .... ·.. ·0)
O~ --Fe H +02 .................
®
+ H20 z-OH·+OH+Fe3+..... ® I
1
0
Lipid peroxidatioo
1
Cell death
NADP+
~ Hox ...
-
manopho~t.
Figure 70.6 Mechanism of toxicity of paraquat. A schematic representation of the mechanism of toxicity of paraquat. I = structure of paraquat and putrescine showing the geometric standards of the distance between the nitrogen atoms; 2 = transport system which recognizes paraquat, minimum separation of charge of approximately 0.5 nrn ; 3 redox cycling or paraquat utilizing NADPH; 4 formation of hydroxyl radical leading to lipid peroxidation; 5 = detoxification of H202 via glutathione reductase/peroxidase couple, utilizing NADPH. From (Smith, 1987). Reproduced with permission from © 1987.
=
phosphate pathway in the lung has been observed following exposure to paraquat (Bassett and Fisher, 1978; Fisher et aI., 1975; Fisher and Reicherter, 1984; Keeling et aI. , 1982; Rose et al., 1976b). Since this pathway represents the major cellular source of NADPH, it is inferred that this response represents an attempt by lung cells to maintain their levels of reducing equivalents under conditions of oxidative stress. In those cells in which paraquat is accumulated, the concentration may be very high and result in very fast rates of NADPH oxidation. If the rate of consumption exceeds the rate of formation via the pentose phosphate pathway, the concentration of NADPH will fall below that required to maintain cell viability. Witschi et al. (1977) first demonstrated that the NADPHlNADP+ ratio in the lungs of rats dosed iv with paraquat was decreased, suggesting that oxidation of the reduced nucleotide had occurred. Later studies by Keeling and Smith (1982) demonstrated that the shift in NADPHlNADP+ ratio in the lung following sc administration of paraquat was the result of NADPH loss from the lung. A consequence of depletion of cellular NADPH is that the cell shuts down it synthetic pathways which are dependent on this nucleotide, such as the synthesis of fatty acids (Keeling et aI., 1982). A loss of NADPH may also have particular importance for alveolar type 11 cells which produce pulmonary surfactant (Brigelius et aI., 1986). NADPH is also consumed in an attempt by the lung to detoxify hydrogen peroxide that is formed via the glutathione peroxidase/reductase enzyme system (Fig. 70.6) to regenerate
=
reduced glutathione (GSH) from its oxidized form (GSSG). In general large changes in lung GSH and GSSG are not seen after paraquat administration (Bus et aI., 1976a; Keeling and Smith, 1982; Shu et aI., 1979; Reddy et aI., 1977). This may explain why lipid peroxidation has not been conclusively demonstrated in vivo as this would not become apparent until both NADPH and GSH were markedly reduced. However, formation of protein mixed disulphides is increased in the lung in vivo (Keeling et aI., 1982; Keeling and Smith, 1982) and in perfused liver (Brigelius et aI., 1982). These changes in protein mixed disulphides in the lung are presumably a response to oxidative stress and may not be critical to the cellular toxicity. This notion is supported by studies with the bipyridyl diquat which can undergo redox cycling in the lung (Rose et aI., 1976b; Witschi et al., 1977). Diquat also produced increases in protein mixed disulphide content in the lung without affecting NADPH content at a dose that did not cause lung injury (Keeling and Smith, 1982). This indicates that NADPH depletion subsequent to redox cycling is a critical step in the mechanism of paraquat toxicity.
70.2.17 THE ROLE OF MITOCHONDRIA IN THE TOXICITY Another hypothesis that has been proposed is that paraquat toxicity is due to mitochondrial damage, based on morphological findings of early mitochondrial changes in alveolar type 11 cells
70.2 Toxicity to Laboratory Animals
(Hirai et aI., 1985). Ultrastructural studies of the time course of development of paraquat-induced lung injury have also reported early changes to mitochondria such as swelling and altered staining density (Keeling et aI., 1981; Smith and Heath, 1974a; Sykes et aI., 1977). These mitochondrial changes were also observed in the lungs of rats exposed to paraquat and 85% oxygen, which enhances paraquat toxicity to the lung (Keeling et aI., 1981). However, as discussed with regard to the lipid peroxidation hypothesis, the question is: are the effects on mitochondria a cause, or a consequence, of paraquat toxicity? Early studies with isolated liver mitochondria reported only minor changes in mitochondria respiration by paraquat (Gage, 1968b). More recent studies have reported that paraquat cation can be reduced by NADH-ubiquinone oxidoreductase (Complex I) located on the inner mitochondrial membrane (Fukushima et aI., 1993; Shimada et aI., 1998). These authors also showed that paraquat was able to stimulate lipid peroxidation in submitochondrial particles (Yamada and Fukushima, 1993). These findings show that mitochondria have the potential to generate superoxide anion from paraquat provided it can gain access. In general, studies with intact mitochondria support the orginal findings of Gage (1968b) showing that little or no effects are seen (Costantini et aI., 1995; Lambert and Bondy, 1989) unless very high concentrations of paraquat are present (Kopazyk-Locke, 1977; Yamamoto et aI., 1987; Thakar and Hassan, 1988; Palmeira et aI., 1995). Paraquat has been shown to induce a Ca2+dependent permeability transition of the inner mitochondrial membrane leading to membrane depolarization, uncoupling, and matrix swelling in isolate rat liver mitochondria (Costantini et aI., 1995). This opening of the membrane permeability pore does not occur in the absence of added Ca2+and requires the presence of rotenone, leading one to question the relevance of this observation to the in vivo situation. It seems likely that any intracellular increases in Ca2+ would only occur once paraquat had entered the lung cell, undergone redox cycling, and altered mixed disulphide status. In summary, mitochondrial damage has been observed in the lung prior to cell death; it seems likely that this response is secondary to changes taking place in the cytosol. 70.2.18 THE INVOLVEMENT OF OXYGEN
As discussed earlier, the redox cycling of paraquat to form superoxide anion requires oxygen and hence oxygen plays a critical role in the toxic process. It has been known for many years that hyperoxia is toxic to the lung, causing damage to endothelial cells through a mechanism that involves the formation of reactive oxygen species (Frank and Massaro, 1979; Jenkinson, 1982). One of the therapeutic measures for anoxia in human cases of paraquat poisoning was the addition of air supplemented with oxygen (see treatment of human poisoning). However, it has been shown that increasing the oxygen concentration potentiates the lethality of paraquat to rats (Douze and van Heijst, 1977; Fisher et aI., 1973; Keeling et aI., 1981; Kehrer et aI., 1979) by increasing the injury to the lung. The
1573
converse is also true; rats exposed to paraquat in a hypoxic environment are protected relative to those exposed to paraquat in air (Rhodes et aI., 1976). Detailed histopathology on the lungs of rats exposed to paraquat alone or paraquat in an atmosphere of 85% oxygen showed that the damage was primarily localized to the alveolar type I and 11 cells with little evidence of endothelial cell damage, showing that oxygen potentiated paraquat toxicity (Keeling et aI., 1981). These findings have recently been reproduced using isolated rat and human alveolar type 11 cells exposed to either paraquat in air or paraquat and increasing concentrations of oxygen. Increasing the oxygen concentration in the atmosphere potentiated the toxicity of paraquat, while lowering the oxygen concentration to 10% afforded some protection (Hoet et aI., 1997). The mechanism underlying this synergistic effect of oxygen on paraquat toxicity is not entirely clear. It seems unlikely that oxygen would normally be rate limiting for paraquat to redox cycle. A more likely explanation is that the cellular defense mechanisms that protect against oxygen and paraquat toxicity are more rapidly overwhelmed. In summary, the key events leading to cellular toxicity are (1) accumulation of paraquat into the cell and (2) its ability to redox cycle and produce oxidative stress. It seems likely that a combination of depletion of NADPH plus the generation of hydroxyl radical leading to lipid peroxidation and mitochondrial dysfunction is involved but the precise temporal relationships have not as yet been established. 70.2.19 EFFECTS ON THE KIDNEY
The major route of elimination for paraquat once it has entered the bloodstream is via the kidneys where it is actively secreted by organic cation transport systems (see review by Chan et aI., 1998). This process becomes saturated at fairly low plasma concentrations (3-4 nmoIfml; 0.5-0.7 !1-g/ml) in the rat (Chan et aI., 1997). At higher plasma concentrations paraquat is nephrotoxic. Large oral or systemic doses administered to rats or mice produce morphological changes to the proximal renal tubules, including hydropic degeneration with occasional evidence of necrotic epithelial cells and of renal tubular regeneration (Clark et aI., 1966; Lock and Ishmael, 1979). Chronic exposure to mice via their drinking water showed ultrastructural evidence for proliferation of smooth endoplasmic reticulum and the presence oflipid containing bodies in proximal tubule cells (Fowler and Brooks, 1971). Renal tubular necrosis is more marked in the dog and rabbit following large toxic doses with clear evidence of degeneration of proximal tubular cells with the presence of casts in the tubular lumen (Clark et aI., 1966; Giri et aI., 1982; McElligott, 1972; Nagata et aI., 1992a; Yonemitsu, 1986). Prior to the onset of renal tubular necrosis, paraquat-induced renal functional changes occur including diuresis, albuminuria, glucosuria, and elevations in plasma urea and creatinine in the rat (Lock and Ishmael, 1979), dog (Giri et aI., 1982; Nagata et aI., 1992a), and cynomologus monkey (Purser and Rose, 1979). The precise mechanism of renal functional impairment
1574
CHAPTER 70
Paraquat
is not known; it probably involves altered renal hemodynamics as well as accumulation of paraquat into proximal renal tubules leading to cellular necrosis. There is some evidence that paraquat may reduce renal blood flow based on the finding of elevated renal plasma renin activity in the dog after dosing (Giri et aI., 1982) and hypovolaemia in the rat (Lock, 1979). Paraquat is thought to enter renal tubular cells by an organic cation transport system, thereby enabling it to concentrate to many times that present in the plasma (Chan et aI., 1996a, 1996b, 1997; Ecker et aI., 1975a; Groves et aI., 1995; Hawksworth et aI., 1981; Lock and Ishmael, 1979; Wright and Wunz, 1995). The accumulation can be blocked by other organic cations such as tetraethylammonium and quinine but is not affected by the polyamines, putrescine, or spermine (Chan et aI., 1996a; Groves et aI., 1995). Thus the accumulation of paraquat into renal tubular cells occurs via a different transport system to that which leads to its accumulation in the lung. Once inside a renal tubular cell paraquat can redox cycle (Baldwin et aI., 1975; Tomita, 1991), producing superoxide anion and hence trigger the cascade of biochemical events leading to cytotoxicity similar to that discussed for the lung (Lock and Ishmael, 1979; Molck and Friis, 1997). Regardless of the mechanism, the consequence of a reduced renal excretion is that more paraquat is available in the plasma to accumulate into the lung. Thus, maintenance of renal function to facilitate paraquat excretion from the body is critically important for cases of human poisoning (see later). 70.2.20 EFFECTS ON THE CENTRAL NERVOUS SYSTEM
No signs of neurotoxicity or neuropathological changes have been reported following oral gavage or dietary administration of paraquat to rodents or dogs (IPCS, 1984). Paraquat as a di-cation does not readily cross the blood-brain barrier and enter the rat brain after either oral or systemic administration (Corasaniti et aI., 1991; Corasaniti and Nistico, 1993; Dey et aI., 1990; Naylor et aI., 1995; Rose et aI., 1976a; Widdowson et aI., 1996a, b). The concentration associated with the rat brain is always lower than that in the plasma and decreases with time. The initial concentration detected in the brain may be largely associated with blood (Dey et aI., 1990; Naylor et aI., 1995; Rose et aI., 1976a). Paraquat was, however, detected in brain regions such as the olfactory bulb, area postrema, and hypothalamus, which do not possess an effective blood-brain barrier. Autoradiographic studies have detected paraquat in these regions and in the cerebrospinal fluid (ventricles and choroid plexus) but the concentrations were low and only represent a very small percentage of the administered dose, about 0.05% at the time of maximal blood concentration 1 h after dosing (Naylor et aI., 1995; Waddell and Marlowe, 1980). Immunohistochemical localization of paraquat in rat brain has shown it is present in capillary walls and glial cells but was not detected in neurones (Nagao et aI., 1991). Recent studies in the rat, using parenteral doses of paraquat at or above the MLD (20-100 mg/kg, ip), produced signs
of neurotoxicity with muscle fasciculation, some tremors and "wet-dog" shakes, and at the higher doses myoclonus, typically within 30 min of dosing (Bagetta et al., 1992; Corasaniti et aI., 1992; Hara et aI., 1993), which is the time of peak blood and brain concentrations. These authors also reported neuronal cell necrosis in the pyriform cortex of these animals 24 h after dosing (Bagetta et aI., 1992; Corasaniti et aI., 1992). The neuronal cell necrosis could be reduced by administration of atropine but not methyl atropine (Bagetta et aI., 1992), suggesting some involvement of central muscarinic receptors. No effects were seen after 5 mg/kg ip paraquat. The basis for the selective injury to the pyriform cortex is currently not known, but it does not reflect the brain region with the highest concentration of paraquat (Corasaniti and Nistico, 1993; Naylor et aI., 1995). Others have reported that paraquat (20 mg/kg, sc) does not produce neuronal cell necrosis in the pyriform cortex of perfused-fixed material from rats 24 and 48 h after dosing (Naylor et aI., 1995; Widdowson et aI., 1996a) and have suggested the effect reported by the Italian group may be a fixation artifact. The precise basis for this variance is currently not understood. Similarly, daily oral dosing of paraquat at 5 mg/kg/day for 14 days to rats produced no evidence of neuronal cell necrosis, despite particular emphasis on the pathology of the pyriform cortex, nigro-stratial region, and hypothalamus or behavioral changes indicative of neurotoxicity (Widdowson et aI., 1996b). Direct administration of paraquat into the ventricles or infusion into certain brain regions produced signs of neurotoxicity in rats which were associated with neuronal cell damage (Bagetta et aI., 1992, 1994; Calo et aI., 1990; Corasaniti et aI., 1992; De Gori et aI., 1988; Liou et aI., 1996; Liu et aI., 1995; Yoshimura et aI., 1993). These effects were seen at low doses of paraquat 2-20 J.!g injections. These observations lend support to the view that little paraquat enters the brain following systemic administration (20 mg/kg, sc or 4000 J.!g1200 g rat) or oral administration (126 mg/kg or 25,200 J.!g1200 g rat) as no neuronal cell toxicity was seen at these doses. Comparisons have been drawn to the structural similarity between paraquat and 1-methyl-4-phenylpyridinium ion (MPP+), the active metabolite of l-methyl-4-phenyl-1,2,3,6tetrahydropyridine (MPTP) which can induce a Parkinson-like syndrome in monkeys and humans. Administration of MPTP to susceptible animal species produces selective damage to dopaminergic neurons in the substantia nigra leading to a marked loss of dopamine and clear signs of neurotoxicity. The mechanism for MPTP toxicity (see Markey et aI., 1986; Tipton and Singer, 1993) is due to its ability to cross the blood-brain barrier and enter glial cells where it can undergo oxidative metabolism by the enzyme monoamine oxidase B to form MPP+. This metabolite then accumulates selectively into dopaminergic neurons via the dopamine transport system, leading to inhibition of mitochondrial respiration which ultimately leads to the demise of the neurone. Structure-activity relationships suggest that, despite their apparent similarity, paraquat and MPTP are two very different chemicals (Koller, 1986). MPTP is uncharged and lipophilic and thereby able to cross the blood-brain barrier, whereas paraquat is charged and hy-
70.2 Toxicity to Laboratory Animals drophilic and does not readily enter the brain. Also, MPTP is a monoamine whose metabolite MPP+, the proximate toxin, is able to use a specific uptake system, particularly in the substantia nigra, whereas paraquat is a diamine. It is also very relevant that administration of MPP+ to experimental animals did not produce neurotoxicity, due to its poor entry across the bloodbrain barrier (Tipton and Singer, 1993). Thus like paraquat, MPP+ does not readily enter the brain. Consistent with this, systemic administration of paraquat to C57 black mice or rats did not lead to dopamine depletion or neuronal cell death in the striatum, like that seen with MPTP (Perry et aI., 1986; Widdowson et aI., 1996b). Others have reported changes in brain dopamine content following paraquat administration to mice (Endo et al., 1988; Fredriksson et aI., 1993). In the latter case, paraquat was administered to pups on days 10 and 11 after birth at a time when the brain is undergoing rapid growth and hence might be a more vulnerable to chemical insult. The authors reported a small loss of dopamine and its metabolites and a decreased behavioral activity when measured at about 4 months of age (Fredriksson et al., 1993). This suggests that the developing brain is potentially more sensitive to insult. However, adverse effects have not been detected in developmental toxicity or multi generation studies, where paraquat was given to pregnant rats and their offspring (see earlier section). Attempts to reproduce the findings of Fredriksson et al. (1993) in C57 black mice, in another laboratory, have not proved possible (David Ray, personal communication). Thus, paraquat as a charged di-cation does not readily enter the brain. The behavioral effects observed in rats only occur at lethal systemic doses. 70.2.21 EFFECTS ON OTHER ORGANS
Following oral ingestion of paraquat by humans, ulceration of the pharyngeal, oesophageal, and gastric mucosa has been reported (see Section 70.3). In animal studies there is often no direct contact with these tissues when gavage dosing is employed. However, focal necrosis of the gastrointestinal tract has been observed in primates, demonstrating the topical irritant nature of high oral doses of paraquat (Murray and Gibson, 1972). Paraquat administration to the rat produced an increased synthesis of liver glycogen and an increase in blood glucose that appeared to be mediated by the adrenal, since adrenalectomy prevented these changes (Rose et aI., 1974b). These effects seen following paraquat and the related bipyridyl diquat are thought to be due to catecholamine release and high circulating concentrations of corticosteroids (Rose et aI., 1974b). This response is thought to be unrelated to the pulmonary damage produced by paraquat but may account for some of the effects seen with paraquat on the adrenal and lymphoid tissues such as the spleen and thymus (Butler and Kleinerman, 1971; Clark et aI., 1966; Fisher et aI., 1973). The increase in circulating corticosterone seen with paraquat can also be prevented by lesioning the area postrema (Edmonds and Edwards, 1996). This area of the brain also controls the taste aversion to paraquat seen in rats (Dey et aI., 1987; Edmonds and Edwards, 1996).
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Liver damage is not a major finding after paraquat administration: after large doses some central lobular necrosis has been reported in most species examined (Cagen et aI., 1976; Clark et al., 1966; Giri et al., 1982; Murray and Gibson, 1972; Nagata et al., 1992a). Since paraquat is delivered to the liver following dosing and hepatocytes possess the relevant enzymes to facilitate redox cycling, presumably paraquat does not normally accumulate in hepatocytes to a sufficient concentration to overwhelm the protective antioxidant defence enzymes and produce necrosis. However, both mice and rats made selenium deficient show marked liver injury following paraquat administration (Burk et aI., 1980; Cagen and Gibson, 1977) supporting the view that selenium dependent enzymes such as glutathione peroxidase play an important protective role. These findings are consistent with recent studies using transgenic mice where glutathione peroxidase has either been deleted or overexpressed, showing that this selenium-dependent enzyme plays a key role in paraquat-induced tissue injury (Cheng et aI., 1998; de Haan et aI., 1998). 70.2.22 TREATMENT OF POISONING IN ANIMALS
Over the past 30 years a variety of attempts to modify the toxicity of paraquat in experimental animals have been examined. To date the only approach that has been shown to clearly reduce mortality in rats is purgation of the gastrointestinal tract with a diatomaceous clay (bentonite or Fuller's earth) along with a cathartic (e.g., magnesium sulphate) (Clark, 1971; Smith et aI., 1974). Attempts to modify paraquat toxicity have been based on its known mechanism of toxicity and will be briefly discussed under the following headings: (a) prevention of absorption from the gastrointestinal tract, (b) removal from the bloodstream, (c) prevention of accumulation into the lung, (d) attempts to scavenge oxygen free radicals, and (e) attempts to prevent lung fibrosis. This aspect of paraquat toxicity has been reviewed by others. See Bateman (1987), Meredith and Vale (1995), Jaeger et al. (1995), and Section 70.3 on human poisoning in this review. 70.2.23 ADSORPTION FROM THE GASTROINTESTINAL TRACT
As discussed earlier, paraquat is poorly absorbed from the gastrointestinal tract and therefore attempts to reduce its entry into the bloodstream could be beneficial. Peak blood levels are detected within 60-90 min in rats, dogs, and monkeys (Figs. 70.2 and 70.3). Therefore any interventions must be taken quickly after poisoning if they are to be effective. The bipyridilium herbicides have been shown to bind very strongly to soil and clay minerals (Knight and Tomlinson, 1967). Clark (1971) demonstrated that bentonite and Fuller's earth where able to reduce mortality in rats given a lethal dose of paraquat when delayed for 2 or 3 h after paraquat administration. Smith et al. (1974) subsequently showed that repeated doses of a bentonite, castor
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CHAPTER 70 Paraquat
oil, and magnesium sulphate mixture protected rats against the lethal effects of paraquat when given 4 h after exposure and this regimen also reduced mortality when delayed for as long as 10 h after exposure. The basis for the protection was shown to be due to a reduction in the concentration of paraquat in the bloodstream and a concomitant reduction in the amount accumulated into the lung (Smith et aI., 1974). Several other absorbent or binding agents have been examined. Activated charcoal was shown to be very effective in rats (Okonek et aI., 1982) and in mice in combination with magnesium citrate, the magnesium salt affording some protection on its own (Gaudreault et aI., 1985). Kayexa1ate (sodium polystyrene sulphate), Kalimate (calcium polystyrene sulphate), sodium dextrin sulphate, sodium glucose sulphate, and a variety of alkylsulphates and alkylsulphonates have been shown to afford some protection in rats and mice (Nokata et aI., 1984; Tsuchiya et aI., 1995; Ukai et aI., 1987). The use of this approach in clinical practice will be discussed in more detail later. 70.2.24 REMOVAL FROM THE BLOODSTREAM Both peritoneal dialysis and hemodialysis have been suggested for removing paraquat from the bloodstream and thereby reducing availability to the lungs. Charcoal hemoperfusion was initially demonstrated to remove paraquat from the blood of beagle dogs (Maini and Winchester, 1975). Hemoperfusion appeared to reduce mortality in dogs when given within 12 h of administration of paraquat (Widdop et aI., 1975), although more recent studies in the dog have indicated that unless started within 2 h of exposure it is unlikely to reduce the paraquat content in the lungs (Pond et aI., 1993). 70.2.25 PREVENTION OF ACCUMULATION INTO THE LUNG Paraquat is actively transported into alveolar type I and 11 cells where it accumulates. Studies in vitro using polyamines, diaminoalkanes, and a number of other chemicals have identified chemicals that can reduce paraquat accumulation (Gordonsmith et aI., 1983; Lock et aI., 1976; Maling et aI., 1978; Ross and Krieger, 1981; Smith and Wyatt, 1981). However, attempts to reduce paraquat mortality in rats with these agents have failed to demonstrate significant protection (Dunbar et aI., 1988; Maling et aI., 1978). Another approach has been to use antibodies to paraquat (polyclonal, monoclonal, or specific Fab fragments) to try and reduce toxicity to the lung. This approach has been shown to reduce paraquat uptake and cytotoxicity in rat lung slices and isolated alveolar type 11 cells (Chen et aI., 1994; Wright et aI., 1987a). However, treatment of paraquat-intoxicated mice (Cadot et aI., 1985; Wright et aI., 1987b) or rats (Nagao et aI., 1989) by immunotherapy did not reduce the concentration of paraquat in the lung or affect the mortality.
70.2.26 FREE RADICAL SCAVENGING As discussed earlier, once inside a cell paraquat can redox cycle and produce superoxide anion, singlet oxygen, and hydroxyl radicals. Many studies have been aimed at attempting to scavenge the radicals formed to reduce or protect the lung injury. In many of these cases significant protection can be demonstrated using isolated cell systems, but in whole animals the protection is limited or equivocal. Superoxide dismutase has been reported to increase survival in rats exposed to paraquat (Autor, 1974; Wasserman and Block, 1978), while other studies have failed to confirm these observations (Frank, 1983; Patterson and Rhodes, 1982). The short plasma half-life of exogenous superoxide dismutase and the fact that it does not enter cells accounts for the lack of protection. A more recent report indicated that a low molecular weight metalloporphyrin superoxide dismutase mimetic afforded some protection against paraquat-induced injury to the lung, but its effect on mortality was not examined and its effect is likely to have been marginal (Day and Crapo, 1996). Desferrioxamine (DF) is an iron chelating agent which has been used to scavenge free iron and thereby reduce hydroxyl radical production. Studies in mice suggested that DF given 24 h before and regularly after an acute dose of paraquat reduced mortality (Kohen and Chevion, 1985). In this same model these workers showed that iron increased paraquat toxicity. In rats, however, DF appeared to afford no protection (Hoffer et aI., 1992; Osheroff et aI., 1985). Van Asbeck et al. (1989) gave DF by continuous infusion to vitamin E deficient rats and showed it prevented the lung injury and hence reduced mortality. This group also examined the effect of DF and CP51 an hydroxypyridin-4-one iron chelator in rats with a normal vitamin E status and found no protection with DF while CP51 increased survival (Van der Wal et aI., 1992). Xanthine oxidase inhibitors may also reduce superoxide anion formation and rats fed a diet rich in tungstenate showed a better survival following paraquat exposure than rats fed the diet alone (Kitazawa et aI., 1991). Clofibrate induces hepatic peroxisomes in rodents and thereby increases hepatic catalase activity, and it was postulated that a similar effect in the lung might afford protection against paraquat toxicity. Prior administration of clofibrate to rats for 6 days followed by paraquat afforded significant protection. However, when clofibrate was administered after paraquat it gave no protection (Frank et aI., 1982). Vitamin E is a lipid soluble antioxidant and radical scavenger. Some early studies showed that vitamin E deficient animals were more susceptible to paraquat than those with a normal vitamin E status (Block, 1979; Bus et aI., 1975). Acute administration of vitamin E to normal mice or rats did not, however, significantly protect against the toxicity (Bus et aI., 1976a; Redetzki et aI., 1980) even when instilled into the trachea in a lipo some either alone or in combination with reduced glutathione (Suntres and Shek, 1995, 1996). The protective effect of selenium has been reported, animals fed selenium deficient diets being more sensitive to paraquat
70.3 Toxicity to Humans toxicity (Cagen and Gibson, 1977; Omaye et aI., 1978). This is probably related to the selenium-dependant enzyme glutathione peroxidase which plays an important role in protecting cells against oxidative stress. Evidence that glutathione peroxidase plays a key role in protecting animals against paraquat toxicity comes from recent studies in transgenic mice where deletion of this enzyme enhances toxicity while addition affords some protection (Cheng et aI., 1998; de Haan et aI., 1998). Vitamin C, water soluble antioxidant, has provided equivocal data with one study suggesting it might protect while others showed it either had no effect or enhanced paraquat toxicity (Matkovics et aI., 1980; McArn et aI., 1980; Minakata et aI., 1996; Montgomery et aI., 1982; Sullivan and Montgomery, 1984). A combination of vitamin C and riboflavin in rats produced a significant improvement in paraquat mortality (Schvartsman et aI., 1984), while vitamin C or riboflavin alone was not protective. These authors suggested that perhaps the combination of antioxidant plus an effect of riboflavin on glutathione reductase activity may have contributed to the protection. Niacin has been reported to modestly reduce paraquat mortality in rats. This may be due to an effect of niacin on NAD synthesis which is reduced by paraquat (Brown et aI., 1981); however, subsequent studies were unable to confirm any protection with niacin (Hooper et aI., 1983). A number of sulphydryl compounds have been examined based on their antioxidant ability, and on an early observation by Bus et al. (1976b) showing that diethylmaleate, which depletes glutathione, enhanced paraquat toxicity. In general precursors of glutathione synthesis which increase intracellular cysteine content have been shown by some workers to provide some increased survival in mice or rats, while others have found these reagents to produce equivocal effects. The protection may be due to alteration of the pharmacokinetics of paraquat or induction of some of the enzymes involved in providing protection against free radical damage. The following have been examined N-acetylcysteine (Cramp, 1985; Hoffer et aI., 1993; Hybertson et aI., 1995; Shum et aI., 1982; Wegener et aI., 1988), glutathione (Matkovics et aI., 1980; Szabo et aI., 1986), cysteine and cystine (Kojima et aI., 1992; Szabo et aI., 1986), L-2-oxothiazolidine-4-carboxylate (Ali et aI., 1996); Dpenicillamine (Szabo et al., 1986), and sulphite or thiosulphate (Yamamoto, 1993). The effect of the lung-surfactant stimulating drug ambroxol has been examined in rats and was shown to increase the rate of survival after paraquat (Salmona et aI., 1992) while Nemery et al. (1992) found no protective effect. 70.2.27 PREVENTION OF LUNG FIBROSIS
Since delayed deaths with pulmonary fibrosis are a characteristic of paraquat poisoning in experimental animals and humans (see later) a number of agents have been examined to ameliorate the fibrotic response. Immunosuppressants such as methylprednisolone, dexamethasone, and cyclophosphamide have been
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examined in experimental animals and in general they were either without effect (Seidenfeld, 1985) or only afforded some protection when give prior to paraquat, but not when given simultaneously (Kitazawa et aI., 1988; Reddy et aI., 1976; Smith and Watson, 1987). Lung irradiation (Saenghirunvattana et aI., 1992) and collagen synthesis inhibitors such as D,L-3,4dehydroproline (Akahori and Oehme, 1983) were not effective in reducing paraquat lung damage. A recent suggestion has been mechanical ventilation with additional inhalation of nitric oxide, based on nitric oxide's vasodilatory effect on the lungs (Berisha et aI., 1994). This approach has not been examined in experimental animals but some clinical experience in combination with other antidotes has been examined (see later). In summary, removal of the ingested material by emesis and purgation of the gastrointestinal tract is currently the most effective method after paraquat exposure in experimental animals. As discussed later, a cocktail of many of these approaches is often used in cases of human poisonings.
70.3 TOXICITY TO HUMANS 70.3.1 EXPERIMENTAL EXPOSURE
The percutaneous absorption of radiolabelled paraquat has been determined in humans (Wester et aI., 1984). Following application of 9 f..lg/cm 2 the amount absorbed was 0.29% for the leg, 0.23% for the hand, and 0.29% for the forearm. This gave a calculated in vivo absorption rate of 0.03 f..lg/cm 2 for the 24 h exposure period. Paraquat was thus only minimally absorbed, especially in comparison with other commonly available pesticides (Wester and Maibach, 1985). 70.3.2 ACCIDENTAL AND INTENTIONAL POISONING
The first case fatalities described involved accidental ingestion of the 20% paraquat concentrate (Bullivant, 1966; Campbell, 1968; Oreopoulos et aI., 1968; Swan, 1967). A major source of poisoning was the decanting into unlabelled drinks bottles and other containers (Malone et aI., 1971). Throughout the 1970s the number of reported cases continued to rise; however, there was a noticeable shift in the circumstances. For example, in the Republic of Ireland the number of accidents due to decanting decreased between 1967 and 1977 from 45% to 4% of total cases (Fitzgerald et aI., 1978b). Further analysis of the circumstances of poisoning showed that before 1975 there was an approximately equal proportion of accidental and suicidal cases, whereas after that date suicides accounted for over 90% of cases and all fatalities. A similar pattern was described in Northern Ireland (Carson and Carson, 1976) and the United Kingdom (Bramley and Hart, 1983; Howard, 1979a). A review of deaths from pesticide poisoning in the United Kingdom between 1945 and 1989 showed that the number of paraquatassociated deaths rose continuously from 1973 onward and
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CHAPTER 70
Paraquat
peaked in 1981. Since then, the number has steadily declined to pre-1973 levels (Casey and Vale, 1994). With the increasing use of paraquat throughout the world during the 1970s and 1980s it became apparent that the problem of accidental and intentional poisoning had shifted away from the British Isles and Europe (Onyon and Volans, 1987). A high incidence was reported in particular from Asian countries such as Japan (Naito and Yamashita, 1987), Malaysia (Amarasingham and Lee, unpublished report), Sri Lanka (Hettiarachi and Kodithuwakku, 1989), and Fiji (Goundar, 1984). Paraquat was also the most widely used chemical suicidal agent in Trinidad (Hutchinson et aI., 1991) and Surinam (Perriens et aI., 1989). In Costa Rica, Wesseling et al. (1993) examined records of the Forensic Medical Department which showed that over the 7 year period from 1980 to 1986 a total of 169 fatalities had occurred from paraquat poisoning. The pathologists had classified the overwhelming majority as suicide-related, although the authors suggested that misclassification occurred in some cases. However, a detailed examination of case records of the Forensic Medical Department between 1990 and 1992 showed that 74 out of 76 paraquat related fatalities were due to suicide from oral ingestion, with 2 fatalities occurring from accidental ingestion (Vargas and Sabapathy, 1995). Government statistics for 1995 and 1996 showed that 62 out of a total of 72 pesticide related fatalities (no compound mentioned) were due to suicide and 2 due to homicide, 8 fatalities were classified as nonoccupational, and there were no occupationally related fatalities (Ministerio de Salud, 1997). Paraquat poisoning is uncommon in the United States the world's largest market for paraquat-containing products. A 10 year survey of calls to U.S. poison centers showed that paraquat (and diquat)-related enquiries accounted for only around 0.01 % of the total (Hall, 1995). Most cases showed either no or minor symptoms, with less than two fatalities occurring annually, almost all of them related to suicides. Data on mortality from paraquat poisoning are difficult to compare because of differences in circumstances, treatment, and reporting systems. In a collection of data from 14 publications compiled by the International Programme on Chemical Safety (lPCS, 1984), mortality ranged from 36% to 100%, with an overall mortality of 48% (446 of 925 cases). A difference in mortality between ingestion of the liquid concentrate (20% paraquat ion) and a granular product (2.5% paraquat, 2.5% diquat) has been described by some authors. Park et al. (1975) found that the fatality rate was 15 of 23 (65%) in patients who had ingested liquid concentrate and 3 of 8 (38%) in patients ingesting the granular product. Fitzgerald and Barniville (1978) reported no deaths in 14 patients ingesting the granular product compared to a mortality of74% in 118 cases of ingestion of the liquid concentrate. In the series published by Howard (1979a) there were 36 deaths from 41 cases (88%) where liquid concentrate was ingested, and 5 deaths from 27 cases (19%) involving the granular product. These differences are largely a reflection of the size of dose ingested. While suicidal ingestion of paraquat concentrate accounts for most of the recorded fatalities, the problem of accidental
ingestion prompted the principal manufacturer of paraquat to introduce formulation changes to the liquid concentrate in the late 1970s and early 1980s (Sabapathy, 1995). A blue color was added to prevent confusion with drinks, a stenching agent was introduced to alert users, and an emetic was included. In addition, packaging and labelling was improved to prevent decanting of the product, and education and training efforts were directed in particular toward smallholder farmers in developing countries, where the majority of incidents occurred. The effect of these efforts is believed to have made a significant contribution to the decrease of accidental paraquat ingestion in many countries (Sabapathy, 1995; Wesseling et aI., 1997). Although ingestion is the route of entry into the body for the overwhelming majority of poisoning cases, there are a few reports of systemic effects from inhalation and dermal exposure (localized skin, eye, and upper respiratory effects will be discussed in Section 70.3.4). Inhalation exposure is not a prominent feature in paraquat poisoning cases because of the extremely low (not measurable) vapor pressure of paraquat. Respiratory exposure to paraquat during spray applications is very low because the large droplet size will prevent the material from going beyond the nasal cavity. Concerns about oral exposure to spray droplets as a result of drainage into the oral cavity and swallowing appear unwarranted because the typical spray concentration of paraquat for hand-held spray applications is 0.1-0.2% and would thus require a dose of 1-2 1iters of spray solution directly into the nose and into the oral cavity to achieve a lethal dose (Howard, 1980). It is therefore not surprising that there are no reports in the published literature of deaths arising from inhalation exposure. A review of 30 cases of presumed inhalation exposure found no evidence for systemic poisoning (Vlachos and Kontoes, 1987). Where paraquat was measured it was undetectable or at the limit of detection. Patients were either asymptomatic or had nonspecific symptoms such as headache, nausea, or feeling unwell. Two patients described nosebleeds. In two patients who presented with cough and fever, pneumonia was established as clinical diagnosis. A recent review (Gamier, 1995) concluded that there was only one convincing reported case of possible systemic poisoning following inhalational exposure to paraquat and signs of toxicity were very mild and the patient made a full recovery (Fitzgera1d et aI., 1978a). In this case, a 43 year old market gardener sprayed a "stronger than usual" solution (no details of spray concentration available) in a greenhouse and complained of a burning sensation in throat and mouth and weakness. There was biochemical evidence of mild renal failure, but liver function tests and chest x-ray were normal. Paraquat tested positive in urine. Renal function parameters returned to normal within 10 days after exposure. It has already been mentioned that paraquat absorption across intact human skin is extremely low both in vitro (Walker et al., 1983) and in vivo (Wester et al., 1984). In 15 cases of single exposures of the skin and eyes during work with paraquat solutions only localized lesions (dermatitis, vesicles, bums, conjunctivitis) were found (Hoffer and Taitelmann, 1989). Paraquat was undetectable in plasma except for three
70.3 Toxicity to Humans
cases where it was at the limit of detection. There were no manifestations of systemic toxicity. A small number of case reports describe systemic paraquat poisoning and fatalities from dermal exposure. In six cases there was deliberate or accidental application of paraquat concentrate to the skin, usually in the unfortunate mistaken belief that it could act against parasitic disease (Binns, 1976; Gamier et aI., 1994; Ongom et aI., 1974; Tungsangaet aI., 1983; Wohlfahrt, 1982 (2 cases». Three cases (Okonek et aI., 1983; Waight, 1979; Wesseling et aI., 1997) involved widespread accidental contamination of the lower abdomen and legs with the 20% concentrate. In two cases (Jaros et at., 1978; Levin et at., 1979) it was evident that a far too concentrated paraquat dilution (28 gl1; 2.8% and 40 gl1; 4%, respectively) was applied combined with faulty leaking spray equipment and lack of skin decontamination. In a further case (Athanaselis et aI., 1983) it is explicitly claimed that a correct dilution of 0.5% paraquat was used (the maximum recommended rate for knapsack). However, subsequent investigation (Hart, 1984) led to the conclusion that, in fact, a more concentrated paraquat solution, probably in excess of 1.5%, was used. In one case (Fitzgerald et al., 1978a) the combination of paraquat exposure and pre-existing skin disease caused the death of the person involved, although very few details are given. Another case (Gamier et at., 1994) involved the application of multiple herbicidal mixtures, including paraquat, over several days by a man with a history of psoriasis. This man suffered a febrile lung disease but made a complete recovery. Four cases involving prolonged skin contact with "diluted" paraquat without pre-existing skin lesions should be mentioned. The two cases described by Wohlfahrt (1982) give very few date which would be useful in this context. In the third case (Papiris et at., 1995), a farmer was exposed for 5-6 h to diluted paraquat from a leaking sprayer which caused burning, blisters, and erosions in his scrotal area. This patient survived after hospital treatment. In the fourth case (Wesseling et aI., 1997), a plantation worker experienced chemical burns on his back, scrotum, and inner parts of both thighs after spraying paraquat with a leaking knapsack sprayer for three consecutive days. He subsequently died from interstitial fibrosis of the lung. Thus, there is no indication that paraquat has caused fatal poisoning through skin contact in normal occupational use. The few cases described in the literature occurred as a result of a combination of factors such as misuse (wrong dilution), preexisting extensive skin disease, faulty equipment, prolonged extensive skin contact, and disregard of safety procedures (no decontamination following significant exposure). 70.3.3 USE EXPERIENCE
Exposure to paraquat under actual field conditions has been assessed in studies with hand-held (knapsack), vehicle mounted, and aerial applications. Dermal exposure was measured either in patches placed on different body regions or, more recently, using whole body exposure assessments. Inhalation exposure
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(including oral exposure) was determined using personal air sampling and the air concentration of different particle sizes was measured. Internal dose was assessed using biological monitoring, for which paraquat is an ideal candidate: it is not metabolized, it is rapidly and completely excreted via the kidneys, it is stable in urine, and there are sensitive analytical techniques available. The data from these studies are summarized in Table 70.5. There is an enormous variation in dermal exposure evident in the studies found in the literature. This is not surprising given the differences in spray strength, volume applied, application technique, environmental conditions, use of personal protective equipment, and differences in study design. Nevertheless, some patterns emerge across the variety of study conditions encountered. It is evident that skin exposure represents by far the most significant route of exposure for paraquat. For handheld applications, total dermal exposure was more than an order of magnitude higher than exposure to uncovered body parts (Chester and Woollen, 1981; Van Wendel de Joode et aI., 1996). A similar difference was seen for vehicle-mounted spray applications (Staiff et aI., 1975; Wojeck et at., 1983). The lowest dermal exposure was seen for pilots applying paraquat (Chester and Ward, 1984), whereas the total dermal exposure of Baggers is comparable to exposure of uncovered body parts in other spray applications. Inhalation exposure was approximately three orders of magnitude lower than skin exposure (Chester and Ward, 1984; Chester and Woollen, 1981; Singmaster and Liu, 1998; Staiff et aI., 1975; Van Wendel de Joode et aI., 1996; Wojeck et aI., 1983). Paraquat proved to be below the limit of detection in most samples. Furthermore, the inhalation potential of respirable droplets was found to be negligible since no respirable paraquat could be measured in the breathing zone of exposed workers (Chester and Ward, 1984). The most recent study (Singmaster and Liu, 1998) showed that even under difficult spraying conditions (heavy exertion while spraying on hillsides) paraquat was below the limit of detection. Paraquat is an ideal candidate for biological monitoring because it is excreted unchanged in urine, where it is comparatively stable. Most of the worker exposure studies mentioned above included measurement of paraquat in urine. Overall, the paraquat concentration in urine was low, with the majority of samples being below the limit of detection. None of the samples contained paraquat at levels which would be indicative of a risk of poisoning (see below). Topical effects from contact with paraquat during spray operations can occur due to a delayed caustic action of paraquat as a result of poor working practice and hygiene (Howard, 1980). Discoloration (white bands), paronychia, and partial or complete loss of nails has been described following contact with concentrated (Samman and Johnston, 1969) and prolonged exposure to diluted paraquat solutions (Hearn and Keir, 1971). Upon cessation of exposure, normal nail growth resumes. Irritant dermatitis, burns, and blistering can occur from skin exposure to paraquat concentrate or as a result of prolonged skin contact with contaminated clothing or from leaking spray
1580
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Paraquat
Table 70.5 Worker Exposure and Absorption of Paraquat Spray
Oennal
Inhalation
Application
dilution
exposure
exposure
level
(mg/h)
(mg/h)
(mg/l)
Reference
Country
method
(%w/v)
Swan, 1969
Malaysia
Hand held
0.05
Urine
<0.01 -0.32
Hogarty, 1976
Ireland
Hand held
Staiff et aI., 1975
USA
Vehicle
<0.003
NO
0-0.002
<0.02
0.0I-O.57a
<0.001
<0.02
<0.01-12a
0-0.005
0.1
0.01-3.4a
Hand held
0.2
Hand held
0.1-0.2
mounted Chester and Woollen, 1981
Malaysia
12-170b Wojeck et aI., 1983
USA
Vehicle
0.05-0.1
mounted Chester and Ward, 1984 Chester et aI., 1993
USA Sri
Aerial
0.3
7.0-42a 12-169b
0-0.07
0.1-2.4b,c
O-O.047 c
0.05-0.26b ,d
0-0.06d
Hand held
0.03-0.04
0.94-2.71 b ,e
Hand held
0.1-0.2
0.2-5.7a
<0.05 -0.76 <0.02 -0.03
<0.03
Lanka Van Wendel de Joode et aI., 1996
Costa
0-0.043
Rica Singmaster and Liu, 1998
Puerto
<0.03 -0.24
Hand held
0.1
<0.007
Rico NO = not detected. to uncovered skin. bTotal dennal exposure. C Aerial-flagger. d Aerial-pilot. eMg/g paraquat sprayed. fExtrapolation from indirect measurement using copper as marker.
a Exposure
equipment (Swan, 1969; Van Wendel de Joode et aI., 1996). Epistaxis has been described (Swan, 1969; Van Wendel de Joode et aI., 1996), most likely from breathing in spray mist or contact with contaminated fingers. No serious or long-term effects have been described. There are a number of case reports of eye damage resulting from splashes with paraquat concentrate (Cant and Lewis, 1968; Deveckova and Mydlik, 1980; Joyce, 1969; Peyresblanques, 1969; Watanabe et aI., 1979). Apart from eye irritation and blepharitis, more serious, delayed ocular damage may occur such as destruction of the bulbar and tarsal conjunctiva and erosion of the corneal epithelium. Anterior uveitis has also been noted. Progressive keratitis and decreased visual acuity may occur and persist for several weeks. However, complete restoration of vision is normal. Attempts have been made to establish the frequency of top ical effects from paraquat exposure, particularly for hand-held applications in developing countries. Surveys have been carried out interviewing 400 smallholder farmers using paraquat in Malaysia (Whitaker, 1989a), 365 smallholders in Central America (Whitaker, 1989b), and 732 smallholders in Thailand (Whitaker et aI., 1993). These surveys showed that, in general, farmers were aware of the potentially fatal consequences of swallowing small quantities of the concentrate. Spray practices
and standards of personal hygiene were generally adequate, although the wider use of gloves and eye protection when handling the concentrate needed to be encouraged. In all three surveys, approximately 10% of respondents had experienced health effects attributed to the use of paraquat. These were predominantly skin irritation (mainly on hands and feet), nausea and headaches associated with the smell of the product (due to the added stenching agent), and, to a lesser extent, eye irritation, nail damage, and epistaxis. Ramasamy and Nursiah (1988) interviewed 1219 Malaysian estate workers, rice farmers, vegetable growers, and smallholders about health effects from pesticide use. They found that exposure to organophosphorous insecticides was associated with giddiness and nausea, whereas the main effects associated with paraquat exposure were eye irritation, nail damage, and nasal bleeding. However, their survey did not establish cause effect relationships with exposure to specific products. Only three cases of hospitalization were described among their study population. The State of California has probably the most comprehensive surveillance system of pesticide-related illness in the world. Between 1971 and 1985 a total of 231 cases of ill health attributed to paraquat were notified to the Worker Health and Safety Branch, California Department of Food and Agri-
70.3 Toxicity to Humans culture (Weinbaum et aI., 1995). Of these, 38.5% were listed as systemic effects (mainly dizziness, nausea, lightheadedness, headache, chest pain, vomiting, and tiredness), 32% were eye effects (burning, itching, redness), 26% were skin effects (rash and irritation, itching) and 3.5% were local respiratory irritant effects (epistaxis, sore throat). There were no cases of pulmonary fibrosis. Analysis of data from 1981 to 1985 showed that the overall incidence of illness was low at 0.6 per 1000 paraquat applications. Detailed medical surveys have been carried out to determine whether the long term exposure to paraquat leads to chronic health effects in workers and spray applicators. Swan (1969) found no abnormalities in chest radiographs of groups of Malaysian rubber plantation workers during paraquat applications over several weeks. Howard (1979b) studied two groups of paraquat formulation workers in the United Kingdom and Malaysia. Mean exposure duration for the UK workers was 5 years, and 2.3 years for the Malaysian workers. A history of skin rashes was found in half of the Malaysian workers, but not in the UK workers where the most common finding was epistaxis and nail damage. Eye irritation was more common in the Malaysian than in the UK workers. There was no evidence of any longterm or permanent skin or eye damage. The most comprehensive medical surveys in paraquatexposed spray operators were carried out in Malaysia (Howard et aI., 1981) and Sri Lanka (Senawayake et aI., 1993). In both studies there were detailed clinical examinations, lung function measurements (including CO diffusion capacity), hematological and biochemical investigations, and, in the Sri Lankan study, a chest radiograph was taken. In the Malaysian survey, 27 paraquat spraymen (mean spraying time 5.3 years; mean individual annual quantity of paraquat handled 67.2 kg as paraquat ion) were compared with two control groups comprising 24 general plantation workers and 23 latex factory workers, respectively. In the Sri Lankan survey, 85 paraquat spraymen (mean spraying time 12 years) were compared with two groups of 76 factory workers and 79 general workers, respectively. In both studies there were no clinically significant differences in any of the parameters studies; in particular, the results of the lung function tests showed similar results for exposed and control groups. It was concluded that the long term spraying of paraquat was not associated with any measurable adverse health effects. A recently published study was carried out in Nicaragua (Castro-Gutierrez et aI., 1997), although the investigation dates back to 1987/88. A population of 134 spray workers with at least 2 years spraying experience with paraquat from 15 banana plantations was interviewed, 63 out of which had not experienced skin irritation, and 71 who had a history of skin rash or bum (used as a surrogate measure of intensity of exposure). A questionnaire was used to check for symptoms of respiratory illness and Forced Vital Capacity (FVC) and Forced Expiratory Volume in 1 second (FEV 1) were measured. The results were compared with a control population of 152 unexposed workers. There was a difference in male: female ratio between the exposed and unexposed groups (100:34 and 88:64, respectively). Paraquat-exposed workers gave a significantly more frequent
1581
history of Grade 3 dyspnea, but not Grade 1 or 2 dyspnea. There was no difference in the occurrence of chronic bronchitis, and episodic dyspnea with wheezing was more frequent in the group with topical effects only. However, there were no differences between exposed and control workers with regard to restrictive (FVC < 80% of predicted value) or obstructive (FEV1:FVC < 70% of predicted value) spirometry parameters. In fact, the lowest incidence of restrictive changes was found in the "intensive exposure" group. 70.3.4 ATYPICAL CASES OF VARIOUS ORIGINS
In a case described by Newhouse et al. (1978), a farner's wife had been spraying paraquat in an orchard for many days. This case is unique in that her complaints started with scratches on arms and legs which proved nonhealing over four weeks. She was then hospitalized for two weeks and discharged without diagnosis. Two and a half weeks later she was readmitted to the hospital because of increased dyspnoea and wheeziness. She was diagnosed as suffering from systemic arteritis and died 12 days after final admission, some 8 weeks after initial exposure. Although the paper links her disease to paraquat exposure, it is doubtful if paraquat was the cause. First, at no time was paraquat measured in blood or urine. Second, the time from exposure to her death was more than 8 weeks, which is highly unusual for paraquat poisoning. Third, she had a clinical diagnosis (systemic arteritis) which did not include any reference to paraquat poisoning. George and Hedworth-Whitty (1980) attributed a case of nonfatal lung disease to the inhalation of nebulised paraquat. A 64 year old woman noticed spray mist drifting into her garden from a spraying operation in an adjacent field. After some 10 minutes she noticed a chest tightening, and over the next week she became gradually more breathless. She was initially treated with a short course of steroids without much effect. Pulmonary function evaluation some two months later showed severe restriction, but there were no abnormalities in the chest radiograph. She was kept on systemic steroids and her lung function had markedly improved some 7 months after the original incident. Hart (1980) commented that the diagnosis of paraquat-induced lung injury was doubtful. The woman had a history of allergic rhinitis and chronic sinusitis. No previous lung function recording was available and no transfer factor was measured at the time of assessment. The chest radiograph was clear and the description of exposure did not provide convincing arguments for a significant inhalation exposure. In the case described by Katopodis et al. (1993), a 31 year old woman was admitted 4 days after ingestion of 2 g paraquat. The urine test for paraquat was still positive, but her plasma concentration was only 10 I-lg/l. Charcoal hemoperfusion was carried out over the next 5 days. Paraquat levels became undetectable in plasma on day 6 and in urine on day 8. The patient survived without evidence of pulmonary involvement. The authors attributed the favorable outcome to the hemoperfusion
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Paraquat
therapy even at such a late stage after ingestion. However, the low paraquat plasma concentration at the time of admission would have suggested a good chance of survival anyway (see below and Table 70.5). Ragoucy-Sengler and Pileire (1996b) reported a case of paraquat poisoning in an HIV positive patient. Indices of severity of the poisoning suggested a survival probability of 30% on admission, and 3% after 72 h. The clinical course included acute renal failure and severe hypoxia; however, pulmonary fibrosis did not develop. The patient was discharged with normal pulmonary function 18 days after admission. The authors suggested that the immune deficiency on the basis of the patient's HIV infection may have prevented the development of pulmonary fibrosis. In a case described by Ernouf et al. (1998) a 47 year old man, while under the influence of alcohol, ingested paraquat which had been decanted into an unmarked a red wine bottle. The patient was a chronic alcoholic. He was admitted to hospital within 3 h and treated with gastric elimination and antioxidant therapy. Evolution of plasma paraquat concentrations pointed toward a prognosis of delayed death from pulmonary fibrosis. However, the patient died on the fourth day after admission from persistent hemodynamic shock and hypoxaemia. The authors speculated that co-ingestion of ethanol may have enhanced the toxicity of paraquat through increased absorption from the gastrointestinal tract and/or decreased renal clearance. However, it has also been suggested that alcoholism may have a protective effect against paraquat toxicity on the basis of increased synthesis of superoxide dismutase (Ragoucy-Sengler et aI., 1991). Methemoglobinemia was described in a patient who ingested "Gramonol," formulation containing 100 g/l paraquat and 140 g/l monolinuron (Ng et aI., 1982). The authors speculated that the superoxide anion and hydrogen peroxide generated by paraquat could oxidize hemoglobin to methemoglobin. However, in response, Proudfoot (1982) pointed out that monolinuron, along with other substituted urea herbicides, is metabolized to aniline derivatives which are well known methemoglinemia and hemolysis causing agents. Furthermore, administration of monolinuron alone had produced methemoglobinemia in experimental animals. Instead of a new feature of paraquat poisoning, it appeared therefore that Ng et al. had reported the first human case of monolinuron toxicity. Since then, a further case of paraquat-monolinuron poisoning has been described (Casey et aI., 1994) in which the severe methenoglobinemia (52%) was successfully treated with methylene blue. However, the patient died after 10 days from the consequences of paraquat poisoning. In 1975 the Government of Mexico began an aerial spraying program, financed by the United States, to destroy marijuana fields with paraquat. In 1978 analyses showed that 21 % of 61 marijuana samples confiscated in California, Arizona, and Texas contained paraquat residues between 3 and > 2000 ppm (Turner et aI., 1978). Further work demonstrated that, nationally, 0.63% of over 100,000 kg marijuana seized contained detectable paraquat levels with a median of 52 ppm (Liddle et
aI., 1980). Over 70% of the contaminated samples were found in the Southwest region of the United States, originating almost exclusively from Mexico. Combustion testing suggested that around 0.2% of the paraquat residue would pass unchanged into marijuana smoke (Brine et aI., 1981). On the basis of a worst case epidemiological risk assessment it was suggested that some marijuana smokers in the Southwest region might have been at risk of health effects from paraquat inhalation (Landrigan et aI., 1983). However, no clinical cases were identified during these studies. A possible association between paraquat exposure and the development of Parkinson's disease has been the subject of much speculation. The reason for this is, as previously discussed, the apparent structural similarity between paraquat and the synthetic pyridine MPTP which produced severe neuropathies in several dozen drug users in southern California (Langston et aI., 1983; Lewin, 1984). The first epidemiological work to draw attention to a possible role of pesticides in Parkinson's disease was published by Barbeau et al. (1986), who showed that the regional incidence of the illness in Canada was nonuniform and correlated with a genetically determined enzyme deficiency. While there was certainly a strong correlation between disease incidence and pesticide use, such a correlation was also found for industrial areas and wood processing regions. Since then a number of case-control studies have been published, with varying methodologies and conflicting results. Some studies suggested that the use of herbicides was significantly associated with the development of Parkinson's disease (Golbe et aI., 1990; Ho et aI., 1989; Semchuk et aI., 1991); in two studies this was specifically linked to paraquat exposure (Hertzman et aI., 1990; Liou et aI., 1997). Others have found no such association (Koller et aI., 1990; Ohlson and Hogstedt, 1981; Tanner et al., 1989; Tanner et aI., 1990; Zayed et aI., 1990). Structure-activity relationships suggest that, despite their apparent similarity, paraquat and MPTP are two very different chemicals (Koller, 1986; see above). Barbeau's hypothesis that Parkinson patients may be more likely to have a specific hydroxylation defect in the P450 enzyme system which might inhibit their ability to metabolize toxins (Barbeau et aI., 1985) does not apply to paraquat because it is not metabolized in mammals. Furthermore, none of the health surveys of paraquat-exposed workers (see above) has revealed any neurological deficits, let alone Parkinson's disease. The strongest evidence against paraquat as a causative factor in Parkinson's disease, however, comes from the many published case reports of paraquat poisoning. There is no evidence of a specific effect of paraquat on the nervous system, nor have neurological sequelae been noticed in survivors of paraquat poisoning (Vieregge et aI., 1988). Zilker et al. (1988) carried out detailed neurological follow-up examinations in four survivors of paraquat poisoning (latency period between ingestion and follow-up 5-10 years) and three patients who had had skin contact with paraquat. It was possible to exclude Parkinsonism in all patients. One patient exhibited tardive dyskinesia most likely due to long
70.3 Toxicity to Humans term therapy with neuroleptic drugs. The authors concluded that acute paraquat exposure does not lead to Parkinson's disease. 70.3.5 CLINICAL FINDINGS AND DOSAGE RESPONSE Information on the clinical course of paraquat poisoning is mainly based on case reports of patients who swallowed paraquat concentrate with suicidal intent. However, the systemic toxic effects are similar regardless of the route of absorption. Paraquat causes nausea which may be prolonged especially following ingestion of emeticized formulations (Meredith and Vale, 1987), as well as vomiting and diarrhea as a result of its local irritant effect on the gastrointestinal tract. Patients may develop a burning sensation, soreness, and pain in the mouth, throat, chest, and abdomen (Vale et al., 1987). Ulceration in the mourn and throat, an inability to swallow saliva, dysphagia, and aphonia are common. The presence of buccopharyngeallesions has no prognostic value (Bismuth et aI., 1995), in contrast to oesophageal and, in particular, gastric ulcerations which indicate a poor prognosis (Bismuth et al., 1982). Prominent pharyngeal membranes ("pseudodiphtheria") have been reported (Stephens et aI., 1981) and perforation of the oesophagus may result in mediastinitis, surgical emphysema, and pneumothorax (Ackrill et aI., 1978). The further clinical course is dependent on the amount of paraquat absorbed into the body (usually following ingestion). Attempts have been made to quantify the toxic dose from estimates based on the information given by patients. Although such estimates are often unreliable, a consensus has emerged which is based on experience with many patients. This has allowed the identification of three degrees of intoxication which are summarized below (for further details see Vale et aI., 1987, and Bismuth et aI., 1995).
1583
necrosis may occur. Both these lesions are fully reversible. Delayed development of pulmonary fibrosis is responsible for the generally poor prognosis in this group. Clinically and radiologically this appears around 7 days after ingestion, but subtle abnormalities are present much earlier, such as a decreased diffusing capacity. The x-ray often shows patchy infiltration which may progress to opacification in one or both lungs. In thin section computerized tomography, the most common pattern on initial scans is ground-glass attenuation, followed by consolidation with bronchiectasis (Lee et aI., 1995). In most cases, pulmonary fibrosis leads to development of refractory hypoxaemia, resulting in death over a period of 5 days to several weeks. 70.3.5.3 Fulminant or Hyperacute Poisoning
In cases of massive ingestion (usually well above 40-55 mg/kg paraquat ion) patients survive less than 4 days and die in cardiogenic shock and multi organ failure. Apart from renal and hepatic failure, alveolitis and noncardiogenic pulmonary oedema are observed. Other organ systems (adrenal glands, pancreas, heart) are affected and mortality in this group has been suggested to approach 100%. While this categorization reflects experience with a large number of cases, it has to be emphasized that there are a significant number of cases reported in the literature where there was survival following the ingestion of alleged doses well above of what is usually considered to be fatal. Table 70.6 shows that there are 52 case reports where a dose apparently in excess of 55 mg/kg has been survived. While inaccuracies in estimating the dose may have led to exaggeration of the dose in some cases, this appears unlikely in many others. Talbot et al. (1988b) reported a series of nine cases of suicidal paraquat poisoning in pregnant women. In the cases where the outcome was known, one fetus died probably unrelated to paraquat, three died in utero or after delivery but associated with respiratory distress in the mothers, two died in utero (one 70.3.5.1 Mild or Subacute Poisoning mother survived and subsequently had a normal pregnancy with The smallest fatal dose has been quoted as 16.7 mg/kg (Stevens no evidence of teratogenicity from the previous paraquat inand Sumner, 1991). However, the original reference (FAOIWHO, toxication), one and fetus was aborted. Previously, Fennelly 1973) makes clear that this value is erroneously low, since et al. (1968) had reported the case of a woman who was 28 the formulation ("Weedol") also contained an equal amount weeks pregnant and died 20 days after paraquat ingestion. Upon of diquat, so that the total bipyridyl ingestion was approxi- autopsy the fetus showed no abnormalities. A 20 week pregmately 35 mg/kg. This is in line with clinical experience which nant patient survived the ingestion of a small dose of paraquat shows that ingestion of less than 20-30 mg paraquat ion/kg has and subsequently delivered a normal child (Musson and Porter, rarely serious consequences. Patients are either asymptomatic 1982). There are now sufficient case reports in the literature to or develop nausea and vomiting. Renal and hepatic lesions are demonstrate that the development of pulmonary lesions is not minimal or absent. An initial decrease of the diffusing capacity inevitably fatal. Fitzgerald et al. (1979b) examined 13 survivors may be apparent in lung function measurements, but full recovof acute paraquat poisoning after a minimum of 1 year. In ery is normal. two children, no clinical, functional, or radiological abnormalities were seen. Of the 11 adults, 5 nonsmokers also showed 70.3.5.2 Moderate to Severe Acute Poisoning no evidence of pulmonary disease. Four smokers were conThis occurs following ingestion of more than 20-30, but less sidered normal on clinical and radiological criteria, but had a than 40 to 50 mg/kg. Apart from the localized lesions described mild deficit in pulmonary function which could reasonably be above, patients in this group develop renal failure, usually be- attributed to smoking. Two patients had pronounced arterial hytween the second and fifth day after ingestion. Hepatocellular poxemia, both having had pre-existing pulmonary disease. In
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CHAPTER 70
Paraquat
Table 70.6 Summary of Case Reports with Doses above 55 mg/kg Taken by Survivors of Paraquat Poisoning by Ingestion Calculated
Body
Age
Number
weight
Range
Dose statedb
of cases
(kg)"
(years)
15-20 ml
13
25-70
3-75
ingested dose (mg/kg)a
55-75
References Addo et aI., 1984
(9 cases)
If[ et
1-2 mouthful
Lloyd 1969 (cited in Cavalli, 1977)
one sachet
Mahieu et aI., 1977
(2 cases)
Ming et aI., 1980
al., 1971
Mirchev, 1977 Taki et aI., 1996 Talbot et aI., 1988a 76-100
10-40 ml
18
25-70
3-65
Addo et aI., 1984
(14 cases)
Douze et aI., 1977
1-2 mouthful
Malone et aI., 1971
(4 cases)
McKean, 1968 Ragoucy-Sengler et aI., 1991 Shahar et aI., 1980 Tabei et aI., 1982 Taki et aI., 1996 Thomas et aI., 1977 Tsatsakis et aI., 1996
101-200
40->50 ml
9
17-59
70
(7 cases)
Addo et aI., 1984 Douzeetal.,1974,
3-4 mouthful
1977
(2 cases)
Florkowski et aI., 1992 Grundies et aI., 1971 Lheureux et aI., 1995 Okonek et aI., 1980
>201
50-150 ml
12
(8 cases)
40-70
10-50
Addo et aI., 1984 Douze et aI., 1974,
3-4 mouthful
1977
(2 cases)
Malone et aI., 1971
one glass or cup
Okonek et aI., 1979
(2 cases)
1980 Tabei et aI., 1982 Tsatsakis et aI., 1996
a All
doses expressed as paraquat ion. bVolumes (ml) refer to the 20% liquid concentrate. A volume of 17.5 ml has been used for "a mouthful." "Sachet" refers to a granular formulation containing 2.5% paraquat and 2.5 g diquat. cWhere the body weight was not explicitly stated, the following assumption were used: 3-6 years, 25 kg; 7-11 years, 40 kg; 12-16 years, 50 kg; 17 years and above, 70 kg.
one of these two patients new and persistent infiltrates were seen in radiography which could be ascribed to paraquat lung damage. Hudson et al. (1991) described persistent radiological changes in three survivors of paraquat poisoning. In one case the patient died a year after her first intoxication from a second massive dose of paraquat. Upon autopsy pulmonary changes from the first as well as the second intoxication were present. Lin et al. (1995) studied 16 survivors of moderate to severe paraquat poisoning after 3 months. Detailed lung function
showed significant improvements over time. This was confirmed by improvements in chest radiographs which showed some residual interstitial fibrosis, especially in the lower lobes. Bismuth et al. (1995) reported five cases, all of which had developed a restrictive pulmonary lesion, but who survived. Two patients were followed up for 4 and 10 years, respectively. In the first patient there was an obstructive component to his pulmonary insufficiency (from smoking) which persisted over time. However, the restrictive component gradually improved
70.3 Toxicity to Humans Table 70.7 Predictive Plasma Paraquat Concentrations beyond 24 Hours Separating Surviving and Nonsurviving Patients (from Scherrrnann, 1995) Plasma paraquat Time (h) 24
concentration (ng/m1) 100
48
86
72
74
96
63
120
54
144
48
168
42
192
37
216
32
240
27
264
23.5
288
20
312
18
over several years, with eventual return to near baseline state. In the second patient (a 13 year old adolescent at the time of intoxication) pulmonary function tests were completely normal 10 years after the poisoning. He had also been able to actively participate in sports. The measurement of paraquat plasma concentration has proved to be a reliable indicator of the prognosis of the intoxication. Levitt (1979) was the first to demonstrate a relationship between plasma concentration of paraquat, the estimated time after ingestion, and the eventual outcome. Based on results from 79 patients with a reasonably well established time of ingestion, Proudfoot et al. (1979) found that those patients whose plasma paraquat concentration did not exceed 2.0, 0.6, 0.3, 0.16, and 0.1 mg/l at 4, 6, 10, 16, and 24 h after ingestion survived. This semilogarithmic plot has become known as the predictive line, or "Proudfoot's curve." Because of the rapidly decreasing plasma concentration in the first few hours following ingestion no accurate prognosis could be given prior to 4 h. The authors emphasized that the line to separate survivors and non survivors was meant to be an approximate guide, and the main use should be to help clinicians in deciding which patients needed urgent aggressive treatment. Subsequently, several other methods have been described to establish the prognosis from plasma paraquat concentrations. None of those methods have been found to invalidate the original estimate by Proudfoot et aI., but they have added other dimensions which may be of help to clinicians. Scherrmann et al. (1987) used data from 30 patients to extrapolate the predictive line beyond 24 h up to 15 days after intoxication; dlis was later modified (Scherrmann, 1995) with data from a total of 52 patients (Table 70.7). The same authors evaluated the relationship between early urine concentrations and clinical prognosis. They also attempted to correlate urine results obtained by radioimmunoassay with those given by the simple colorimetric dithionite test. Data from 75 patients showed a wide variation in urine concentrations within
1585
24 h of ingestion. All 17 patients with concentrations of less than 1 J.1.g/ml survived, whereas 51 out of 58 patients with urine paraquat concentrations of more than 1 J.1.g/ml died. No color was observed in the dithionite test at paraquat concentrations below 0.5 J.1.g/ml (Scherrmann et aI., 1987; Scherrmann, 1995). Using a sample size of 219 patients, Hart et al. (1984) were able to calculate the probability of survival of the patient from the initial paraquat plasma concentration (Fig. 70.7). It was noted that the line denoting a 50% probability of survival correlated well with Proudfoot's curve. Sawada et al. (1988) categorized their patients into three groups: survivors (n = 10), nonsurvivors who died from respiratory failure (n = 9), and nonsurvivors who died from circulatory failure (n = 11). They calculated a severity index of paraquat poisoning (SIPP) from time to treatment since ingestion of paraquat multiplied by the serum level at admission (J.1.g/ml). A boundary SIPP of 10 separated survival from death by either cause, whereas a SIPP of 50 separated deaths from respiratory failure and deaths from circulatory failure. Using data from 128 patients, Ikebuchi et al. (1993) separated survivors and fatal cases by multivariate analysis and established a discriminate function D. Their toxicological index of paraquat (TIP) could then be divided into three types. TIP 1 is characterized by D > 0.1 (100% survival probability). TIP 2 has the characteristic -0.1 < D < 0.1 and here urgent treatment may influence the outcome. In TIP 3 the discriminate function D < -0.1, and the probability of a fatal outcome is 100%. All these methods depend on the availability of paraquat analysis, and this is often not the case, or at least not in a timely fashion. Investigators have therefore attempted to predict the outcome of the intoxication using biological indices rather than plasma paraquat concentrations. Suzuki et al. (1989) measured the respiratory index (RI) from blood gas analysis and used it as an index of lung oxygenation in 51 patients. Progressive deterioration of the RI above 1.5 was found in 43 nonsurvivors, whereas the RI remained below 1.5 in the 8 survivors. Furthermore, the time taken from ingestion for the RI to exceed 1.5 was found to be a good indicator for predicting the survival period in fatal cases. The major weakness of this method is that it cannot predict the outcome at the point of first contact with the patient, unlike the methods relying on plasma paraquat analysis. Also, conditions which may influence the RI such as pneumothorax, cardio-pu1monary rescuscitation, septic shock, pulmonary edema, and pneumonia limit the usefulness of this method. On the other hand, it can be used at any time after the intoxication, and it is independent from an estimate of time of ingestion. Yamaguchi et al. (1990) reviewed the medical records of 160 patients who had ingested paraquat and calculated an equation derived from serum creatinine and potassium concentrations and arterial blood bicarbonate level. When plotted against time of ingestion they were able to estimate the probability of survival in three categories (90%, 38%, and 3%). Most recently, a different biological index using creatinine measurement from 18 patients has been proposed by Ragoucy-Sengler and Pileire (1996a). They found that the time evolution of blood creatinine in intoxicated patients was linear during the first 24 h
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5.5 Percentages denote the probability of survival
5.0 4.0 3.0 2.0 1.0
-----
.~----
900/"",--
o o
4
8
12
16
20
24
28
Hours after swallowing Figure 70.7 Relationship between the concentration of paraquat in the plasma and the survival of the patient. From (Hart et aI., 1984), reproduced with permission © 1984.
after admission. The rate of increase of creatinine in the patients with fatal outcome was equal to a constant (zero order kinetics). A rate of creatinine increase over 5 h (dCreat/dt) of >3 J.lmol/llh was found in the 12 fatal cases whereas this value remained < 1.26 for the survivors. As with the method of Suzuki et al. (1989) this biological index is independent of an estimate of time elapsed since ingestion. It has the advantage that a prognosis can be established within a few hours after admission of the patient using a standard biochemical analysis. However, it is currently based on data from a relatively small number of patients and will thus require further confirmation from a larger dataset. 70.3.6 LABORATORY FINDINGS If performed early and serially, pulmonary function tests may be of diagnostic value. However, it has been pointed out that any changes seen are not specific for paraquat poisoning, since they may also occur in other clinical conditions such as pneumonia, pulmonary edema, pulmonary thromboembolism, and advanced degrees of the alveolar capillary block syndrome (Cooke et al., 1973). The abnormalities must be interpreted in conjunction with the clinical picture. As mentioned above, pulmonary function tests in patients with moderate to severe paraquat poisoning are likely to be abnormal much earlier than clinical or radiological findings. A decrease in the carbon monoxide diffusing capacity or transfer factor (DLco or TLco) can be noted as early as the first day after intoxication (Baguley et aI., 1983). Beginning between the fifth and sixth
day there may be a restriction of the FEV 1 and the FVC. These changes are followed by a drop in the arterial oxygen tension and an increase in the gradient of alveolar to arterial tension. Finally, there is the development of a functional shunt by which a decreasing fraction of the blood passing through the lungs is oxygenated (Cooke et aI., 1973). In patients who died within 11-14 days, the extent of lipid peroxidation, expressed as malondialdehyde, was higher than in controls or in patients who survived. Massive doses (death in 1-3 days) did not result in increased levels of malondialdehyde (Yasaka et aI., 1981, 1986). Serum superoxide dismutase (SOD) levels were significantly decreased in cases of lethal paraquat poisoning (Nemeth et aI., 1985). Better clinical courses were detected if SOD levels were normal or slightly elevated. Extremely increased levels were measured several times in the terminal state and were interpreted as the consequence of liver cell necrosis and intravascular haemolysis. Other laboratory findings, including those reflecting renal and hepatic failure, are nonspecific. Detailed renal function studies were performed in three cases of paraquat poisoning who developed acute renal failure (Vaziri et aI., 1979). The glomerular filtration rate (estimated by using creatinine clearance) improved for two patients who survived two weeks, illustrating the reversible nature of the renal failure. A mild to moderate transient proteinuria but little albuminuria was observed during the first two weeks after intoxication. Other findings consistent with proximal tubular dysfunction included glucosuria, amino aciduria, and increased fractional excretion of phosphorus, sodium, and uric acid.
70.3 Toxicity to Humans Many case reports have shown a transient rise in liver enzymes such as ALT and AST, reflecting the centrilobular necrosis and cholestasis often seen at autopsy (Vale et al., 1987). Serum protein was decreased in one case (Bullivant, 1966) but increased in another with a large increase in the globulin fractions (Matthew et al., 1968). Peak total serum bilirubin concentration correlated significantly with the alveolar-arterial oxygen difference in a series of 21 patients (Lin et aI., 1995). Normochromic anemia developed rapidly in five cases reported by LautenschIager et al. (1974). This was accompanied by suppression of erythropoietin in the bone marrow but had little effect on other aspects of hematopoesis. The bone marrow had returned to normal in one patient who survived and was re-examined 6 months after the intoxication. In the above mentioned study by Lin et al. (1995), the alveolar-arterial oxygen difference also showed a negative correlation with the initial platelet count. Paraquat analysis in plasma and urine has already been mentioned as the key to diagnosis and prognosis of paraquat poisoning. A simple spot test can be performed with urine or gastric aspirate and is based on the reduction of paraquat cation to a blue radical in the presence of alkali and sodium dithionite (Berry and Grove, 1971; Widdop, 1976). These methods can detect concentrations of paraquat in urine down to 1-2 Il-g/ml and may be made semiquantitative if a range of standards are prepared in control samples. Quantitative methods based on the dithionite reaction with a spectrophotometric endpoint have also been described to determine paraquat in plasma (Jarvie and Stewart, 1979; Knepil, 1977). An improved spot test using extraction with a silica cartridge has allowed lower detection limits between 0.1 and 0.5 Il-g/ml (Woollen and Mahler, 1987). The lower limit of detection for paraquat using spectrophotometry following solid phase extraction was 45 ng/ml (Smith et aI., 1993). Other methods which have been described include a radioimmunoassay with a sensitivity of 6 ng/ml (Levitt, 1979). Gas chromatography and mass spectroscopy have been used (Draffan et aI., 1977), giving a sensitivity of 25 ng/ml. A fluoroimmunoassay achieved a sensitivity of 20 ng/ml (Coxon et al., 1988). Gill et al. (1983) described a high performance liquid chromatography method involving ion-pair extraction on disposable cartridges of octadecyl silica. Most of these methods can be applied to the analysis of plasma, urine, and tissue samples.
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excreted in urine (Bennett et aI., 1976; Davies et aI., 1977). Limited clinical data suggest that having a full stomach may effectively decrease the bioavailability of paraquat (Bismuth et aI., 1982, 1995). In humans, the precise time at which the plasma paraquat concentration peaks is unknown. However, paraquat may be detected in urine as early as 1 h after ingestion (Meredith and Vale, 1987). To judge by the plasma concentration data published by Proudfoot et al. (1979), peak plasma concentrations in humans are certainly attained within 4 h. This is in line with the toxicokinetic analysis of data from 18 patients by Houze et al. (1990), who estimated peak plasma concentrations to occur between 2 and 4 h. However, most patients were admitted to hospital comparatively late, and they could measure peak plasma concentrations in only two case; in both they were seen around 3.5 h after ingestion. 70.3.8 DISTRIBUTION The distribution of paraquat appears to be similar in humans and dogs (Davies et aI., 1977; Van den Bogaerde et al., 1984), suggesting that the three compartment model described by Hawksworth et al. (1981) (see above) in the dog is also applicable to humans. Smith (1987) pointed out that the concentration of paraquat in plasma in human poisoning cases falls rapidly to much lower levels than described in the rat. In their series, Houze et al. (1990) found that the concentration-time curve in 15 adult patients (not hemodialyzed) was best described by a biexponentia1 curve, with the elimination half-lives of the early and late phase being 5 and 84 h, respectively. These patients could be divided into three groups:
1. Patients admitted early and having a rapidly fatal course from cardiovascular collapse showed only monoexponential decreases with a mean half-life of 7 h. However, because of the early death of the patients, evaluation of the late phase was precluded. 2. The second group included patients who were admitted early and survived long enough for an evaluation of the late phase. They showed a biexponential decrease with mean half-lives of7 and 103 h, respectively. 3. In the third group hospital admission was delayed and only late paraquat plasma concentrations could be measured. Accordingly, a monoexponential decrease in plasma paraquat concentrations was observed with a mean half-life of 101 h.
70.3.7 ABSORPTION No adequate data exist on absorption of paraquat in humans. However, Davies (1987) has pointed out that early estimates of an absorption of less than 5% of an ingested dose (Conning et aI., 1969) may be an underestimate. He suggested that absorption kinetics in man may be more similar to those seen in the dog, where a rapid but incomplete paraquat absorption occurs, with peak plasma levels occurring at 75-90 minutes, and almost 40% of the dose absorbed in 6 h, as judged by the amount
Acute renal failure occurred in all but one of the patients. The terminal half-life, however, was very long even in the patient with normal renal function, suggesting that the prolonged elimination phase depends not only on renal function but also on the gradual release of paraquat by extravascular tissue into the blood circulation. In six of their cases with fatal outcome, Houze et al. (1990) also determined tissue paraquat concentrations. High concentrations were found in the lungs, kidneys, heart, and liver and
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CHAPTER 70 Paraquat
much lower concentrations in lipophilic organs such as brain and adipose tissue. The apparent volume of distribution ranged from 1.2 to 1.5 l/kg, compared to 2.75 l/kg in the study by Davies et al. (1977). The mean value of the distribution half-life in humans is greater than that reported from animal studies (see above). Assuming a first-order distribution rate constant and an early half-life of 5 h, paraquat distribution would be achieved within approximately 30 to 40 h (Houze et aI., 1990). The active transport of paraquat into lung tissue in different species, including humans, has been described in detail above. Paraquat accumulation in tissue could be considered as a slow process from a pharmacological point of view, but it is rapid in clinical terms (Bismuth et aI., 1987). In a study of the kinetics of paraquat through the heart-lung block, Baud et al. (1988) showed that concentrations in the radial artery were usually higher than or equal to the corresponding value in the pulmonary artery. Only one patient who was examined approximately 4 h after ingestion showed a pulmonary artery concentration clearly higher than that in the radial artery, providing evidence of pulmonary uptake of paraquat. The arteriovenous difference disappeared approximately 8 h after ingestion followed by inversion of this ratio. This suggests that lethal concentrations of paraquat in the lung may be reached less than 10 h after ingestion. Paraquat crosses the placenta and a case reported by Talbot et al. (1988b) suggests that it is concentrated in the fetus. Following suicidal ingestion of paraquat a premature infant (32 weeks) was delivered by Caesarean section. Both mother and infant died shortly thereafter. Paraquat was measured in maternal blood at 5.6 !-lg/ml and in the infant's blood at 20.6 !-lg/ml. 70.3.9 METABOLISM
As in experimental animals, paraquat is not metabolized in humans but is reduced to an unstable free radical which is then reoxidized to produce a superoxide radical (see above). Paraquat is excreted unchanged in urine. 70.3.10 EXCRETION
As in experimental animals, paraquat elimination is essentially renal via glomerular filtration with an element of tubular secretion (Bismuth et aI., 1988). With normal renal function, clearance of paraquat is greater than creatinine clearance, which enables excretion of high concentrations and large amounts of paraquat within the first hours after ingestion (Davies et al., 1977; Scherrmann et aI., 1983). However, ingestion of large doses of paraquat causes tubular necrosis with a rapid decrease of glomerular filtration and tubular secretion. In four cases described by Houze et al. (1990), renal paraquat clearance was lower than creatinine clearance, even in a patient with apparently normal creatinine clearance. Urinary and plasma elimination half-lives correlated well. Paraquat may be detectable in urine for a long period of time. Beebeejaun et
al. (1971) found paraquat excreted in urine until 26 days after ingestion. In the case of a 14 month old boy, Houze et al. (1990) could detect paraquat in urine for up to 3 months after ingestion, suggesting ongoing release of paraquat from a deep body compartment. Small amounts of paraquat have been recovered in bile samples at postmortem examination, suggesting that a minor enterohepatic cycle may exist in humans (Van Dijck et aI., 1975). As in experimental animals, the amount of paraquat excreted in feces corresponds to 60-70% of the ingested dose in humans. This excretion may be prolonged (Van Dijck et aI., 1975). 70.3.11 PATHOLOGY
Pathological findings upon autopsy in humans fatalities from paraquat poisoning are similar to those seen in experimental animals, in particular the rat (for a detailed review see Smith and Heath, 1976). The lung is the organ showing the most severe changes in paraquat poisoning. Pulmonary pathology has been divided into two phases which correspond with the early and late stages of the clinical signs and symptoms (Smith and Heath,1975). 70.3.11.1 The Destructive Phase
This occurs during the first few days after paraquat poisoning and is rarely seen in human autopsy cases, but it has been described in a case where an early biopsy was performed (Toner et aI., 1970). It is characterized by swelling of the alveolar epithelium which sloughs off and is thought to be related to early development of pulmonary edema with congestion and fibrin exudate (Smith and Heath, 1974a). Death due to this pulmonary pathology is rare. 70.3.11.2 The Proliferative Phase
This phase is usually seen in patients who survive for longer than 1 week. Pulmonary congestion with interstitial and alveolar edema continues, sometimes associated with hemorrhage. There is lymphocytic and other inflammatory cell infiltration and occasional proliferation of cells lining the alveolar wall (Bullivant, 1966). The most specific feature is the presence of large quantities of fibroblastic tissue which is perivascular and peribronchial early on, but later more diffuse (Smith and Heath, 1974b). The pulmonary fibrosis is sometimes associated with an early honeycomb appearance of the lung parenchyma. However, in contrast to a true honeycomb lung the cystic air spaces are dilated respiratory bronchioles and their, walls consist of fibrosed, collapsed alveoli. Renal pathology is common, but it is rarely responsible for the death of the patient. Macroscopically, the kidneys are swollen and soft. There is degeneration or necrosis of proximal tubular cells (Bullivant, 1966; Campbell, 1968) with nuclear loss and cast formation (Parkinson, 1980). Depending on the time after poisoning, there may be signs of regeneration.
70.3 Toxicity to Humans While early studies made little mention of liver damage, Mullick et al. (1981) found evidence of cholestasis, usually localized to the centrilobular region in the majority of their 13 autopsy cases. There was cholangiocellular injury involving the small- and medium-sized bile ducts in portal areas. The authors hypothesized that paraquat injury to the liver is biphasic with an initial hepatocellular injury followed after 2 days by a cholangiocellular phase. Toxic myocarditis is frequently seen in cases with ingestion of larger amounts of paraquat. Parkinson (1980) described a patchy but widespread polymorphonuclear leucocyte infiltration in the presence of normal myocardial fibers. In some cases adrenal cortical necrosis has been described (Nagi, 1969; Reif and Lewinsohn, 1983) in patients who died early after ingestion of paraquat. This lesion was diffuse and involved mainly the zona fasciculata and zona reticularis. FitzgeraId et al. (1977) found adrenal cortical necrosis upon autopsy in 12 of 23 patients. The severity of the lesion appeared doserelated with patients showing complete cortical necrosis after ingestion of higher doses. Brain pathology has been studied in a series of eight patients (Grant et aI., 1980). Changes included generalized edema, hemorrhages (these two findings being the most consistent changes), glial reactions, and meningeal inflammation. The authors suggested that paraquat may damage the cerebral blood vessels. These changes were also seen in a case reported by Hughes (1988) who suggested that, apart from a direct toxicity of paraquat on cerebral blood vessels, the neuronal depletion, myelin breakdown, and astrocytic fibrous gliosis seen were a secondary effect due to prolonged anoxia. 70.3.12 TREATMENT OF POISONING The therapy of paraquat intoxication has focused on three main areas: prevention of absorption from the gastrointestinal tract, enhancement of elimination of paraquat from the body, and therapy directed against the mechanisms of toxicity. In addition, there have been attempts to use lung transplantation as a means to overcome the consequences of paraquat lung toxicity. 70.3.12.1 Prevention of Absorption Following the first reports of paraquat poisoning it was suggested that the immediate therapy of paraquat poisoning should be directed toward prevention of absorption from the gastrointestinal tract (Malone et aI., 1971). There is little information available on the use of gastric lavage in paraquat poisoned patients. Bismuth et al. (1982) were not able to establish a beneficial effect from gastric lavage in their series of 28 patients. Bramley and Hart (1983) did not find an improved prognosis resulting from the use of gastric lavage in a series of 262 patients. McDonah and Martin (1970) proposed urgent gastric lavage with a 1% bentonite solution to inactivate paraquat. Following the studies by Clark (1971) who found that bentonite (sodium montmorillonite) and Fuller's Earth (calcium montmorillonite) had a high adsorption capacity for paraquat, Douglas
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et al. (1973) reported three cases of survival after paraquat poisoning, two of which had been treated with 7% bentonite as adsorbent. Smith et al. (1974) suggested a treatment regime of repeated administration of cathartics together with large volumes of Fuller's Earth or bentonite which had been shown to effectively protect rats against an otherwise lethal dose of paraquat. Vale et al. (1977) used this approach together with charcoal hemoperfusion in 10 patients with paraquat poisoning. Only 1 patient who had the initially lowest plasma paraquat concentration survived, prompting the authors to conclude that the treatment was likely to be of benefit only in less severely poisoned patients. This was also the conclusion of Fitzgerald et at. (1979a) who analyzed 62 cases of paraquat poisoning with respect to treatment with Fuller's Earth and survival. They found that the majority of patients who survived had not taken what was regarded as a lethal dose. Also, death occurred in all patients who had ingested more than 30 ml of the concentrate, irrespective of therapy. In the group of patients who ingested between 5 and 30 ml and who received therapy within 6 h after ingestion 4 out of 7 survivors and 2 out of 5 non survivors had received Fuller's Earth. The authors suggested that Fuller's Earth may have been of benefit in a few cases who had taken slightly in excess of the lethal dosage, but it was unlikely to affect the outcome in the majority of patients with paraquat poisoning. While Fuller's Earth is still widely used in the first-line treatment of paraquat poisoning, the original claim by Clark (1971) that activated charcoal did not bind paraquat has been disputed. On the basis of in vitro binding studies and in vivo experiments, Okonek et al. (1982) suggested that the use of activated charcoal instead of Fuller's Earth was equally effective. This has prompted a revision of the advice given to medical practitioners in the United Kingdom (Department of Health, 1996) since activated charcoal is more likely to be immediately available in most hospitals and treatment centers. Other adsorbents such as the cation exchange resin kayexalate have been used (Yamashita et aI., 1987) but it is doubtful whether these have any benefit over the use of Fuller's Earth and activated charcoal. From 1979 onward a potent emetic, the phosphodiesterase inhibitor PP796, was gradually introduced in all paraquat formulations made by the major manufacturer (DenduytsWhitehead et al., 1985). It has been shown that following ingestion of emeticized formulations vomiting occurs earlier and is more profuse and prolonged than following ingestion of nonemeticized product (Meredith and Vale, 1987). However, a comparison of data from patients who had ingested paraquat concentrate with or without added emetic failed to show an overall benefit of the emetic on survival rate (Bismuth et al., 1982; Bramley and Hart, 1983; Onyon and Volans, 1987). Nevertheless, the emetic has been retained with the rationale that in particular accidental paraquat ingestions usually involve small quantities of the product, where early gastric emptying could have an effect on the outcome. It can be concluded that there is no clear evidence that gastric emptying and the use of adsorbents have improved the survival of patients with paraquat poisoning. The main reasons
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for this are the high dose of paraquat ingested by the majority of patients with deliberate ingestion and the frequent delay in hospital admission. Most authors concede that, on theoretical grounds, therapy designed to prevent absorption of paraquat should be able to help those patients who have a realistic chance of survival. However, clear evidence for this from clinical studies has so far not been obtained.
70.3.12.2 Elimination of Paraquat from the Body Since the kidney is the primary excretory organ for absorbed paraquat, enhancement of urinary elimination was one of the first therapeutic options considered. Kerr et al. (1968) published the first case report where forced diuresis had been used to treat paraquat poisoning. The exact fluid volume was not given, but their patient's urine excretion was more than llliters over 24 h. The total urine excretion of paraquat was 46 mg and the patient survived. Another patient was treated with a total of 27 liters of fluid over 48 h (Fennelly et aI., 1971). During the course of the forced diuresis he developed seizures, a metabolic alkalosis, and electrolyte disturbances, but the therapy was successfully completed. He developed transient mild hepatic and renal failure, but the only sign of pulmonary involvement was a slight temporary reduction in transfer factor. The authors suggested that this was a case of severe poisoning but were unable to attribute his survival to the forced diuresis therapy because the patient had also received immunosuppressive therapy with azathioprin and prednisolone. Bismuth et al. (1982) suggested that forced diuresis per se does not enhance the urinary elimination of paraquat. Nevertheless, they believed the therapy might be of value in the prevention of paraquat-induced renal damage because of a reduction in the tubular concentration of paraquat. However, of the 18 patients with developing renal failure who were treated with frusemide, only 1 survived despite the fact that diuresis was maintained in nine patients. Removal of paraquat by means of peritoneal dialysis, hemodialysis, and hemoperfusion has been advocated to reduce paraquat plasma concentrations and enhance elimination. Of these, dialysis procedures were found to be ineffective (Bismuth et aI., 1982; Vale et aI., 1977) and the value of charcoal hemoperfusion remains controversial. Experimental hemoperfusion in dogs was able to improve survival (Widdop et al., 1977), but early results in paraquat poisoned patients were disappointing (Vale et aI., 1977). In 1979, Okonek and co-workers published a report on the successful treatment of two patients with what they described as "continuous hemoperfusion." Plasma paraquat analysis prior to hemoperfusion indicated a very poor prognosis, but under an aggressive hemoperfusion therapy over several weeks both survived. Subsequently, a further 6 patients were treated with this regime and had a positive outcome (Okonek et aI., 1982/83). However, these apparent successes proved to be rare. Hampson and Pond (1988) carried out a meta-analysis of data from 35 cases published in the literature and 7 cases from their own hospital which had sufficient comparative data, as well as details of the haemoperfusion procedure. They showed that none of the patients whose initial
plasma paraquat concentration was higher than 3 mg/l survived, regardless of time after ingestion and treatment. Overall, the outcome was in line with predictions and did not appear to be affected by hemoperfusion, single or repeated. The authors concluded that hemoperfusion should only be considered for patients whose initial plasma concentration was below 3 mg/l, those in whom the probability of survival was between 20 and 70%, and those who present within a few hours of ingestion. Subsequently, B6hler et al. (1992) reported a case where the use of continuous arterio-venous hemoperfusion was effective in lowering the plasma paraquat concentration below the limit of detection. However, the patient died on the second day after ingestion from gastrointestinal complications. Suzuki et al. (1993) compared the effect of "aggressive" (> 10 hours in the first 24 h after ingestion) vs "conventional" « 10 h) hemoperfusion on the outcome of the intoxication in 40 patients. Aggressive hemoperfusion did not improve the overall outcome but significantly increased survival time. Finally, Lee and Lee (1995) found that 8 out of 18 patients treated with hemoperfusion survived, whereas none of 20 who did not receive hemoperfusion died. No plasma paraquat concentrations were measured, but the authors stated that the estimated volume ingested was not significantly different between the two groups. In conclusion, no clear benefit has been demonstrated from therapies aimed at enhancing elimination of paraquat from the body. The best chances appear to lie in the maintenance of renal function through adequate diuresis. As for extracorporeal elimination, hemoperfusion appears to be the only technique which may be of benefit in some patients, and the early and aggressive use of this technique may have contributed to survival in a few cases.
70.3.12.3 Pathophysiological Treatment A wide range of therapeutic substances have been studied experimentally in an attempt to prevent the specific lung toxicity of paraquat from occurring. Some have been used in humans, but most of the published work is based on single or a small number of cases. Usually, more than one therapy was employed, and information on the severity of poisoning and the initial probability of survival is often limited. For these reasons a critical evaluation of the benefit of anyone therapy is difficult and, in many cases, impossible. Since oxygen is required to set off the biochemical cascade of paraquat toxicity the use of supplementary oxygen should be avoided as long as possible. Bismuth et al. (1982) used a hypoxic breathing mixture and hypothermia in six patients. The arterial oxygen tension was maintained below 6.6 kPa. Only one patient survived who had clinical evidence of only mild poisoning. In the other patients, the Fi02 had to be increased on a daily basis, all of them requiring >0.5 (50%) prior to their death. Since redox cycling and the generation of free radicals are considered to be the principal a steps in the development of alveolar epithelial cell damage, a number of agents which, at least theoretically, interfere with this process have been tried
70.3 Toxicity to Humans
therapeutically. One of the first steps in the biochemical cascade of injury is the generation of the superoxide anion which is detoxified by the enzyme superoxide dismutase. This has been given either intravenously (Davies and Connolly, 1975), intramuscularly (Harley et aI., 1977), intrapulmonary during fiberoptic bronchoscopy (Bateman, 1987), or as a nebulized aerosol (Davies and Connolly, 1975; Hong et aI., 1996). In some cases there was co-administration with the antioxidants vitamin C (Hong et aI., 1996) or vitamin E (Harley et aI., 1977) which has also been given on its own (Shahar et aI., 1980). The doses given appeared to have been determined empirically, and no conclusive evidence of a beneficial effect has so far been shown. N-acetyl cysteine (NAC) is a glutathione precursor which readily crosses the cell membrane, and glutathione depletion is one of the features of paraquat-induced cellular damage, Lheureux et al. (1995) treated a patient with high doses of NAC (300 mg/kg/day) over 3 weeks. However, the patient who survived also received early hemodialysis and desferrioxamine. The latter, an iron chelating agent, has been proposed because iron has a catalytic effect in the production of hydroxyl radicals. However, no other data exist on its clinical use. 70.3.12.4 Prevention of Lung Fibrosis The development of the paraquat lung lesion is characterized by early infiltration of inflammatory cells, followed by fibroblast proliferation. Attempts have therefore been made to halt this process by giving immunosuppressive therapy. A few case reports involved the use of azathioprine, in one case with successful outcome (Laithwaitte, 1975); in two other cases the patients died (Malcolmson and Beesley, 1975). In one patient who survived, bleomycin was used over 3 days (Mahieu et aI., 1977). Most experience exists with a combination treatment of cyclophosphamide and corticosteroids which was first advocated by Malone et al. (1971). Addo et al. (1984) claimed a 75% survival rate in 20 patients treated with, cyclophosphamide (5 mg/kg/day to a maximum total of 4 g) and dexamethasone (8 mg t.i.d. over 2 weeks). Two years later they published a case series using the same regime with 72 patients, 52 (72%) of which survived (Addo and Poon-King, 1986). However, the plasma paraquat data of 25 patients showed that 7 survivors had no measurable paraquat levels, and of the other 18 only the 6 patients with the lowest plasma concentration survived. Following a preliminary report (Lin et aI., 1996) on the use of pulse therapy with cyclophosphamide (1 g/day over 2 days) and methylprednisolone (1 g/day over 3 days), Lin et al. (1999) reported results of a prospective study in 142 patients. Seventyone patients died from fulminant poisoning within 1 week, and cyclophosphamide did not make any difference. In the group of moderately to severely poisoned patients, only 4122 patients treated with cyclophosphamide died, compared to 16128 in the control group. Plasma paraquat concentrations were not available, but the authors stated that there was no difference in severity of poisoning between the two groups based on the urine dithionite test. However, the beneficial effects of the cyclophosphamide-dexamethasone regime have been disputed
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(Nogue et aI., 1989), and in a prospective study Perriens et al. (1992) did not find any difference in mortality between 14 patients who had received standard treatment and the 33 patients who had received high-dose cyclophosphamide and dexamethasone. A final answer regarding the usefulness of this therapy can therefore not been given at this stage. Because of the radiosensitivity of fibroblasts in vitro, Webb et al. (1984) treated a patient who had developed diffuse alveolar damage following paraquat ingestion initially with cyclophosphamide and, after further deterioration, with fractionated radiotherapy over 11 days. The patient survived. It was noted that the severity of poisoning in this patient was mild (Proudfoot et aI., 1984) and the majority of patients in subsequent reports died (Bloodworth et aI., 1986; Williams and Webb, 1987). This may have been due to differences in the severity of intoxication, as well as the therapy employed. Following the successful treatment of a patient with poor prognosis (Talbot et aI., 1988a), Talbot and Barnes (1988) treated a further eight patients with radiotherapy. Only two survived and the authors suggested that a definite benefit of radiotherapy could not be demonstrated in their study. 70.3.12.5 Other Treatments Beta-blocking agents such as propranolol have been shown to block the uptake of paraquat into the lung (Maling et aI., 1978). However, their limited therapeutic use has not been successful (Davies and Connolly, 1975; Fairshter et aI., 1976, 1979). Recently, there have been two case reports on the use of nitrogen oxide inhalation (NO) in paraquat poisoning. On the basis that NO is a potent endogenous vasodilator and that NO inhalation exerts a beneficial effect on pulmonary gas exchange, Koppel et al. (1994) treated a 52 year old patient with severe paraquat poisoning (plasma concentration 4 days after ingestion 1 mg/l). She received 25 ppm in the inhalation mixture, her respiratory parameters improved immediately, and she was stabilized for 3 days. However, the patient died with massive pleural effusions and ventilatory failure on day 11 after ingestion. In the second case, Eisenman et al. (1998) treated a 52 year old male whose plasma paraquat concentration predicted only a 30% chance of survival, with NO because of developing respiratory distress. In addition, the patient had received Fuller's Earth, forced diuresis, hemofiltration, N -acetyl cysteine, methyl prednisolone, cyclophosphamide, vitamin E, and colchicine. Because of the multiple therapy it was impossible to be sure which of the therapeutic measures had contributed to this patient's survival. Nevertheless, it was felt that the use of NO deserved further evaluation (Hall, 1998). There are five reports in the literature where lung transplantation has been performed after paraquat poisoning. Matthew et al. (1968) described a single lung transplantation 6 days after accidental paraquat ingestion in a 15 year old boy whose plasma paraquat levels at the time of the operation were still at toxic levels (0.4 IJ-g/ml). The patient died l3 days after the operation in respiratory failure and the autopsy showed changes typical for paraquat poisoning, although no paraquat was measurable in the transplanted lung. A contribution of rejection to
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CHAPTER 70 Paraquat
the disease process could not be excluded. The same group subsequently reported a further unsuccessful lung transplantation in an 18 year old farm worker (Cooke et aI., 1973). A further single lung transplantation with fatal outcome in a 25 year old man was reported in 1984 by Kamholz et al. This patient died after 45 days from the consequences of a bronchopleural fistula. A sequential bilateral lung transplantation was described by the Toronto Lung Transplant Group (Saunders et aI., 1985). This 31 year old patient had received a lung transplant at a time when his plasma paraquat levels were below what was considered to be a toxic level (0.03 f-Lg/ml). In the postoperative period there was a sevenfold increase in paraquat plasma levels, possibly due to the release of paraquat from muscle stores. The transplanted lung failed and following several days of intensive therapy including hemoperfusion a second transplant was undertaken. The transplant worked well over 2 months; however, the patient developed a progressive toxic myopathy and ultimately died from a cerebrovascular accident. Nevertheless, this case showed the feasibility of lung transplantation in paraquat poisoning. The most recent case (Licker et aI., 1998) is the only one with a successful outcome. A 17 year old farmer developed respiratory failure of unknown origin. Repeated plasma paraquat measurements were negative. Following mechanical ventilation for 5 weeks a single lung transplantation was carried out. Recovery was complicated by myopathy, and paraquat was confirmed in the excised lung and a muscle biopsy. The patient subsequently admitted to having taken paraquat. The patient was discharged after 88 days and was able to lead an independent life at the last follow-up 13 months after transplantation. These cases demonstrate that, over the years, lung transplantation has become feasible in cases of paraquat poisoning. While the early attempts were hampered by problems with immune suppression as well as a lack of understanding of the pathophysiological events following paraquat poisoning, these problems appear now to have been satisfactorily resolved. However, the authors of the latest paper make the point that the use of such a scarce and expensive resource is questionable in cases of deliberate self-harm. Disclaimer The positions on certain aspects of the toxicology of paraquat in this chapter may not be aligned with the Syngenta positions; the latter are derived mainly from internal Syngenta reports many of which have not been published in the open literature.
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CHAPTER 70
Paraquat
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CHAPTER
71 Diquat Edward A. Lock Syngenta Central Toxicology Laboratory
Martin F. Wilks Syngenta Crop Protection AG
71.1 IDENTITY, PROPERTIES, AND USES
71.1.4 PHYSICAL AND CHEMICAL PROPERTIES
71.1.1 CHEMICAL NAME Diquat is 1,l'-ethy1ene-2,2'-bipyridyldiy1ium ion (IUPAC), or 6,7-dihydrodipyrido[1,2-a : 2',l'-c]pyrazinediium (CAS RN [2764-72-9J).
71.1.2 STRUCTURE
The molecular formula of the cation is C 12H 12N2 with a molecular weight of 184.24. The dibromide salt has the formula C12H12Br2N2 and a molecular weight of 344.1. Diquat dibromide forms colorless to yellow crystals which decompose above 300°C. With a vapour pressure of <0.013 mPa (monohydrate), it is practically nonvolatile. It is very soluble in water (700 g/l at 20°C), slightly soluble in alcohols and hydroxylic solvents, and insoluble in nonpolar organic solvents. It is stable in neutral and acidic media but readily hydrolyzed in alkaline media. Diquat is photochemically decomposed by ultraviolet radiation.
71.1.5 HISTORY, FORMULATIONS, AND USES
Diquat dibromide
Figure 71.1
71.1.3 SYNONYMS The common name diquat is in general use (BSI, E-ISO, (m) F-ISO, ANSI, WSSA, JMAF), except in Germany (deiquat) and Russia (regIon). Diquat is formulated as the dibromide salt (CAS NR [85-00-7J and [6385-62-2J for the dibromide monohydrate). A code designation for the material is FBI2. Trade names for diquat dibromide formulations include Desiquat®, Midstream®, Reglone®, and Reglex®. Mixtures of diquat with paraquat are sold under trade names including Actor®, Dukatalon®, Opal®, Pathclear® (also includes simazine and aminotriazole), Preeglox®, Preglone®, Seccatutto®, Spray Seed®, and Weedol®. Handbook of Pesticide Toxicology Volume 2. Agents
The herbicidal properties of diquat were first described by Brian et al. (1958). Diquat is mainly formulated as an aqueous solution. In some countries, a low-strength granular formulation (also containing paraquat) is available. Diquat is a fast-acting, nonselective contact herbicide and desiccant, absorbed by the foliage with some translocation in the xylem. It is used for preharvest desiccation of cotton, flax, alfalfa, and many other crops, as a defoliant on hops, for destruction of potato haulms, for general weed control on noncrop land, for weed control and tassel inhibition in sugar cane, and for control of emergent and submerged aquatic weeds. Diquat is rapidly deactivated upon contact with the soil and does not leach.
71.2 TOXICITY TO LABORATORY ANIMALS 71.2.1 SIGNS OF TOXICITY Following a median lethal oral dose of diquat to rats, few signs of toxicity were seen over the first 24 h. Subsequently the an-
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CHAPTER 71
Diquat
imals became lethargic, lost weight, showed slight pupillary dilatation, and excreted mucoid, ropey feces of a characteristic green color which was associated with mild abdominal distension. Renal function was impaired and the animals died 2-14 days after dosing (Clark and Hurst, 1970; Crabtree et aI., 1977; Lock, 1979). Similar signs of toxicity were seen in mice, guinea pigs, rabbits, dogs (Clark and Hurst, 1970), and cynomologus monkeys (Cobb and Grimshaw, 1979). Following a single median lethal subcutaneous (sc) injection, rats showed a marked diuresis, became lethargic about 6 h after dosing, and had marked pupillary dilatation which persisted. By 24 h the rats had lost weight and were drinking less, the animals generally became weaker, with deaths occurring between 2 and 10 days, although some animals died as late as 8 weeks. All animals dying after 10 days had grossly distended abdomens due to a grossly swollen caecum. The cause of death is unclear; hypovolaemia and renal tubular necrosis leading to renal shutdown are contributory factors. Histological findings showed injury to the lining of the stomach and gastrointestinal tract but these were not life threatening. A large injection of diquat (four or five times a median lethal dose, sc) to rats produced subdued behavior within a few minutes and labored respiration within 1 h. Muscular twitching then occurred, leading to generalized convulsions and death within a few hours (Clark and Hurst, 1970). Lung injury similar to that seen with paraquat is not a prominent feature with diquat. 71.2.2 SPECIES AND DOSE-RESPONSE RELATIONSHIPS 71.2.2.1 Acute Toxicity
The acute oral toxicity of diquat to a number of species is shown in Table 71.1. The median lethal dose (MLD) of pure diquat dibromide or dichloride expressed as the cation is between 121 to 230 mg/kg in male and female rats. Diquat is, however, more toxic when given by the intraperitoneal (ip) or sc routes. The MLD by injection is about 11 mg/kg, indicating that following oral dosing the compound is poorly absorbed from the gastrointestinal tract. For most other laboratory species the MLD is in the range of 100-200 mg/kg (Table 71.1). Dermal exposure of rabbits to diquat for 24 h produced no ill effects at 400 mg/kg, the maximum dose that could be applied to the skin (Clark and Hurst, 1970). For the mouse and guinea pig the MLD was approximately 400 mg/kg and for the rat about 650 mg/kg (Table 71.2). This finding is consistent with in vitro studies showing that diquat is poorly absorbed across the skin of laboratory animals and humans (see later) from an aqueous solution (Scott and Corrigan, 1990). Diquat is not volatile; inhalation exposure of rats to an aerosol of respirable droplets ( < 5 Jl.m diameter) at 23 Jl.g/l for 30 min produced no ill effects (Gage, 1968a). The MLD for diquat by inhalation is about 35 Jl.g/l for rats and guinea pigs (Table 71.2) while a value of 83 Jl.g/l was reported for rats by Bainova and Ve1cheva (1977), the difference presumably being due to particle size or analytical differences. Direct intratracheal administration of diquat to
rats has been shown to produce lung injury and fibrosis (Manabe and Ogata, 1986, 1987). 71.2.2.2 Irritation and Sensitization
A single application of diquat to the skin of mice (Bainova, 1969b) or rabbits (Clark and Hurst, 1970) did not cause local irritation. Diquat is not a skin sensitizer (Bainova, 1969b). Daily applications of diquat at 20 mg/kg/day for 20 days in water to the skin of rabbits provoked some mild erythema, thickening of the skin, and some scabbing at the site of application (Cl ark and Hurst, 1970). Instillation of one drop of a 20% aqueous solution of diquat into the rabbit eye produced slight conjunctival irritation which persisted for 2 days (Clark and Hurst, 1970). No pupillary dilatation was observed. 71.2.2.3 Subchronic and Chronic Toxicity
The primary target organ for toxicity following prolonged exposure to diquat is the eye, where cataracts develop. Daily oral dosing of rats with diquat at 6.5, l3, and 40 mg/kg/day for 40 days or 2.1 or 4.3 mg/kg/day for 4.5 months produced histological changes in the liver, kidney, lung, and gastrointestinal tract (Bainova, 1969a, 1975). The effects on these organs will be discussed later. Diquat, when administered to rats in their drinking water at 2 and 4 mg/kg/day for up to 2 years, did not increase mortality, although some histological changes were seen in the lungs at the high dose (Bainova and Ve1cheva, 1978). Daily administration in the diet of rats at 10, 50, 100, 250, 500, or 1000 ppm diquat dichloride for 2 years produced no compound-related deaths. Some reduction in food consumption and body weight gain was seen at 1000 ppm but not at any of the lower doses. Gross and microscopic examination revealed no compound-related changes other than in the eye where cataracts developed in rats receiving 50 ppm and above during the course of the study. The time to onset of cataracts was dose-related. At 1000 ppm partial or complete opacities were present in one or both lenses within six months. Bilateral cataracts were seen at 12 months following 500 ppm (about 25 mg/kg/day), lens opacity was detected in all animals at 250 ppm at 18 months, while at 100 and 50 ppm about 25% of the animals were affected at 2 years. No effects on the eye were seen at 10 ppm (about 0.5 mg/kg/day) for 2 years in any animals (Clark and Hurst, 1970). Rats fed diquat at 500 ppm for 8 weeks and then returned to normal diet for 1 year did not develop cataracts, showing that continuous and prolonged exposure is necessary in the rat for cataract formation (Clark and Hurst, 1970). In a subsequent dietary study, no cataracts were observed when rats were fed 20 ppm diquat dichloride (about 1 mg/kg/day) in their diet for 2 years (FAOIWHO, 1978). Dogs can tolerate doses of 15 mg/kg/day diquat dichloride for 2 years without affecting growth or producing any histological effects, except in the eye. Bilateral cataracts appeared at 15 mg diquat/kg/day after about 10-11 months. However, dogs exposed to 1.7 mg diquat/kg/day for 4 years did not develop cataract (Clark and Hurst, 1970). Subchronic exposure following dermal application to rabbits at 20 mg/kg/day for 20 days caused some mild erythema,
71.2 Toxicity to Laboratory Animals
1607
Table 71.1 Acute Toxicity of Diquat to Laboratory Animals (Data Expressed as mglkg Diquat Ion) Route of
Median lethal
Species
Sex
Strain
administration
dose
Reference
Rat
F
Alderley Park-
po
231 (194--274)
Clark and Hurst, 1970
Wistar Rat
M
Sherman
po
147 (138-155)
Gaines and Linder, 1986
Rat
F
Sherman
po
121 (108-136)
Gaines and Linder, 1986
Rat
M
Alderley Park-
po
226
Crabtree et aI., 1977
sc
11 (5-15)
Clark and Hurst, 1970
sc
11 (9-12)
Clark and Hurst, 1970
Rat
po
281
Verbetskii and Pushkar, 1968
Rat
po
215
Makovskii, 1972
Wistar Rat
M
Alderley ParkWistar
Rat
F
Alderley ParkWistar
Rat
po
130
Bainova, 1969a
ip
<11
Smith and Rose, 1977
Rat
M
Alderley Park-
Mouse
M
Alderley Park
po
125 (106-146)
Clark and Hurst, 1970
Dog
F
Alderley Park-
po
100-200
Clark and Hurst, 1970
Wistar
beagle Guinea pig Guinea pig
F
Alderley Park
Rabbit
F
Albino
Cynomolgus
M
po
123
Verbetskii and Pushkar, 1968
po
approx 100
Clark and Hurst, 1970
po
101 (72-138)
Clark and Hurst, 1970
po
100-300
Cobb and Grimshaw, 1979
monkey
thickening of the skin, and scabbing. Increasing the dose to 40 mg/kg/day caused weight loss and muscular weakness. Four of the six animals in this group died after 8 to 20 applications. The MLD was therefore between 20 and 40 mg/kg/day in the rabbit (Clark and Hurst, 1970). Dermal application of diquat to rats at 5 to 120 mg/kg/day for 20 days, without occluding the skin, produced slight skin irritation at the site of application. An increase in signs of toxicity was observed at 10 mg/kg/day and above, including distension of the abdomen. Mortality was seen at the
Table 71.2 Acute Toxicity of Diquat to Laboratory Animals Following Dermal Application or Inhalation Exposure (Data Expressed as mg/kg Diquat Ion) Route of
Median lethal
Species
administration
dose
Reference
Rat
Dermal
650
Makovskii, 1972
Mouse
Dermal
430
Bainova, 1969b
Guinea pig
Dermal
400
Makovskii, 1972
Rabbit
Dermal
>400
Clark and Hurst, 1970
Rat
Inhalation
35 ",g/l
Makovskii, 1972
Rat
Inhalation
83 ",g/l
Bainova and Ve1cheva,
Guinea pig
Inhalation
38 ",gIl
Makovskii, 1972
1977
higher doses and at toxic doses there was histological evidence of injury to the liver, kidneys, lung, and gastrointestinal tract (Bainova, 1969b). The median lethal dose was 35 mg/kg/day with no effects being seen at 5 mg/kg/day. Repeated exposure of rats to diquat by inhalation for 6 h per day for 15 days over 3 weeks at 2 ).Lg/l caused signs of respiratory irritation in females. Histological examination showed evidence of irritation with peribronchial lymphoid hyperplasia and slight perivascular oedema. No effects were seen at 0.5 ).Lg/l over the same exposure time (Gage, 1968a). Rats, mice, guinea pigs, rabbits, and a dog exposed to 1.06 ).Lg/l for 15 daily exposures of 6 h remained in good condition during exposure, apart from the rabbits which initially showed signs of rapid, shallow breathing which had recovered at the end of the study (Gage, 1968a). Rats exposed to doses ranging from 0.32 to 1.9 mg/l for 4 or 6 h per day over 4 to 4.5 months showed signs of lung irritation and lung injury at the higher doses. No effects were seen at the lowest doses of 0.32 mg/l for 6 h/day for 4.5 months or at 0.4 mg/l for 4 h/day for 4 months (Bainova et a!., 1972; Makovskii,1972). Overall, these studies show that following acute and chronic exposure some toxicity is seen in the gastrointestinal tract, liver, kidney, and lung. The major target organ for toxicity following chronic exposure is the eye with the production of cataracts. Dose levels of diquat that do not cause damage in laboratory
1608
CHAPTER 71
Diquat
animals following acute and chronic exposure have been clearly established. 71.2.2.4 Mutagenic and Carcinogenic Potential Diquat is not carcinogenic in either rats or mice. The activity seen in some short-term assays for mutagenesis is associated with cytotoxicity, believed to arise as a consequence of the redox cycling ability of diquat, leading to superoxide anion formation. Diquat has minimal to no genotoxic activity when evaluated in a wide range of in vitro and in vivo test systems. Many groups have reported the absence of an effect in the Ames assay (Andersen et aI., 1972; Benigni et aI., 1979; Levin et aI., 1982). Diquat was not mutagenic when tested in the mouse dominant lethal assay (Anderson et aI., 1976; Pasi et aI., 1974) or in studies including chromosomal aberrations in mice (Selypes et aI., 1980). Some positive effects have been observed with gene conversion in Saccharomyces cerevisiae (Siebert and Lemperle, 1974), DNA repair in Salmonella typhimurium, gene mutation in Aspergillus nidulans (Benigni et aI., 1979), and sister chromatid exchange in Chinese hamster lung cells (Tanaka and Amano, 1989). These effects were usually associated with cytotoxicity and are believed to arise as a consequence of the redox cycling ability of diquat, leading to the production of superoxide anion (see later). It is well known that DNA damage frequently occurs when cells are exposed to oxidative stress (Brawn and Fridovich, 1981; Repine et aI., 1981). Recent studies using the 32P-postlabeling technique showed no differences in DNA adducts in the liver of diquat-treated rats compared to controls (Vulimiri et aI., 1995). Importantly diquat did not induce tumors in a 2 year feeding study in rats or mice (Clark and Hurst, 1970; FAOIWHO, 1993) or when given to rats in their drinking water for 2 years (Bainova and Velcheva, 1978). 71.2.2.5 Effects on Reproduction, Embryotoxicity, and Teratogenicity Diquat has no effect on fertility, is not teratogenic, and only produces fetotoxicity at doses that are maternally toxic. The main finding in the multi generation study was cataract formation. The testes of male rats dosed orally with diquat dibromide at 6.5 mg/kg/day for 60 days were histologic ally normal as was the sperm count and sperm motility (Bainova and Velcheva, 1974). In a multigeneration reproduction study, rats were fed diets containing 0, 16, 80, or 400 ppm diquat for 12 weeks, mated, and then allowed to rear the litters that resulted (Fla). The process was repeated with animals selected from the Fl a litter, these FI parents being mated 11 weeks after selection. The dose received by the top dose F1 rats was reduced after 4 weeks to 240 ppm. Diquat had no effect on fertility in either sex; the main finding was cataract formation. Decreased body-weight gain was seen at the top dose in both adults (Fo and F 1) and pups. Cataract formation was mostly confined to the top dose although a low incidence was seen at 80 ppm in the F 1 female parents. No cataracts were seen at 16 ppm (equivalent to 0.8 mg/kg/day) (FAOIWHO, 1993).
A single ip injection of diquat to rats at 7 mg/kg during days 6-14 of gestation produced a marked reduction in body weight gain and retarded ossification but was not teratogenic. Repeated doses of 0.5 mg/kg/day diquat ip did not produce embryotoxicity (Khera et al., 1970). Bus et al. (1975) reported that diquat can cross the rat placenta as fetuses contained 4 C] diquat following administration to the mother on days 13, 16, or 21 of gestation. Bus and co-workers also reported that diquat given at 15 mg/kg iv to rats on days 7-21 of gestation resulted in significant fetal reabsorption and maternal death. This is not unexpected as the MLD for diquat given by ip injection is 11 mg/kg (Table 71.1). The embryotoxic effects of high doses of diquat have also been observed in mice receiving 2.7 or 11 mg/kg on days 9, 10, 11, and 12 of gestation. An increase in the number of dead fetuses and postimplantation loss was observed, but no congenital malformations were seen (Selypes et aI., 1980). Oral administration of diquat to pregnant rats at doses of 4, 12, or 40 mg/kg diquat cation/day from days 7-16 of gestation resulted in maternal toxicity, reductions in fetal weight and litter weight, as well as fetal defects in ossification at the top dose. No significant effects were seen at 12 mg/kg diquat cation/day (FAOIWHO, 1993). Oral administration of diquat to pregnant rabbits at doses of 1, 3 or lOmg ion/kg/day for days 7-19 of gestation resulted in maternal toxicity at the top dose with some evidence of fetotoxicity in the form of partially ossified sternbrae. No adverse effect on the mother or fetuses was seen at 1 mg/kg/day (FAOIWHO, 1993).
e
71.2.2.6 Pathology At postmortem following either oral or sc administration of diquat the most obvious effect is gross distension of the gastrointestinal tract with greenish-yellow fluid. The color is believed to be due to bacterial reduction of diquat. Histological changes to the gastrointestinal tract, liver, kidneys, lungs, and eyes have been reported and these will be discussed later under each target organ. 71.2.2.7 Absorption The first studies on the absorption and excretion of diquat were conducted by Daniel and Gage (1966) in rats. They showed, following a single oral dose of 4 C-ethylene]diquat dibromide or dichloride, that most of the radioactivity was excreted within 48 h, although at the higher doses some appeared in the feces on day 3. Between 6 and 10% of the dose was excreted in the urine over 48 h, the remainder being in the feces. In contrast, when diquat dibromide was given sc the bulk of the radioactivity appeared in the urine within 24 h of dosing, showing that diquat is not completely absorbed across the gastrointestinal tract of the rat. Subsequent studies in rats have reported 5.5% of an oral dose of 60 mg/kg diquat ion excreted in the urine over 1-7 days (Litchfield et al., 1973) and 7.5% of an oral dose of 126 mg/kg diquat ion in 24 h (Lock and Ishmael, 1979). Following oral administration of diquat (126 mg/kg) to rats, the peak plasma concentration occurred before 2 h, the earliest time measured (Rose et aI., 1976a), and then remained constant for
e
71.2 Toxicity to Laboratory Animals
up to 30 h. Studies in the dog using a tracer dose of 12 jl.g/kg of 4C-ethylene]diquat did not result in an early plasma peak with only 10-20% of the dose absorbed in 6 h (Bennett and Davies, 1976). The dog appears to absorb a slightly larger percentage of an orally administered dose of diquat than the rat, which is consistent with the greater susceptibility of the dog by this route of administration. Overall, few absorption studies have been reported with diquat. From the available information diquat appears to be rapidly but incompletely absorbed from the gastrointestinal tract of laboratory animals with peak plasma concentrations occurring within a few hours of dosing.
e
71.2.2.8 Distribution In the rat after an oral dose of 126 mg/kg, the plasma diquat concentration remained constant between 2 and 30 h after dosing (Rose et aI., 1976a). During this time no accumulation into the lung was seen, although the concentration in the adrenal gland and to a lesser extent the liver was higher than that found in the plasma. In no other organs, apart from the kidney, which is the major organ for the excretion of diquat, was the concentration above that found in the plasma (Rose et aI., 1976a). Diquat did not appear to enter the brain (Rose et aI., 1976a). These findings, plus the earlier observation of Sharp et al. (1972) who compared the tissue distribution of diquat and paraquat following 20 mg/kg iv, confirmed that diquat was not retained in the lung. The only organs with higher concentrations of diquat compared to paraquat were the liver and at later times the kidney. Others have subsequently studied the distribution of diquat in the rat following oral or systemic administration with similar findings (Kurisaki and Sato, 1979; Matsuura et aI., 1978; Spalding et aI., 1989). Following dietary administration of diquat to rats for 2, 4, or 8 weeks at 250 ppm a time-dependent increase in diquat was detected in the kidneys, with lower levels in the liver and lung, while in the brain diquat was at or below the limits of detection (Litchfield et aI., 1973). Whole body autoradiography studies have also provided valuable information on the tissue distribution of diquat. Early studies by Litchfield et al. (1973) in mice given iv [14C_ ethylene ]diquat showed the compound was distributed throughout most tissues 10 min after injection with higher concentrations associated with cartilaginous tissues, the gall bladder, the small intestine, and the urinary bladder. By 1 h the concentration of radioactivity had declined in most tissues apart from the gastrointestinal tract and urinary bladder. By 24 and 48 h excretion was virtually complete apart from some radioactivity present in the gastrointestinal tract. Whole body autoradiography studies have shown that diquat binds to melanin following iv administration to pigmented C57 black mice, but not albino mice. The association of diquat with melanin is probably due to an ionic interaction (Larsson et aI., 1977).
71.2.2.9 Metabolism Diquat is poorly metabolized, the bulk of the material being excreted in the urine and feces unchanged. Daniel and Gage
1609
(1966) and Hughes et al. (1973) compared the colorimetric assay for diquat with that found by radiochemical detection in the urine and feces of rats dosed with diquat and showed that they agreed very closely. Incubation of diquat with rat caecal contents for up to 24 h showed up to a 50% loss, indicating microbial metabolism, as the loss was prevented when the contents ofthe caecum were heat-treated (Daniel and Gage, 1966). However, in vivo studies in rats have not shown significant biotransformation, indicating that the in vitro studies had overpredicted the likely metabolism. Hughes et al. (1973) reported some unidentified metabolites of diquat in the urine of rabbits and guinea pigs. Subsequent studies in the rat identified diquat monopyridone as a metabolite mainly in the feces, at about 5% of an oral dose, while diquat dipyridone was detected in the urine (FAOIWHO, 1978). Overall these studies indicate that diquat is probably metabolized by gastrointestinal bacteria.
71.2.2.10 Excretion Whole body autoradiography showed that diquat was present in the gall bladder of mice, indicating biliary excretion (Litchfield et al., 1973). The extent of biliary excretion of diquat was < 5% when dosed to bile cannulated rats, rabbits, or guinea pigs and bile collected over a 3 hour period (Hughes et aI., 1973; Spalding et aI., 1989). The major route of diquat elimination from the body is via the kidneys. The renal clearance of diquat is greater than that of inulin in the rat (Lock, 1979), indicating that diquat is actively secreted. Accumulation of the organic cation N 1_ methy Inicotinamide, but not the organic anion p-aminohippuric acid, by slices of rat renal-cortex was reduced by diquat, suggesting that it is actively secreted via a cationic transport system analogous to that for paraquat (Lock and Ishmael, 1979). The renal transport systems for the excretion of organic cations are discussed in detail in the chapter on paraquat toxicology. In the rat in vivo, the fractional excretion of diquat was 1.14 at a plasma concentration that may have saturated the transport system analogous to that reported for paraquat (Chan et aI., 1998). Thus at low plasma concentrations diquat is probably readily cleared from the body; however, at higher plasma concentrations this system will become saturated and less diquat is cleared. At toxic doses it is well established that diquat can cause renal functional impairment. In rats (Lock, 1979) and monkeys (Cobb and Grimshaw, 1979) given 100 mg diquat cation/kg orally, renal impairment was observed 24 h after dosing (Fig. 71.2).
71.2.2.11 Biochemical Mechanisms of Diquat Toxicity Diquat can be reduced to form a free radical, which is bright green in color and stable in aqueous solution in the absence of oxygen:
1610
Diquat
CHAPTER 71
::a
300
~-~~~----
Ol>
§
....,""
200
e
lOO
;;;J
..,2
r.n
0 600
~§
400
r.n
200
...
< 0
o
24
4
48
Time (h) Figure 71.2 The effect of diquat on markers of liver and kidney injury in cynomolgus monkeys following a single oral dose. Results are mean and range for two animals per treatment at lOO and 300 mg/kg and mean ± SE of four animals at 200 mg/kg. 100 mg/kg (_); 200 mg/kg (D) and 300 mg/kg (hatched). From Cobb and Grimshaw (1979).
In the presence of oxygen, in biological systems, the radical will rapidly reoxidize to the cation with the concomitant production of superoxide anion (Oz,). DQ+'
+ 02 ~ DQ2+ + Oz'
Thus, once diquat enters a cell it will undergo alternate reduction followed by reoxidation, a process known as redox cycling. Gage (1968b) first reported that the diquat cation could be reduced by rat liver NADPH-dependent microsomal flavoprotein reductase to form the radical, with the concomitant oxidation of NADPH. Redox cycling of diquat has also been reported in microsomal preparations of lung, liver, and kidney (Baldwin et aI., 1975; Tomita, 1991). Tomita (1991) also demonstrated one electron reduction of diquat by mitochondrial fractions of liver, lung, and kidney with the highest activity in the kidney. Thus, like paraquat, diquat can redox cycle, the major difference being that diquat can more readily accept an electron than paraquat, such that this response is seen at lower intracellular concentrations with diquat. The mechanism is similar to that discussed earlier for paraquat, whereby a cascade of events is triggered, leading to NADPH depletion and lipid peroxidation with the free radical scavenging enzymes such as superoxide dismutase, catalase, and glutathione peroxidase playing a key protective role. The relevance of this mechanism to the toxicity observed in laboratory animals will now be discussed. 71.2.2.12 Effects on the Lung
Diquat is not accumulated into the lung, unlike the situation with paraquat, following oral or systemic administration to rats or mice (Keeling et aI., 1981; Litchfield et aI., 1973; Rose et aI., 1976a; Sharp et aI., 1972; Spalding et aI., 1989; Witschi et aI., 1977). This is consistent with studies using rat lung slices where diquat does not accumulate into the slice (Rose et aI., 1974a, 1976a). Although diquat is not accumulated it is able to block
the entry of paraquat into lung cells via the energy-dependent transport system (Rose and Smith, 1977), suggesting it can interact with the transport system but not undergo transport itself. Sufficient diquat can, however, enter lung cells in vitro to undergo redox cycling and thereby stimulate the pen to se phosphate pathway (Rose et aI., 1976b). Studies in vivo with toxic oral or systemic doses have shown that diquat does not cause histopathological evidence of lung injury, like that seen with paraquat (Clark and Hurst, 1970; Cobb and Grimshaw, 1979). Similarly diquat does not cause pulmonary edema or alter lung function following ip administration, although it does reduce pulmonary cell turnover (Lam et aI., 1980; Smith and Rose, 1977). Diquat can, however, stimulate the pentose phosphate pathway in the lung following toxic oral or iv administration (Rose et aI., 1976b). Studies by Witschi et al. (1977) showed for the first time that a large dose of diquat (40 mg/kg, iv) produced a marked fall in the NADPHlNADP ratio in the lung, indicative of redox cycling of diquat in lung cells, and further that exposure of these rats to 100% oxygen enhanced the toxicity. These workers also reported damage to type I alveolar epithelial cells, following this large dose of diquat. Subsequent studies by Keeling and Smith (1982) in rats given 20 mg/kg diquat iv did not detect changes in NADPH or NADP in the lung at various times after dosing, although they did find a persistent increase in the total disulphide content of the lung, suggesting redox stress. The finding that diquat in the presence of oxygen can enhance the toxicity to rats has also been confirmed and damage to alveolar type 11 cells and pulmonary edema may contribute to the death of the animal under these conditions (Keeling et al., 1981; Kehrer et aI., 1979). Direct intratracheal administration of diquat to rats can produce lung damage and fibrosis, although much larger doses were required than for paraquat (Manabe and Ogata, 1986, 1987). The early studies of Bainova (l969a), Bainova et al. (1972), Bainova and Velcheva (1978), and Makovskii (1972) also reported lung irritation and injury following subchronic and chronic exposure. Overall these studies indicate that if diquat enters a lung cell it can redox cycle which, if extensive, can overwhelm the defense mechanisms, leading to cell death. However, lung injury is not a contributory factor to the death of animals receiving a MLD and exposed to air. 71.2.2.13 Effects on the Gastrointestinal Tract
The most obvious postmortem observation following diquat administration is distension of the terminal ileum and caecal region of the gastrointestinal tract (Clark and Hurst, 1970; Cobb and Grimshaw, 1979; Crabtree et al., 1977; Pushkar, 1969; Verbetskii and Stolyarchuk, 1967). Following an oral dose, diquat produced a dose-related increase in the water content in the lumen of the gastrointestinal tract (Fig. 71.3), which 24 h after an MLD was about 14 ml/rat. This results in tissue dehydration and hypovolaemia (Fig. 71.3) (Crabtree et aI., 1977). The marked decrease in blood volume (Lock, 1979) would be expected to have an effect on peripheral circulation and presumably contributes to the reduced renal function observed
71.2 Toxicity to Laboratory Animals
60
""'....0
= ~.... = '"""
40
"'40 "C..c
..... 20
'" ~ 20 ~ 8
~ 0
(j
....""c:l
~
='g ~ 0
c:l
~""' ,;,:: 0
Eo-<
=
c.!l
0
0 0
23
92
153
230
~o
>
306
Diquat dichloride (mglkg) Figure 71.3 Dose response curve for the effect of diquat on the water content of the gastrointestinal tract and the haematocrit values 24 h after a single oral dose to the rat. Results are mean ± SE with * representing statistical! y significantly different from control. Water content (_) and packed cel! volume (D). Adapted from Crabtree et al. (1977).
(see later). Whether it contributes to the death of the animal in shock remains to be established. Following sc administration the loss of fluid into the lumen of the gastrointestinal tract is delayed until about 6 days after dosing. An effect on blood volume is seen about 17 h after dosing, which is probably related to the diuretic effect of the compound via this route. At later times the hypovolaemia is less marked than after oral administration (Crabtree et aI., 1977). Histological changes in the gastrointestinal tract following oral dosing are minimal, consisting of patchy loss of keratin from the cardiac end of the stomach, edema of the submucosa at the junction of the glandular and nonglandular region, with some dilation of lacteal and submucosal lymphatic vessels of the small intestine and caecum (Clark and Hurst, 1970; Crabtree et aI., 1977). In monkeys given toxic doses of diquat, the lining of the stomach was ulcerated and the large and small intestine congested; histological examination revealed large areas of necrosis with exfoliation of the lining epithelium (Cobb and Grimshaw, 1979). Studies with anaesthetized rats where diquat was added to ligated segments of jejenum or infused into intestinal loops showed that diquat induced a net secretion of fluid into the lumen (Anton et aI., 1998; Rawlings et al., 1992). This response is seen at sublethal doses and can be blocked by nitric oxide synthase inhibitors, suggesting a role for nitric oxide in the mechanism of fluid secretion (Anton et aI., 1998). The addition of diquat to epithelial cells of the rat small intestine increased the activity of the pentose phosphate pathway and produced NADPH depletion, indicating redox cycling of diquat in these cells. If marked and sustained this may contribute to the fluid loss into the lumen of the gastrointestinal tract (Rawlings et aI., 1994). 71.2.2.14 Effects on the Kidney The major route of elimination for diquat once it has entered the bloodstream is via the kidneys where it is actively secreted by organic cation transport systems (see earlier discussion). At high plasma concentrations diquat produced a mild hydropic change to proximal tubular cells of the Alderley Park rat kidney, which was associated with mild proteinuria and glucosuria (Lock and Ishmael, 1979). More recent studies have shown that
1611
the Fischer 344 rat is more sensitive to diquat-induced liver injury (see later) and this strain is also more sensitive to renal injury (Petry et aI., 1992). Renal tubular necrosis was marked in the cynomolgus monkey where a dose-related renal proximal and distal tubular injury was observed, which was associated with a marked elevation in serum urea (Fig. 71.3) (Cobb and Grimshaw, 1979). Oral administration of diquat to rats reduced urine output and glomerular filtration rate and hence the clearance of diquat. This effect may be secondary to hypovolaemia as both total and renal plasma volume were reduced 24 h after treatment (Lock, 1979). As in other organs, diquat, once it enters a proximal tubular cell, can redox cycle and increase the activity of the pen to se phosphate pathway (Lock and Ishmael, 1979), suggesting that if sufficient diquat was concentrated inside the cell it could overwhelm the defense mechanisms, leading to necrosis. 71.2.2.15 Effects on the Eye Chronic exposure in the diet to diquat produced a dose- and time-dependent appearance of cataracts in both rats and dogs as discussed earlier. This finding was confirmed by Pirie and Rees (1970), who fed rats a diet containing 500 or 750 ppm diquat. They reported that the first change observed in the eye was an irregular "lace-work" of opacity in the posterior cortex, which arose following 4-8 months exposure. The next stage was a clearly defined nuclear cataract that could be seen with the naked eye which progressed to shrinkage and complete opacity. The concentration of diquat that enters the eye at the time of peak blood levels following ip injection is about 0.31 nmol/g of lens and 2.7-4.3 nmol/g eye contents, 1 to 3 h after dosing. The concentration in the eye contents resembling that found in the plasma, while that in the lens was much lower than that in the plasma (Pirie and Rees, 1970). Associated with the development of cataract there was a decrease in the concentration of ascorbic acid in ocular fluid, while the glutathione content remained unchanged (Pirie and Rees, 1970). Studies in vitro using bovine or rat lens have shown that diquat can catalyze a time-dependent loss of ascorbic acid with the formation of hydrogen peroxide (Pirie et aI., 1970) providing evidence that redox cycling of diquat can occur in the eye. Subsequent studies by Bhuyan and Bhuyan (1991, 1994) have provided indirect evidence of diquat-induced production of superoxide anion, hydroxyl radical, and hydrogen peroxide in the rabbit eye following intravitreal injection. Overall, these studies indicate that oxidative stress as a consequence of the redox cycling of diquat is the likely mechanism of cataract formation. Following low dose exposure the onset of cataract will only arise once the defense mechanisms in the eye are overwhelmed. Following acute sc injection diquat causes a prolonged dilatation of the pupil of the eye, suggesting a sympathomimetic action of the compound after high exposure. This response was less marked after oral dosing and not seen after direct application to the eye, indicating it is a systemic-mediated effect (Clark and Hurst, 1970).
1612
CHAPTER 71
Diquat
71.2.2.16 Effects on the Liver Burk and co-workers (1980) reported that diquat was very toxic to selenium deficient rats, and a dose of 5 mg/kg ip produced mortality in 2-3 h. The rats had extensive liver and kidney necrosis and exhaled large quantities of ethane, a marker of lipid peroxidation. By contrast, normal rats given 20 mg/kg ip diquat showed little or no liver injury (Table 71.3). Toxic doses of diquat to cynomolgus monkeys elevated plasma alanine aminotransferase (ALT) and aspartate aminotransferase (AST) (Fig. 71.2) which was associated with minimal histological evidence of hepatic single cell necrosis and sinusoidal congestion (Cobb and Grimshaw, 1979). The discovery that Fischer 344 rats are more susceptible to hepatic necrosis than Sprague-Dawley rats stimulated interest in the mechanism of hepatotoxicity and the role of oxidative stress (Smith et aI., 1985). Large ip doses of diquat administered to Fischer 344 rats elevated plasma ALT and AST values (Table 71.3) and produced hepatocyte necrosis. Diquat also increased the biliary efflux of GSSG and decreased hepatic glutathione content at early times after dosing (Smith et aI., 1985). Treatment of rats with 1,3-bis(2-chloroethyl)-N-nitrosourea (BCNU), an inhibitor of glutathione reductase, followed by diquat increased the efflux of GSSG into bile and potentiated the liver toxicity, relative to control animals, indicating a key role for this enzyme in protecting the liver against oxidative stress (Smith, 1987a). NADPH-cytochrome P-450 reductase catalyzes the reduction of diquat to form a diquat cation radical which can release iron from ferritin both in vitro and in vivo (Reif et al., 1988; Thomas and Aust, 1986). The availability of free iron presumably contributes to the free radical mediated lipid peroxidation seen in the liver (Burk et aI., 1980, 1995; Smith, 1987b;
Wolfgang et aI., 1991). Diquat also enhances the biliary excretion of nonhaem iron (Benzick et aI., 1994; Gupta et aI., 1994. Whether this is a result of intracellular iron overload and hence a clearance mechanism requires further study. Many of the findings reported above with diquat also occur in isolated hepatocytes or liver slices where manipulation of the system has enabled a better understanding of the mechanism of toxicity (DeGray et aI., 1991; Eklow-Lastbom et aI., 1986; Nakagawa et aI., 1992; Rikans and Cai, 1993; Sandy et aI., 1986, 1987; Wolfgang et aI., 1991). Overall, these studies indicate that high doses of diquat can overwhelm the hepatocytes' defense mechanisms against oxidative stress, leading to necrosis. 71.2.2.17 Effects on Other Organs Diquat administration to rats prevented the normal depletion of liver glycogen in fasted animals and produced a marked increase in blood glucose that appeared to be mediated by the adrenal gland, since adrenalectomy prevented these changes (Rose et aI., 1974b). These effects seen following both diquat and paraquat are thought to be due to catecholamine release and high circulating concentrations of corticosteroids (Rose et aI., 1974b). Subsequent studies confirmed that diquat administration to rats produced a dose-related increase in plasma corticosteroid concentration and further confirmed that this could be blocked by pretreating the rats with dexamethasone, which also reduced the concentration of circulating adrenocorticotrophic hormone (ACTH). It was concluded, based on both in vivo and in vitro studies that the increase in adrenal steroid synthesis was due to the release of ACTH from the pituitary (Crabtree and Rose, 1976). The high circulating corticosteroid concentration may account for the changes reported in the thymus, spleen and
Table 71.3 Liver Injury Produced by High Parenteral Doses of Diquat in Normal and Selenium Deficient Rats Plasma marker (U/ml) Dietary state
Strain
Dose (mg/kg)
Normala
Sprague-Dawleyb
0
41 ±5
83 ±8
26d
51 ± 7
199 ± 42
52 e
54± 18
96± 12
0
39±3
135 ± 19
Normala
Fischer
Holzmannc
Normala Selenium
344b
deficien~
Holzmannc
ALT
AST
13
186 ± 25*
105 ± 71
261
833 ± 293*
1063 ± 388*
20
19 ±4
608
41 ±5
5h
3490 ± 1940*
a Rats were given a single intraperitoneal injection of diquat and killed 24 h later unless otherwise stated. bData from Smith et al. (1985). CData from Burk et al. (1980). d Mortality 40% at 24 h. eMortality 80% at 24 h. fMortality 14% at 24 h. g Animals died within 80 ± 12 minutes of dosing. h Animals died within ISO ± 37 minutes of dosing. *Statistically significant from control.
71.3 Toxicity to Humans adrenal gland of rats after a large sc dose of diquat (Clark and Hurst, 1970) and the observed suppression of cell turnover in the lung and eye at early times after dosing (Pirie and Rees, 1970; Smith and Rose, 1977). 71.2.2.18 Treatment of Poisoning in Animals
Diquat like paraquat binds tightly to diatomaceous clay and hence the treatment is the same as that discussed for paraquat, namely purgation of the gastrointestinal tract with bentonite or Fuller's earth (diatomaceous clay's) along with a cathartic such as magnesium sulphate (Clark, 1971; Smith et aI., 1974). See the more detailed discussion on cases of human poisoning.
1613
and the outcome were known, and which were reported to the manufacturer between 1969 and 1996 (Zeneca Agrochemicals, unpublished data). As with the data from the published literature, there is a difference in mortality following accidental ingestion and deliberate ingestion: all 13 patients with accidental or unknown aetiology survived, whereas 11 out of 13 patients with suicidal ingestion died. Most cases of accidental ingestion involved decanting of diquat concentrate from its original container into unmarked drinks bottles. There are no reports in the literature of systemic illness or fatalities occurring following dermal exposure to diquat. 71.3.3 USE EXPERIENCE
71.3 TOXICITY TO HUMANS 71.3.1 EXPERIMENTAL EXPOSURE
Following intravenous administration of 1 J..LCi 14C-labelled diquat to six subjects, 61.2 ± 16.0% of the dose was excreted in urine over 5 days (Feldmann and Maibach, 1974). In the same study, 4 J..Lg/cm2 was applied onto the skin, and 0.3 ±0.1 % (corrected for incomplete urine excretion) of the dose was excreted in urine over 5 days. Diquat was the least absorbed of the 12 pesticides studied. Percutaneous absorption was increased to 1.4% when the site of application was occluded, and to 3.8% when the stratum corneum was removed by successive skin stripping (Wester and Maibach, 1985). 71.3.2 ACCIDENTAL AND INTENTIONAL POISONING
There is a relatively small number of reports on human diquat poisoning in the literature. The first case was reported by Oreopoulos and McEvoy (1969) and involved an 18 year old man who accidentally ingested a mouthful of diquat concentrate which had been decanted into a Coca Cola bottle. This patient survived after treatment with forced diuresis. A further 11 cases of systemic poisoning by ingestion have been reported since then with varying levels of detail (Table 71.4). The overall mortality was 8 out ofthe 12 cases (67%); however, because of the small number of cases, this figure must be treated with caution. Two of the 3 patients who accidentally ingested diquat survived. The third patient, a 2 1/2 year old boy, died 6 days after ingesting an unknown amount of diquat concentrate which had been decanted into a soft drink bottle. In contrast, 7 of the 9 patients who ingested diquat with suicidal intent died. In addition, Okonek (1976) mentioned another two fatal diquat intoxication's as a consequence of suicidal ingestion, however, no further information was given regarding these patients. Hall (1995) found that in the 10 year period from 1983-1992 only two diquat-related fatalities were reported by poison control centres in the USA. Table 71.5 gives details of 26 additional diquat poisoning cases where the approximate ingested dose
Diquat levels in air after tractor and manual spraying were determined by Makovskii (1972). The application rates were 1.0-1.3 kg diquatlha. The highest diquat concentrations (mean 0.56 mg/m 3 ) were found in the tractor cabin when the door was open and spraying was in progress in the direction of the wind. Spraying against the wind and manual spraying resulted in lower concentrations (mean 0.17 and 0.25 mg/m 3 , respectively). The diquat concentration in the air decreased rapidly within 10-20 min after completion of the treatment. The dermal exposure of the spraymen ranged from 0.05 mg to 0.08 mg on the face and hands after 2-3 h of daily work. No health effects were reported. In a study of diquat exposure during aerial application (Sawinsky and Pasztor, 1977), the average diquat concentration in the breathing zone of the pilot was 4.5 J..Lg/m 3 . The potential dermal exposure was estimated from filter discs as 61.5 J..Lg/lOO cm 2. No diquat was detected in urine samples of pilots who had sprayed diquat for 3 to 4 weeks. In contrast, potential dermal exposure of mechanics and loading personnel ranged from 3.5-8.7 mg/lOO cm2, and average urine concentration of diquat was measured as 6.3 and 19.6 J..Lg/100 ml, respectively. Wojeck et al. (1983) studied the exposure of workers applying diquat by hand-operated sprayer against water hyacinths or using direct injection into the water for hydrilla control. There was no measurable inhalation exposure. Average dermal exposure of spraymen and airboat drivers was 1.82 and 0.20 mg/h, respectively, during the treatment of water hyacinths. Average dermal exposure of spraymen and mixers of diquat for the treatment of hydrilla was 0.17 and 0.47 mg/h, respectively. No diquat could be detected in urine. Topical effects from exposure to diquat have been described. Inflammation and nose-bleeds were observed in people handling crystalline powder (Clark and Hurst, 1970). Epistaxis in the field has been seen occurring from splashes when mixing the concentrate or prolonged exposure to spray drift. Cases of nail damage from contact with a concentrated paraquatldiquat mixture were first reported by Samman and 10hnston (1969). According to Clark and Hurst (1970), contact of the 20% diquat concentrate with the nail base may result in nail growth disturbances, development of coloured spots and white bands,
1614
CHAPTER 71
Diquat
Table 71.4 Details of Published Cases of Human Diquat Poisoning by Oral Ingestion Calculated ingested dose
Age (years)
Sex
Dose stateda
18
M
I mouthful
25
M
2-3 mouthful
43
F
Unknown
(mg/kg)h 50 100-150
Aetiology
OutcomeC
Reference
Accident
Survival
Oreopoulos and McEvoy, 1969
Suicide
Death
Schonbom et aI., 1971
(7 days)
Suicide
Death
Okonek and Hofmann, 1975
(2 days) 53
M
45
M
Unknown
33
M
200ml'
<50
Accident
Survival
Fel et al., 1976
Suicide
Death
Okonek, 1976
Suicide
Death
(2 days) 860
Narita et aI., 1978
(2 days) 16
F
50ml
200
Suicide
Death
Vanholder et aI., 1981
(I day)
60
F
20ml
57
Suicide
Death
Vanholder et aI., 198 I
(5 days) 29
M
I mouthful
23
M
300ml
2.5
M
Unknown
50
Suicide
Survival
Ferguson et aI., 1983
860
Suicide
Death
McCarthy and Speth, 1983
Accident
Death
(I day)
Powell et aI., 1983
(6 days) 33
M
300ml
860
Suicide
Survival
Mahieu et aI., 1984
aVolumes (m!) refer to the 20% liquid concentrate, except' which stated that a 30% concentrate was ingested. A volume of 17.5 ml has been used for "a mouthful." b All doses expressed as mg diquat ion per kg body weight. Where the body weight was not explicitly stated, the following assumptions were used: 3-6 years, 25 kg; 7-1 I years, 40 kg; 12-16 years, 50 kg; 17 years and above, 70 kg. cTime interval between ingestion and death indicated in parentheses.
and eventual shedding of the nail. Nonnal nail growth follows upon cessation of exposure. Concentrated diquat fonnulations have also been reported to delay the healing of superficial cuts of the hands of spray workers (IPCS, 1984). Perineal and scrotal burns caused by leaking from a knapsack sprayer containing a paraquat/diquat mixture were seen in two patients (Ronnen et aI., 1995). The lesions responded well to treatment with topical silver sulfadiazine and oral antibiotics, and the damaged skin healed within a few days without scarring. Ocular damage due to exposure to a concentrated paraquat/diquat mixture has been described by Cant and Lewis (1968). This patient had received a splash in his eye and developed conjunctivitis, uveitis, and corneal epithelial damage. Healing was well progressed after 11 days of treatment. In two cases of eye splashes with a paraquat/diquat mixture reported by Nirep et al. (1993), there were delayed corneal epithelial defects (1-2 weeks) with gradual recovery. It was suggested that the surfactants contained in the concentrate may have contributed to the development of this lesion. Diquat is not known to cause cataract in humans. This may be due to a lack of sufficient exposure, but the absence of any ocular signs in poisoned patients may also indicate a true species difference in susceptibility.
71.3.4 ATYPICAL CASES OF VARIOUS ORIGINS
In a case described by Wood et al. (1976), a 45 year old man was admitted to hospital with confusion, high-grade pyrexia, and a 4 day history of productive cough. Chest x-rays showed areas of pneumonic consolidation. Despite antibiotic treatment the patient's condition continued to deteriorate until he was treated with oral prednisone. Shortly thereafter he improved. An episode of diquat spraying preceded the illness, and the patient described being exposed to a cloud of aerosol from a clogged nozzle. No diquat was measured, and no renal damage was seen. The radiological picture was described as being more typical of Laffier's pneumonia. The absence of any facial, upper airways, or eye irritation as well as the atypical picture make a diagnosis of diquat poisoning very doubtful. A 24 year old man was admitted to hospital five hours after exposure to diquat (Williams et aI., 1986). He stated that he had been spraying the herbicide and had experienced a salty taste on the lips and therefore stopped. Two hours later he developed severe abdominal pain and blurred vision and felt generally unwell. A urine test for diquat was positive. Plasma diquat was 0.56 mg/l upon admission, urinary diquat
71.3 Toxicity to Humans
1615
Table 71.5 Cases of Human Diquat Poisoning by Oral Ingestion where Dose and Outcome Were Known (Zeneca Agrochemicals, Unpublished Data) Calculated ingested dose
Age
Country/
(years)
Sex
Dose stateda
36
M
20ml
57
Unknown
F
250ml
715
19
M
200ml
570
Suicide
Death
Canada 1975
Unknown
M
200ml
570
Suicide
Death
Japan 1975
Unknown
F
50ml
140
Accident
Survival
Japan 1977
Unknown
F
300ml
860
Suicide
Death
Japan 1977
55
M
120ml
340
Suicide
Death
USA 1979
(mg/kg)b
OutcomeC
year
Unknown
Survival
Japan 1969
Suicide
Death
Canada 1974
Aetiology
(I day)
21
M
15 ml
43
44
F
150ml
430
Unknown
Survival
UK 1979
Suicide
Death
N Zealand 1981
(I day)
21
M
200ml
570
Suicide
Death
France 1981
(2 days) 41
F
30--40 ml
86-115
Suicide
Survival
France 1984
23
M
300ml
860
Suicide
Died
USA 1981
(I day)
2
F
5ml
lOO
Accident
Survival
15
M
15 ml
60
Accident
Survival
UK 1984 USA 1984
2
M
<5ml
<100
Accident
Survival
USA 1985
29
M
240ml
685
Suicide
Death
USA 1987
(4 days) 21
M
I mouthful
50
Accident
Survival
USA 1988
39
M
<5ml
<14
Accident
Survival
USA 1988
Unknown
F
800ml
2285
Suicide
Death
UK 1989
(I day)
10
I mouthful
88
Accident
Survival
Germany 1989
10
M
I mouthful
88
Accident
Survival
Ireland 1989
52
M
300ml
860
Suicide
Survival
UK 1996
I mouthful
140
Accident
Survival
USA (year
20-30ml
57-86
Accident
Survival
USA (year
<5ml
<100
Accident
Survival
USA (year
180ml
510
Suicide
Death
USA (year
5
unknown) 20
unknown) 0.5
unknown) 50
F
unknown) (5 days) aVolumes (ml) refer to the 20% liquid concentrate. A volume of 17.5 ml has been used for "a mouthful." b All doses expressed as mg diquat ion per kg body weight. Where the body weight was not explicitly stated, the following assumptions were used: 0-2 years, 10 kg; 3-6 years, 25 kg; 7-11 years, 40 kg; 12-16 years, 50 kg; and 17 years and above, 70 kg. cTime interval between ingestion and death indicated in parentheses.
was 52 mg/l, and diquat in gastric contents was 74 mg/l. The patient developed a degree of polyuric renal failure which resolved spontaneously. Initial treatment consisted of gastric lavage, Fuller's Earth, activated charcoal, and mannitol. He also received two haemoperfusions. The patient survived and was asymptomatic at follow-up 6 months later. Although he denied that the exposure was intentional, the severity of poi-
soning and the clinical course raise significant doubts on the description of what appeared to be a minor occupational exposure. A 72 year old farmer with a history of diabetes and transient right-sided hemiparesis developed erythema of the skin with hyperkeratosis and conjunctivitis after exposure of the hands to a 10% diquat solution for about 10 minutes (Sechi et aI.,
1616
CHAPTER 71
Diquat
1992). About 10 days later he developed akathisia with moderate hyperexcitability and insomnia. Over a period of 5 days he became dysphonic, bradykinetic, and rigid. Treatment with carbidopa/levidopa and bromocriptine significantly improved his symptoms. An MRI scan 4 months after the onset of the illness showed small, multiple, bilateral, symmetric areas of high signal intensity in the caudate nuclei and putamen and in the white matter near the ventricular wall. The authors suggested a causal relationship to diquat exposure. However, there were no clinical signs suggestive of systemic diquat poisoning, nor has there been anywhere else a description of Parkinson-like illness after diquat exposure.
However, in contrast to paraquat poisoning, pulmonary fibrosis has not been seen after diquat poisoning. 71.3.5.3 Fulminant Poisoning All organ systems can be affected and death occurs in the majority of patients within 1-2 days. Initial signs are extensive vomiting and diarrhea with massive fluid loss. Typically, patients develop pulmonary edema, acute liver and kidney failure, cardiac arrhythmia, and coma. Death occurs from multiple organ failure. 71.3.6 LABORATORY FINDINGS
71.3.5 CLINICAL FINDINGS AND DOSAGE RESPONSE It is possible from a review of the published literature (Fel et aI., 1976; Ferguson et aI., 1983; Mahieu et aI., 1984; McCarthy and Speth, 1983; Narita et aI., 1978; Okonek, 1976; Okonek and Hofmann, 1975; Oreopou10s and McEvoy, 1969; Powell et aI., 1983; SchOnborn et aI., 1971; Vanholder et aI., 1981) and unpublished cases to distinguish three categories of severity of diquat poisoning and correlate them to the amount ingested (Table 71.6). This is similar to the situation in paraquat poisoning, however, important differences exist both in terms of the clinical presentation and prognosis. Furthermore, since in the case of diquat poisoning this categorization is based on a relatively small number of cases, the dose-response relationship is less certain. Following ingestion, nausea, vomiting, abdominal pain, and diarrhea (often bloody), may occur. Ulcerations of mouth, lips, and back of the throat have also been seen as a result of the caustic action of diquat which may lead to oesophageal and intestinal ulceration within 24-48 h after swallowing.
71.3.5.1 Mild Poisoning In addition to the localized effects on mucous membranes, urea and creatinine may be elevated as a sign of transient renal functional impairment. Patients will make a full recovery regardless of whether treatment is given. 71.3.5.2 Moderate to Severe Poisoning Depending on the amount ingested, the clinical course can be protracted over several weeks. Although nausea, vomiting, and diarrhea may persist for 2-3 days, there may sometimes be a an asymptomatic period, extending for up to 48 hours (Vanholder et aI., 1981). Intestinal paralysis and fluid loss may lead to abdominal distension, tissue dehydration, and hypotensive shock. Within 3-4 days a progressive decline of renal function may occur and continue into complete anuric renal failure. Evidence of reversal may show after 7-10 days. Severe neurologic and neuropsychiatric complications due to brain stem infarction and/or intracranial hemorrhage have been described. In this group of patients, death may occur if treatment is delayed or inadequate.
In cases of diquat poisoning, laboratory findings are generally nonspecific. The changes seen reflect organ failure, affecting in particular kidneys and liver. A rise in serum creatinine and urea is frequently found, although it is not a very sensitive parameter of renal functional impairment. In a case described by Mahieu et al. (1984) a 33 year old man who was said to have ingested about 300 ml of diquat concentrate showed little variation in serum creatinine over the course of 2 weeks. However, evidence of renal damage was seen by a massive increase in albuminuria on day 3 and 4 after the intoxication, accompanied by a rise in urinary excretion of retinol binding protein and beta-2-microglobulin. All parameters had returned to normal by day 15. Increases in liver enzymes such as ALT and AST are a reflection of hepatocellular necrosis which can be seen at autopsy (Schonborn et aI., 1971). They also occur after 3-4 days, although maximum values may be delayed as much as 10-12 days (Mahieu et aI., 1984). A rise in serum bilirubin is a reflection of intracellular cholestasis. Thrombopenia without accompanying changes in other blood parameters has been described after diquat poisoning. In the case reported by SchOnborn et al. (1971) this was severe enough (4000 platelets/mm3 ) to cause multiple bleeding and may have contributed to the brain stem hemorrhage which was seen upon autopsy. This patient received hemodialysis, but not hemoperfusion. Mahieu et al. (1984) found a reduction in platelet count in their patient to 64,000 mm 3 on day 6 with a return to normal values by day 12. No signs of intoxication were seen. Many of the analytical methods described for the determination of paraquat are also applicable to diquat. These include the dithionite spot test (Berry and Grove J, 1971; Widdop, 1976) which, although less sensitive than for paraquat, gives a green color in the presence of diquat. The improved spot test using extraction with a silica cartridge gives a detection limit between 0.5 and 2 J.ig/ml for diquat in plasma (Woollen and Mahler, 1987). A variation of the quantitative method with a spectrophotometric endpoint has been described to determine diquat in plasma (Williams et aI., 1986). The high performance liquid chromatography method described by Gill et al. (1983) can also be used for measuring diquat. The significance of diquat plasma concentrations in terms of the prognosis has not
71.3 Toxicity to Humans
1617
Table 71.6 Severity Grade, Aetiology, and Outcome in 38 Cases of Human Diquat Poisoning by Oral Ingestion (Compiled from Tables 71.4 and 71.5)
Severity
Dose
No. of
(mglkg)
Cases
Accidents
Suicides
No.
(0/0)
No.
(0/0)
Fatalities No.
(0/0)
Mild
:::::50
sa
3
(60)
(20)
0
(0)
Moderate-severe
50--200
15 a
11
(73)
3
(20)
3
(20)
Fulminant
~200
18
0
(0)
18
(100)
16
(89)
38 b
14
(37)
22
(58)
19
(50)
All cases
QIncludes one case with unknown aetiology. bIncludes two cases with unknown aetiology.
been established. However, in two separate studies with a total of 71 patients ingesting a product containing equal levels of paraquat and diquat, the combined plasma concentrations of the two chemicals were above the predictive line established by Proudfoot et al. (1979) in the patients with fatal outcome, and below in the survivors (Ameno et aI., 1994; Yoshioka et aI., 1992). 71.3.7 ABSORPTION
There is little information on absorption of diquat in humans. There was no difference in serum concentration of paraquat and diquat in the first 24 h after ingestion of a combined herbicide, suggesting a similar absorption of the two cations from the gut (Ameno et aI., 1994; Yoshioka et aI., 1992). However, after 24 h the plasma concentration of diquat was consistently lower than that of paraquat. It has been suggested that this finding may be related to an increased biliary excretion and possibly metabolism of diquat (Ameno et aI., 1994, see below). 71.3.8 DISTRIBUTION
The distribution of diquat and paraquat appears to be similar in humans (Ameno et aI., 1994). Following intravenous administration of trace amounts of 14C-labelled diquat to volunteers, the plasma half-life was 4 h (Feldmann and Maibach, 1974). In the case described by Williams et al. (1986), the peak plasma concentration of diquat was measured in the first sample which was taken approximately 5 h after ingestion. Diquat could no longer be detected in plasma 23 h after ingestion. Powell et al. (1983) found a rise in diquat plasma concentrations between 2 and 10 h after ingestion. They also suggested extensive sequestration of diquat in tissues because of a marked rebound of plasma concentrations shortly after the end of haemoperfusion treatments. Tissue levels of diquat in postmortem samples from five patients are shown in Table 71.7. These show marked interindividual differences in tissue distribution which are most likely related to the time interval between ingestion and death. In the patients from the study by Ameno et al. (1994), who died within 48 h, by far the highest concentration of diquat (excluding bile) was measured in the kidneys. In contrast, patients who
died later showed a smaller variation between tissue concentrations (Powell et aI., 1983; SchOnbom et al., 1971). 71.3.9 METABOLISM
Two metabolites of diquat have been identified in humans (Fuke et aI., 1996). Diquat dipyridone and monopyridone were found in both serum and urine following ingestion of a paraquat/diquat mixture in three patients. Serum concentrations of the metabolites accounted for less than 3% of diquat when serum concentrations of diquat were above 10 !J.g/ml. However, once the diquat concentration had fallen below 1 !J.g/ml, diquat dipyridone reached up to 20% of diquat in the serum of one patient, with diquat monopyridone being considerably lower. In contrast, urinary excretion of the monopyridone was up to 10 times higher than for the dipyridone. The authors suggested that the monopyridone was the primary metabolite which would also be excreted more rapidly because it remained partially ionized. At high serum diquat concentrations, metabolism would make little difference to the elimination kinetics of the parent compound. However, this would be different at lower serum concentrations, possibly accounting for the more rapid elimination of diquat when compared to paraquat. 71.3.10 EXCRETION
Following intravenous administration of radiolabelled diquat to human volunteers, 37.3% of the administered dose was recovered in urine in the first 4 h (Feldmann and Maibach, 1974). A further 17.0% was excreted in urine between 4 and 24 h after administration with a total of 61.2% of the dose recovered over 5 days. In the case described by Mahieu et al. (1984) most of the amount recovered in urine was found on the first day although trace amounts could be detected until day 13. Ameno et al. (1994) found the biliary concentration of diquat about 3.5 times higher than that of paraquat after ingestion of a product containing equal amounts of the two herbicides. These authors suggested that biliary excretion was a significant route of diquat elimination which could partially explain the significantly lower plasma concentration of diquat compared to paraquat after 24 h following ingestion.
1618
CHAPTER 71
Diquat
Table 71.7 Diquat Tissue Concentrations in Postmortem Samples of Five Patients
(~g/g)
Ameno et aI., 1994 Reference
Schiinborn et aI., 1971
PoweII et aI., 1983
Case 4
Case 5
Case 6
ingestion
7 days
6 days
26 hours
46 hours
18 hours
Brain
-
0.03
0.23
0.11
Lung
0.56
0.06
0.32
0.32
Heart
0.11
<0.01
0.18
Liver
0.33
0.15
0.34
0.12
Spleen
1.04
0.46
0.31
1.20
0.26
0.05
0.56
Time after
,
Pancreas Kidney
0.23 1.68 0.51 2.48
0.04
3.36
2.40
4.34
0.1
0.26
0.07
0.40
Fat
0.08
0.01
0.15
Blood
0.09
0.20
0.59
28.70
5.56
34.00
1.19
Muscle
Bile Large intestine
0.37
Small intestine
0.45
'Indicates not determined.
71.3.11 PATHOLOGY
71.3.12 TREATMENT OF POISONING
Pathological findings upon autopsy in humans fatalities have been described in a number of cases (McCarthy and Speth, 1983; Powell et aI., 1983; SchOnbom et aI., 1971; Vanholder et aI., 1981). Most autopsy reports describe infarction and purpura of the brain stem as a specific complication of diquat intoxication. In particular the pons may show areas of hemorrhage in association with multiple small and sometimes confluent areas of infarction. Necrosis of the capillary walls may be evident. Other cerebral structures appear unaffected. Depending on the amount of concentrate ingested, the oropharynx, oesophagus, and gastrointestinal tract will show areas of hemorrhagic necrosis, sometimes with pseudomembraneous inflammation. Small and large intestines will be distended with fluid accumulation. The kidneys are pale and swollen and show signs of acute tubular necrosis with severe degenerative lesions, formation of eosinophilic necrosis, and exfoliation of necrotic cells into the tubular lumen. Perivascular round cell infiltration has been seen in the cortex. The liver may show fatty generation of the epithelium, lipid storage and vacuolisation of the Kupffer cells, and evidence of intracellular cho1estasis. In contrast to paraquat poisoning, pulmonary findings are of an acute nature, such as hemorrhagic, fibrin-rich edema, localized intraalveolar bleeding, formation of hyaline membranes, and focal bronchopneumonia. Thickening of alveolar walls and accumulation of reactive alveolar epithelial and chronic inflammatory cells may also occur, but the typical fibroblastic proliferation seen in paraquat lung has not ben described after diquat ingestion.
The therapy of diquat intoxication is based on the same principles as described for paraquat poisoning with prevention of absorption and enhanced elimination being the mainstay of the therapy. Because of the absence of pulmonary fibrosis, no specific pathophysiological therapy has been attempted in most cases of diquat poisoning. Gastric lavage and the use of Fuller's Earth, Bentonite, or activated charcoal together with administration of a cathartic have been advocated as an early treatment to minimize absorption from the gastrointestinal tract (Vanholder et aI., 1981). However, these authors have also pointed out that gastric and intestinal decontamination should be performed cautiously because of the risk of perforation, particularly when therapy is delayed. Adsorbent material should be instilled with care during intestinal paralysis since massive sequestration may occur. Because of the massive fluid losses into the gastrointestinal tract and its potential circulatory and renal consequences, special attention must be given to adequate hydration of the patient, if possible under control of the central venous pressure (Vanholder et aI., 1981). Anticoagulants should be administered with great caution because of the risk of brain stem hemorrhage. However, the use of heparin is often inevitable when hemodialysis or hemoperfusion are needed. Forced diuresis has been used to enhance the elimination of diquat (Mahieu et aI., 1984; Oreopoulos and McEvoy, 1969). However, there is no conclusive evidence of its therapeutic value. Extracorporeal hemodialysis was found to be ineffective in removing diquat from the circulation with an average clearance of 3.17 mllmin and a total removal of 0.84 mg diquat during 11.5 h of dialysis (Okonek and Hofmann, 1975).
References
Hemoperfusion with activated charcoal has been suggested as a more effective way of lowering the plasma diquat concentration (Okonek, 1976). Powell et al. (1983) showed that use of a polystyrene resin cartridge did not remove diquat, but charcoal haemoperfusion achieved clearance rates between 39 and 104 ml/min. However, a slow but marked rebound in diquat concentrations was observed between treatments, indicating extensive tissue sequestration of diquat. Williams et al. (1986) suggested that haemoperfusion was probably ineffective in removing significant amounts of diquat in their patient, despite a lowering of the diquat plasma concentration during the treatment. In conclusion, the treatment of diquat poisoning is directed at preventing gastrointestinal absorption and enhancing elimination from the circulation. There is no conclusive evidence that these therapeutic interventions have contributed significantly to the overall survival of patients. However, in cases of moderate to severe poisoning the prognosis is often favourable, provided the complications of intestinal fluid loss, renal failure and brain stem haemorrhage can be avoided or successfully managed. Disclaimer The positions on certain aspects of the toxicology of diquat in this chapter may not be aligned with the Syngenta positions; the latter are derived mainly from internal Syngenta reports many of which have not been published in the open literature.
REFERENCES Ameno, K, Fuke, c., Shirakawa, Y., and Ogura, S. (1994). Different distribution of paraquat and diquat in human poisoning cases after ingestion of a combined herbicide. Arch. Taxicol. 68, 134--137. Andersen, K J., Leighty, E. G., and Takahashi, M. T. (1972). Evaluation of herbicides for possible mutagenic properties. J. Ag. Fd. Chem. 20, 649. Anderson, D., McGregory, D. B., and Purchase, 1. F. H. (1976). Dominant lethal studies with Paraquat and Diquat in male CD-l mice. Mutat. Res. 40, 347358. Anton, P., Theodorou, v., Fioramonti, J., and Bueno, L. (1998). Low-level exposure to diquat induces a neurally mediated intestinal hypersecretion in rats: involvement of nitric oxide and mast cells. Taxicol. Appl. Pharmacal. 152,77-82. Bainova, A. (1969a). Chronic oral toxicity of bipyridilium herbicides. Hig. Zdrav. 12,325-332. Bainova, A (1969b). Experimental assessment of the effect of dipyridylium herbicides on the skin. Letapisi HEI9, 25-30. Bainova, A, Zlateva, M., and Vulcheva, V. S. (1972). Chronic inhalation toxicity ofbipyridilium herbicides. Hig. Zdrav. 15, 25-31. Bainova, A, and Velcheva, V. S. (1974). Experimental assessment of the effects of dipyridylium on sex glands. In "Works of the Research Institute of Hygiene and Laboratory Protection," Vo!. 22, pp. 111-122. Sofia, Medzina I Fizkultura. Bainova, A. (1975). Cumulative action of Gramoxone and Reglone. In "Problemi na higienta," Vo!. 1, pp. 31-38. Sofia, Medizina I Fizkultura. Bainova, A., and Velcheva, V. S. (1977). Experimental verification of the maximum allowable concentration of Reglone in the air of the workplace. Prabl. Khig.3, 11-18. Bainova, A. 1., and Velcheva, V. S. (1978). Chronic action of Diquat on lungs. Dakl. Balg. Aknd. Nauk 31, 1369-1372. Baldwin, R c., Pasi, A, MacGregor, J. T., and Hine, C. H. (1975). The rates of radical fonnation from the dipyridylium herbicides, paraquat, diquat and
1619
morfamquat in homogenates of rat lung. Kidney and liver: An inhibitory effect of carbon monoxide. Taxical. Appl. Pharmacal. 32, 298-304. Bennett, P. N., and Davies, D. S. (1976). In viva absorption studies with paraquat and diquat in the dog. Br. J. Pharmacal. 58, 284P. Benigni, R. M., Bignami, A, Carere, P., Comba, G., Conti, L., and Conti, R (1979). Mutagenicity studies in salmonalla, streptomyces, aspergillus and unscheduled DNA synthesis in eye cells of paraquat and diquat. Mutat. Res. 64, 127-128. Benzick, A. E., Reddy, S. L., Gupta, S., Rogers, L. K., and Smith, C. V. (1994). Diquat- and acetaminophen-induced alterations of biliary efflux of iron in rats. Biachem. Pharmacal. 47, 2079-2085. Berry, D. J., and Grove, J. (1971). The detennination ofPQ 1,1' -Dimethyl-4,4'bipyridy lium cation in urine. Clin. Chim. Acta 34, 5-11. Bhuyan, K. c., and Bhuyan, D. K (1991). Oxy radicals in the eye tissues of rabbits after diquat in viva. Free Rad. Res. Camm. 12-13, 621---{)27. Bhuyan, D. K, and Bhuyan, K C. (1994). Assessment of oxidative stress to eye in animal model for cataract. Methads Enzymal. 233, 630---{)39. Brian, R. c., Homer, R F., Stubbs, J., and Jones, R. L. (1958). A new herbicide, 1: l-ethylene-2,2' -dipyridylium dibromide. Nature (Londan) 181, 446. Brawn, K, and Fridovich, I. (1981). DNA strand scission by enzymatically generated oxygen radicals. Arch. Biachem. Biaphys. 206,414--419. Burk, R. E, Lawrence, R A., and Lane, J. M. (1980). Liver necrosis and lipid peroxidation in the rat as the result of paraquat and diquat administrationEffect of selenium deficiency. J. Clin. Invest. 65, 1024--1031. Burk, R E, Hill, K E., Awad, J. A., Morrow, J. D., Kato, T., Cockell, K A, and Lyons, P. R (1995). Pathogenesis of diquat induced liver necrosis in selenium deficient rats: Assessment of the roles of lipid peroxidation and selenoprotein P. Hepatalagy 21, 561-569. Bus, J. S., Preache, M. M., Cagen, S. Z., Posner, H. S., Eliason, B. c., Sharp, C. W, and Gibson, J. E. (1975). Fetal toxicity and distribution of paraquat and diquat in mice and rats. Taxical. Appl. Pharmacal. 33, 450460. Cant, J. S., and Lewis, D. R H. (1968). Ocular damage due to paraquat and diquat. Br. Med. J. 3, 59. Chan, B. S. H., Lazzaro, V. A., Seale, J. P., and Duggin, G. G. (1988). The renal excretory mechanisms and the role of organic cations in modulating the renal handling of paraquat. Pharmacal. Therap. 79, 193-203. Clark, D. G., and Hurst, E. W (1970). The toxicity of diquat. Br. J. Ind. Med. 27,51-55. Clark, D. G. (1971). Inhibition of the absorption of Paraquat from the gastrointestinal tract by absorbents. Br. 1 Ind. Med. 28, 186-188. Cobb, L. M., and Grimshaw, P. (1979). Acute toxicity of oral Diquat (1,1'ethylene-2,2' -bipyridinium) in cynomolgus monkeys. Taxicol. Appl. Pharmacal. 51, 277-282. Crabtree, H. C., and Rose, M. S. (1976). Early effects of diquat on plasma corticosteroid concentrations in rats. Biachem. Pharmac. 25, 2465-2468. Crabtree, H. c., Lock, E. A., and Rose, M. S. (1977). Effects of Diquat on the gastrointestinal tract of rats. Taxical. Appl. Pharmacal. 41, 585-595. Daniel, J. W., and Gage, J. C. (1966). Absorption and excretion of diquat and paraquat in rats. Br. J. Ind. Med. 23, 133-136. DeGray, J. A., Rao, D. N. R, and Mason, R P. (1991). Reduction of paraquat and related bipyridylium compounds to free radical metabolites by rat hepatocytes. Arch. Biachem Biaphys. 289, 145-142. Eklow-Lastbom, L., Rossi, L., Thor, H., and Orrenius, S. (1986). Effects of oxidative stress caused by hyperoxia and diquat a study in isolated hepatocytes. Free Radical Res. Cammun. 2, 57---{)8. FAOIWHO (1978). Diquat. In "1977 Evaluations of Some Pesticide Residues in Food." Food and Agricultural Organization of the United Nations, Rome. FAOIWHO (1993). Diquat. In "Evaluation 1993, Part II-Toxicology, Pesticide Residues in Food." Food and Agricultural Organization of the United Nations, Rome. Fel, P., Zela, 1., Szule, E., and Varga, L. (1976). Reglone diquat-dibromide poisoning case successfully treated by haemodialysis. Drv. Hetil. 117, 17731774. Feldmann, R J., and Maibach, H. I. (1974). Percutaneous penetration of some pesticides and herbicides in man. Taxicol. Appl. Pharmacal. 28, 126-132.
1620
CHAPTER 71
Diquat
Ferguson, A. H., Jacobsen, J. B., and Nielsen, H. (1983). Severe diquat poisoning. Ex. Med. Toxicol. 1, 778. Fuke, c., Ameno, K., Ameno, S., Kinoshita, H., and Ijiri, 1. (1996). Detection of two metabolites of diquat in urine and serum of poisoned patients after ingestion of a combined herbicide of paraquat and diquat. Arch. Toxicol. 70, 504-507. Gage, J. C. (1968a). Toxicity of paraquat and diquat aerosols generated by a size-selective cyclone. Effect of particle size distribution. Br. J. Ind. Med. 25,304-314. Gage, J. C. (1968b). The action of paraquat and diquat on the respiration of liver cell fractions. Biochem. J. 109,757-761. Gaines, T. B., and Linder, R. E. (1986). Acute toxicity of pesticides in adult and weanling rats. Fundam. Appl. Toxicol. 7, 299-308. Gill, R., Qua, S. c., and Moffat, A. C. (1983). High-performance liquid chromatography of paraquat and diquat in urine with rapid sample preparation involving ion-pair extraction on disposable cartridges of octadecyl-silica. 1. Chromat. 255, 483-490. Gupta, S., Rogers, L. K., and Smith, C. v. (1994). Biliary excretion of lysosomal enzymes, iron and oxidised protein in F344 and Sprague-Dawley rats and the effects of diquat and acetaminophen. Toxicol. Appl. Pharmacol. 125, 42-50. Hall, A. H. (1995). Paraquat and diquat exposures reported to U.S. poison centers 1983-1992. In "Paraquat Poisoning" (c. Bismuth and A. H. Hall, eds.), pp. 53-63. Dekker, New York. Hughes, RD., Millburn, P., and Williams, R T. (1973). Biliary excretion of some diquaternary ammonium cations in the rat, guinea-pig and rabbit. Biochem. J. 136, 979-984. IPCS (1984). "Paraquat and Diquat." Environmental Health Criteria 39, World Health Organization. Keeling, P. L., Pratt, 1. S., Aldridge, W. N., and Smith, L. L. (1981). The enhancement of paraquat toxicity in rats by 85% oxygen-lethality and cellspecific lung damage. Br. J. Exp. Path. 62, 643-654. Keeling, P. L., and Smith, L. L. (1982). Relevance of NADPH depletion and mixed disulphide formation in rat lung to the mechanism of cell damage following paraquat administration. Biochem. Pharmacol. 31, 3243-3249. Kehrer, J. P., Haschek, w., and Witschi, H. P. (1979). The influence of Hyperoxia on the acute toxicity of paraquat and diquat. Drug Chem. Tox. 2, 397-408. Khera, K. S., Whitta, L. L., and Clegg, D. J. (1970). Embryopathic effects of diquat and paraquat in rats. Ind. Med. Surg. 37, 257-261. Kurisaki, E., and Sato, H. (1979). Tissue distribution of paraquat and diquat after oral administration in rats. Forensic Sci. Int. 14, 165-180. Lam, H. E, Takezawa, J., Gupta, B. N., and Van Stee, E. W. (1980). A comparison of the effects of paraquat and diquat on lung compliance, lung volumes and single breath diffusing capacity in the rat. Toxicology 18, 111-123. Larsson, B., Oskarsson, J. A., and Tjalve, H. (1977). Binding of paraquat and diquat in melanin. Exp. Eye Res. 25, 353-359. Levin, D. E., Hollstein, M., Christman, M. E, and Ames, B. K. (1982). A new salmonella tester strain (TA 102) with A.T. base pairs at the site of mutation detects oxidative mutagens. Proc. Natl. Acad. Sci. 79, 7445-7449. Litchfield, M. H., Daniel, J. w., and Longshaw, S. (1973). The tissue distribution of the bipyridylium herbicides diquat and paraquat in rats and mice. Toxicology 1, 155-165. Lock, E. A. (1979). The effect of paraquat and diquat on renal function in the rat. Toxicol. Appl. Pharmacol. 48, 327-336. Lock, E. A., and Ishmael, J. (1979). The acute effects of paraquat and diquat on the rat kidney. Toxicol. Appl. Pharmacol. 50, 67-76. Mahieu, P., Bonduelle, Y., Bernard, A., De Cabooter, A., Gala, M., Hassoun, A., Keonig, J., and Lauwerys, R (1984). Acute diquat intoxication interest of its repeated determination in urine and the evaluation of renal proximal tubule integrity. J. Toxicol. Clin. Toxicol. 22, 363-369. Makovskii, V. N. (1972). "Toxicological and Hygiene Studies of the Bipyridilium Herbicide Diquat and Paraquat." Ph.D. Thesis, Vinniza, USSR. Manabe, J., and Ogata, T. (1986). The toxic effect of diquat on the rat lung after intratracheal administration. Toxicol. Lett. 30, 7-12. Manahe, J., and Ogata, T. (1987). Lung fibrosis induced by diquat after intratracheal administration. Arch. Toxicol. 60,427-431.
Matsuura, N., Takinami, M., Kurisaki, E., and Satoo, O. (1978). Distribution of paraquat dichloride and diquat dibromide in the living body. Fukishima Igakkai Zasshi 28,212-215. McCarthy, L. G., and Speth, C. P. (1983). Diquat intoxication. Ann. Emerg. Med. 12,294-396. Nakagawa, Y., Moldeus, P., and Cotgreave, 1. (1992). The S-Thiolation of hepatocellular protein thiols during diquat metabolism. Biochem. Pharm. 43, 2519-2525. Narita, S., Matojuku, M., Sato, J., and Mori, H. (1978). Autopsy in acute suicidal poisoning with diquat dibromide. Nippon Igakkai Zasshi 27, 454-455. Nirep, M., Hayasaka, S., Nagata, M., Tamap, A., and Tawara, T. (1993). Ocular injury caused by Preeglox-L, a herbicide containing paraquat, diquat and surfactants. Jpn. J. Ophthalmol. 37, 43-46. Okonek, S. (1976). Poisoning by paraquat or diquat. Med. Welt. 27,1401-1404. Okonek, S., and Hofmann, A. (1975). On the question of extracorporeal haemodialysis in diquat intoxication. Arch. Tox. 33, 251-257. Oreopoulos, D. G., and McEvoy, J. (1969). Diquat poisoning. Postgrad. Med. J. 45, 635-637. Pasi, A., Embree, J. W., Eisenlord, G. H., and Hine, C. H. (1974). Assessment of the mutagenic properties of diquat and paraquat in the murine dominant lethal test. Mutat. Res. 26, 171-175. Petry, T. w., Wolfgang, G. H. 1., Jolly, R. A., Ochoa, R, and Donarski, W. J. (1992). Antioxidant-dependant inhibition of diquat-induced toxicity in vivo. Toxicology 74, 33-43. Pirie, A., and Rees, J. R (1970). Diquat cataract in the rat. Exp. Eye Res. 9, 198-203. Pirie, A., Rees, J. R, and Holmberg, N. J. (1970). Diquat cataract: Formation of the free radical and its reaction with constituents of the eye. Exp. Eye Res. 9,204-218. Powell, D., Pond, S. M., Alien, T. B., and Portale, A. A. (1983). Hemoperfusion in a child who ingested diquat and died from pontine infarction and hemorrhage. 1. Toxicol. Clin. Toxicol. 20, 405-420. Proudfoot, A. T., Stewart, M. S., Levitt, T., and Widdop, B. (1979). Paraquat poisoning: Significance of plasma-paraquat concentrations. Lancet 11, 330332. Pushkar, M. S. (1969). Morphological changes in the organism under the effect of the herbicide reglone (diquat). Vrach. Delo. 9, 92-96. Rawlings, J. M., Foster, J. R, and Heylings, J. R (1992). Diquat-induced intestinal secretion in the anaesthetised rat. Human Exp. Toxicol. 11,524-529. Rawlings, J. M., Wyatt, 1., and Heylings, J. R. (1994). Evidence for redox cycling of diquat in rat small intestine. Biochem. Pharmacol. 47, 1271-1274. Reif, D. w., Beales, 1. L. P., Thomas, C. E., and Aust, S. D. (1988). Effect of diquat on the distribution of iron in rat liver. Toxicol. Appl. Pharmacol. 93, 506-510. Repine, J. E., Pfinninger, O. S., Talmage, D. W., Berger, E. M., and Pettijohn, D. E. (1981). Dimethyl sulphoxide prevents DNA nicking mediated by ionising radiation or iron/hydrogen peroxide-generated hydroxyl radical. Proc. Natl. Acad. Sci. U.SA 78, 1001-1003. Rikans, L. E., and Cai, Y. (1993). Diquat-induced oxidative damage in BCNUpretreated hepatocytes of mature and old rats. Toxicol. Appl. Pharmacol. 118, 263-270. Ronnen, M., Klin, B., and Suster, S. (1995). Mixed diquat/paraquat-induced burns. Int. J. Dermatol. 34, 23-25. Rose, M. S., and Smith, L. L. (1977). The relevance of paraquat accumulation by tissues. In "Biochemical Mechanisms of Paraquat Toxicity" (A. P. Autor, ed.), pp. 71-91. Academic Press, San Diego. Rose, M. S., Smith, L. L., and Wyatt, 1. (l974a). Evidence for the energydependant accumulation of paraquat into rat lung. Nature 252,314-315. Rose, M. S., Crabtree, H. c., Fletcher, K., and Wyatt, 1. (1 974b). Biochemical effects of diquat and paraquat: Disturbance of the control of corticosteroid synthesis in rat adrenal and subsequent effects in the control of liver glycogen utilisation. Biochem. J. 138,437-443. Rose, M. S., Lock, E. A., Smith, L. L., and Wyatt, 1. (1976a). Paraquat accumulation. Tissue and species specificity. Biochem. Pharmacol. 25,419-423. Rose, M. S., Smith, L. L., and Wyatt, 1. (l976b). The relevance of pentose phosphate pathway stimulation in rat lung to the mechanism of Paraquat toxicity. Biochem. Pharmacol. 25, 1763-1767.
References
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Thomas, C. E., and Aust, S. D. (1986). Reductive release of iron from ferritin by cation free radicals of paraquat and other bipyridyls. 1. BioI. Chem. 261, 13064-13070. Tomita, M. (1991). Comparison of one-electron reduction activity against the bipyridylium herbicides, paraquat and diquat, in microsomal and mitochondrial fractions ofliver, lung and kidney (in vitro). Biochem. Pharmacol. 42, 303-309. Vanholder, R., Colardyn, F., De Reuck, J., Praet, M., Lameire, N., and Ringoir, S. (1981). Diquat intoxication-Report of two cases and review of the literature. Am. J. Med. 70, 1267-1271. Verbetskii, V. E., and Stolyarchuk, A. A. (1967). Toxicological and pharmacological properties of herbicides which are bipyridine derivativesGramoxone and Reglone. Vop. Gig. Toksikol. Pestits. 164-168. Verbetskii, V. E., and Pushkar, M. s. (1968). Pathological changes in the organs of animals on acute poisoning with the herbicide diquat (Reglone). In "Some Questions Concerning Human and Animal Morphology," pp. 51-52. Medizina, Odes sa. Vulimiri, S. v., Gupta, S., Smith, C. v., Moorthy, B., and Randerath, K. (1995). Rapid decrease in indigenous covalent DNA modifications (I-Compounds) of male Fischer-344 rat liver DNA by diquat treatment. Chem.-Biol. Inter-
act. 95, 1-16. Wester, R. c., and Maibach, H. I. (1985). In vivo percutaneous absorption and decontamination of pesticides in humans. J. Toxicol. Environ. Hith. 16, 2537. Widdop, B. (1976). Detection of Paraquat in urine. Br. Med. J. 2, 1135. Williams, P. F., Jarvie, D. R., and Whitehead, A. P. (1986). Diquat intoxication: Treatment by charcoal haemoperfusion and description of a new method of diquat measurement in plasma. J. Toxicol. c/in. Toxicol. 24, 11-20. Witschi, H., Kacew, S., Hirai, K. I., and Cote, M. G. (1977). In vivo oxidation of reduced nicotinamide adenine dinucleotide phosphate by paraquat and diquat in rat lung. Chem. BioI. Interact. 19, 143-160. Wojeck, G. A., Price, J. F., Nigg, H. N., and Stamper, J. H. (1983). Worker exposure to paraquat and diquat. Arch. Environ. Contam. Toxicol. 12, 6570. Wolfgang, G. H. I., Jolly, R. A., Donarski, W. J., and Petry, T. W. (1991). Inhibition of diquat induced lipid peroxidation and toxicity in precision cut rat liver slices by novel antioxidants. Toxicol. Appl. Pharmacol. 108,321-329. Wood, T. E., Edgar, H., and Salcedo, J. (1976). Recovery from inhalation of diquat aerosol. Chest 70, 774-775. Woollen, B. H., and Mahler, J. D. (1987). An improved spot-test for the detection of paraquat and diquat in biological samples. CUn. Chim. Acta 167, 225-229. Yoshioka, T., Sugimoto, T., Kinoshita, N., Shimazu, T., Hiraide, A., and Kuwagata, y. (1992). Effects of concentration reduction and partial replacement of paraquat by diquat on human toxicity. A clinical survey. Human Exp. Toxicol. 11,241-245.
CHAPTER
72 Phenoxy Herbicides (2,4-D) Elke Kennepohl and Ian C. Munro Cantox Health Sciences International
72.1 INTRODUCTION Phenoxy herbicides have been commercially available for over 50 years and are the most widely used family of herbicides worldwide. 2,4-Dichlorophenoxyacetic acid (2,4-D), the most common of the phenoxy herbicides, is one of the best-studied agricultural chemicals. This chapter focuses primarily on 2,4-D since it is the most widely used herbicide and the majority of the literature on phenoxy herbicides pertains to studies with 2,4-D. The safety of using phenoxy herbicides was first questioned when a series of case-control studies was published by Lennart Hardell in the late 1970s, in which he hypothesized that the occurrence of three rare forms of cancer (Hodgkin's disease, soft tissue sarcoma, and non-Hodgkin's lymphoma) in workers was related to exposure to these herbicides along with dioxins known to contaminate 2,4,5-trichlorophenoxyacetic acid (2,4,5-T). Since that time, several human and animal studies have been conducted which do not lend support to his hypothesis. As well, several expert panels have been convened to assess the safety of 2,4-D, and all have concluded that there is no evidence to suggest that 2,4-D poses any risk to human health under its intended conditions of use. In fact, 2,4-D has been classified by the U.S. Environmental Protection Agency (EPA) as a Group D (not classifiable as to human carcinogenicity) because "the evidence is inadequate and cannot be interpreted as showing either the presence or absence of a carcinogenic effect." Because of the vast amount of data available on 2,4-D, this chapter provides a brief summary and overview of the available studies.
72.2 PHYSICAL AND CHEMICAL PROPERTIES Several phenoxy acids have been used as herbicides, including 2,4,5-T, 4-(2,4-dichlorophenoxy) butyric acid (2,4-DB), 2-(2,4-dichlorophenoxy propionic acid) (dichlorprop), 2-(2methyl-4-chlorophenoxy) propionic acid (MCPP or mecoprop), 2-methyl-4-chlorophenoxyacetic acid (MCPA), and 2-(2,4,5trichlorophenoxy) propionic acid (Silvex), with the most comHandbook of Pesticide Toxicology Volume 2. Agents
monly and widely used herbicide being 2,4-D. 2,4,5-T and Silvex are no longer manufactured or sold. Figure 72.1 shows the chemical structures of the phenoxy acids. 72.2.1 2,4-D ACID, SALTS, AND ESTERS The basic form of 2,4-D is the acid, but 2,4-D is often formulated as an inorganic salt, amine, or ester through various manufacturing processes, and is used in many commercial products. CAS number: Chemical name: Trade names:
94-75-7 2,4-dichlorophenoxyacetic acid Agrotect, Amidox, Amoxone, AquaKleen, Brush-Rhap, Chloroxone, Crop Rider, Crotilin, Dacamine, Dacamine, Ded-weed LV-69, Dicopur, Dinoxol, DMA-4, Dormone, Emulsamine BK, Envert 171, Esteron, Esteron 99C, Estone, Farmco, Femesta, Femimine, Femoxone, Ferxone, Foredex 75, Formula 40, Hedonal, Hebidal, Ipaner, Krotiline, Lawn-Keep, Macrondray, Miracle, Monosan, Moxone, Netagrone, Pennamine, Phenox, Pielik, Planotox, Plantgard, Rhodia, Salvo, Transamine, Tributon, Trinoxol, Vergemaster, Vidon 638, Visko-Rhap, Weed-Ag-Bar, Weedar-64, Weedatul, Weed-B-Gon, Weedez Wonder Bar, Weedone LV4, Weed-Rhap, Weed Tox, Weedtrol
Appearance:
white powder
Empirical formula: Molecular weight: Melting point: Boiling point: Water solubility: Vapor pressure: Partition coefficient:
CgH6Ch 0 3 221.04
1623
140.soC
130°C at 1 mm Hg (isopropyl ester) 900 mg/liter at 25°C (acid) 0.02 mPa at 25°C (acid) 2.81 Copyright © 200 1 by Academic Press. All rights of reproduction in any form reserved.
1624
CHAPTER 72
Phenoxy Herbicides (2,4-D)
CI--Q--o- CH COO 3-
l!
Cl 2,4-DaCld
Cl
2,4-DB
~ ..--cH3 ~ OCH,COOH-HN_ ,
Dichlorprop
CH
Cl 2,4-D dimemylamine salt
-
IT
r'CH;
Mecoprop
2,4,5-T
C I - q - 0 - C H , - C - O - CH3-CH,HW
Cl
2,4-D 2-ethylhcxyl ester
Cl
~ ~_
'"
r'
O--CH--C()(lH
Cl
C~(}--CH3COOCH'CH30CH'CH3CH'CH'
MCPA
Silvex
Cl 2,4-D bu(oxycthano! ester
Figure 72.1
Chemical structures of phenoxy acid herbicides.
72.2.4 DICHLORPROP (2,4-DP)
72.2.2 2,4,5-T CAS number:
93-76-5
CAS number:
Chemical name:
2,4,5-trichlorophenoxyacetic acid
Trade names:
no longer manufactured or sold
Appearance:
white crystals
Empirical formula:
CsHs Cl 30 3 255.49 156.6°C (pure acid); 150-151°C (technical acid)
Chemical name: Trade names: Appearance: Empirical formula: Molecular weight: Melting point: Water solubility:
Molecular weight: Melting point: Boiling point: Water solubility: Vapor pressure:
decomposes 278 mg/liter at 25°C (acid) 0.022 mm Hg at 25°C
72.2.3 2,4-DB CAS number:
94-82-6
Chemical name: Trade names:
4-(2,4-dichlorophenoxy) butyric acid Butoxone, Butyrac, Butirex, Embutone, Embutox, Legumex, Venceweed
120-36-5 2-(2,4-dich10rophenoxy) propionic acid Cornox, Hedona1, Weedone, Estaprop colorless crystals C9 H SCh 0 3 235.07 117-IlSOC 350 mg/liter at 20°C
72.2.5 MECOPROP (MCPP) CAS number: Chemical name:
7085-19-0 2-(4-chloro-2-methy1phenoxy) propionic acid
Trade names:
Kilprop, Mecopar, Triester-II, MecAmine-D, Triamine-II, Triplet, TriPower, Trimec, Trimec-Encore, U46 KV Fluid white to light brown crystalline solid
Appearance:
colorless to white crystals
Appearance: Empirical formula:
Empirical formula: Molecular weight:
ClO H IOCh 0 3 249.10
Molecular weight: Melting point:
Melting point:
117-119°C
Water solubility:
46 mg/liter at 25°C
Vapor pressure:
negligible (acid and salts)
Water solubility: Vapor pressure: Partition coefficient:
CloHllCI03 214.65 93-95°C very soluble at 25°C 0.31 mPa at 20°C 1.26 at pH 7
72.4 Fonnulations
72.2.6 MCPA CAS number: Chemical name: Trade names:
Appearance: Empirical formula: Molecular weight: Melting point: Water solubility: Vapor pressure:
94-74-6 2-methyl-4-chlorophenoxyacetic acid Agritox, Agroxone, Agrozone, Agsco MXL, Banlene, Blesal MC, Bordermaster, Cambilene, Cheyenne, Chimac Oxy, Chiptox, Class MCPA, Cornox Plus, Dakota, Ded-Weed, Empal, Envoy, Legumex, Malerbane, Mayclene, Mephanac, Midox, Phenoxylene, Rhomene, Rhonox, Sanaphen-M, Sharnrox, Selectyl, Tiller, Vacate, Weed-Rhap, Zhelan colorless crystals C9 H 9CI03 200.62 118-1l9°C 825 mg/liter at 25°C (acid) 0.2 mPa at 20°C
72.2.7 SILVEX CAS number: Chemical name: Trade names: Appearance: Empirical formula: Molecular weight: Melting point: Water solubility:
93-72-1 2-(2,4,5-trichlorophenoxy) propionic acid no longer manufactured or sold white powder C9 H7 Cl}03 269.51 181.6°C 200 mglliter at 25°C
72.3 HISTORY OF USE For over 50 years, 2,4-D has been the most commonly and widely used herbicide throughout the world. When applied to plants, 2,4-D is absorbed through the roots and leaves within 4 to 6 hours and is distributed in the plant via the phloem (WHO, 1984). Once absorbed, 2,4-D selectively eliminates broadleaf plants (due to their larger leaf area and hence, greater absorption) by mimicking the effect of auxins (i.e., plant growthregulating hormones) and stimulating growth, rejuvenating old cells, and overstimulating young cells, leading to an abnormal growth pattern and death in some plants (Mullison, 1987). In addition, 2,4-D affects plant metabolism, which leads to interference with food transport (Mullison, 1987). 2,4-D is primarily used as a herbicide in agriculture, forestry, and lawn care practices, with the majority (>60%) of the total usage in the United States being reported for use as weed control in agriculture (i.e., corn and small grains) (EPA, 1997). 2,4-D is reported to be effective against dandelion, plantain,
1625
chickweed, henbit, white clover, heal-al1, red sorrel, curly dock, chicory, yellow rocket, speedwell, ground ivy, spurge, oxalis, knotweed, purslane, thistle, wild violet, wild onion, wild garlic, lespedeza, yellow nutsedge, crabgrass, sumac, willow, sagebrush, ragweed, Eurasian water milfoil, and water hyacinth (Lefton et aI., 1991; Mullison, 1987; WHO, 1975). To a lesser extent, 2,4-D is used as a growth regulator on various crops ranging from potatoes to citrus fruits (WHO, 1975, 1984). In the past, 2,4-D was combined with 2,4,5-trichlorophenoxyacetic acid (2,4,5-T) for brush and weed control. Over 10 million gallons of a special concentrated mixture called Agent Orange were applied in the Vietnam War to defoliate trees (Wolfe, 1983).
72.4 FORMULATIONS 2,4-D is formulated into end-use products to facilitate application. Water-soluble salts and amines are usually prepared as aqueous solutions with small amounts of additives such as water conditioners and antifoam agents. The oil-soluble esters are often formulated with petroleum solvents (e.g., kerosene or naphtha) plus emulsifiers. Such formulations are then diluted with relatively large amounts of water to make the final herbicide spray mixture. Several forms of 2,4-D can be combined with dry fertilizer ingredients to form lawn "weed and feed" products. In the past, concerns arose regarding possible contamination of2,4-D formulations with dioxins, notably the polychlorinated dibenzo-p-dioxins (PCDDs) and more specifical1y 2,3,7,8tetrachlorodibenzo-p-dioxin (T4CDD), and nitrosamines. 2,4D formulations have been reported to contain 2,3,7,8-T4CDD but only when 2,4,5-T was present (Cochrane et al., 1981, 1982a, 1982b; Woolson et aI., 1972). 2,4-D formulations currently sold in the United States contain very few PCDD contaminants. In fact, analytical studies of the more recent formulations have repeatedly shown dioxin levels to be below the limit of quantitation set by the U.S. Environmental Protection Agency (Berry, 1989; Cramer, 1996). In Canada, a limit of 10 ppb per isomer PCDD (nondetectable for 2,3,7,8-T4CDD at 1 ppb) in 2,4-D formulations has been set (Agriculture Canada, 1983). In an older analysis of 200 samples of various forms of 2,4-D (Cochrane et aI., 1981, 1982a, 1982b), al1 but a few samples tested below the 10-ppb limit. In the past, there was some possible contamination with nitrosamines formed from nitrates used in preserving metal storage containers; however, plastic or epoxy-lining has replaced the metal used for storage containers and nitrosamine formation no longer presents a concern.
72.5 HUMAN EXPOSURE TO 2,4-D 2,4-D is one of the most commonly used herbicides both domestically and commercially, and exposure to it can occur via inhalation, ingestion, and dermal contact. Respiratory exposure
1626
CHAPTER 72
Phenoxy Herbicides (2,4-D)
to 2,4-D is less than 2% of total exposure (Grover et al., 1986), and residual levels in foodstuffs or drinking water are essentially nondetectable or only detected in trace amounts (Duggan and Corneliussen, 1972; Duggan and Lipscomb, 1969; Gartrell et al., 1985). By far, dermal contact during use of the product accounts for the greatest potential for exposure, with estimates that approximately 90% of total exposure occurs through dermal exposure (Feldman and Maibach, 1974). Since 2,4-D use is typically seasonal and short-term, the duration of exposure is considered repeated subchronic. In a 1992 review of 2,4-D, Munro et al. (1992) summarized exposures to 2,4-D in a variety of occupational and home-use settings. Based on several epidemiological studies, the highest exposures were obtained in occupational settings, with reported average estimated internal doses ranging from 0.01 to 40 J.tg/kg body weight/day for forestry workers, and 0.35 to 6.3 J.tg/kg body weight/day for commercial applicators and farmers (Frank et al., 1985; Grover et al., 1986; Lavy and Mattice, 1984; Lavy et al., 1987; Yeary, 1986). Average estimated internal doses reported to be reached by bystanders or home and garden users were below 0.2 J.tg/kg body weight/day (Harris et al., 1992; Lavy and Mattice, 1984). Most of the past epidemiological studies do not reflect the growing trend toward using protective apparel when applying herbicides. With an increased awareness of worker safety and the new proposed labeling directions, workers are required to wear protective clothing consisting of eye protection, chemical-resistant gloves, long-sleeved shirt, long pants, socks, and shoes. In addition, following use of2,4-D, it is recommended that workers thoroughly wash their hands, face, and arms with soap and water and wash any contaminated clothing separately.
72.6 TOXICOLOGICAL STUDIES 72.6.1 ABSORPTION 2,4-D is rapidly absorbed through the gastrointestinal tract following oral exposure, with peak plasma levels being reached in as little as 10 minutes or up to 24 hours depending on the dose and chemical form of 2,4-D (Erne, 1966a; Khanna and Fang, 1966; Knopp and Schiller, 1992; Kohli et al., 1974; Pelletier et aI., 1989; Sauerhoff et al., 1977). The rate of absorption is related to dose, with absorption occurring more rapidly at lower doses (i.e., 0.4 mg/kg body weight/day) than at higher doses (i.e., 1 mg/kg body weight/day) (Pelletier et al., 1989). Absorption of 2,4-D esters has been reported to occur more slowly than for acid or salt forms (Erne, 1966a); however, the excretion rates for the various forms are reported to be similar (Khanna and Fang, 1966; Knopp and Schiller, 1992; Pelletier et aI., 1989). Dermal contact is the major route of exposure to 2,4-D. In occupationally exposed humans, dermal absorption was reported to occur rapidly based on the detection of 2,4-D in urine within 4 hours (Feldman and Maibach, 1974), and although the percentage absorbed is variable, it is usually less than 6 % (EPA,
1996; Feldman and Maibach, 1974; Harris and Solomon, 1992). Studies in rats and monkeys showed these percentages to be highly variable and dependent on chemical form, vehicle, and animal species (Grisson et aI., 1987; Knopp and Schiller, 1992; Moody et aI., 1990, 1991; Pelletier et aI., 1989, 1990). Although no controlled studies have been conducted to assess the absorption rate via inhalation exposure, epidemiological studies of occupationally exposed workers indicate that absorption is rapid by both dermal and inhalation routes (Frank et aI., 1985; Kolmodin-Hedman and Erne, 1980). 72.6.2 DISTRIBUTION 2,4-D is highly water soluble and therefore is widely distributed, but does not accumulate, in the body. It also does not readily cross lipid membranes, and at physiological pH, it exists predominately in the ionized form. 2,4-D uses active transport systems to enter tissues and cross the bloodlbrain barrier (Kim and O'Tuama, 1981; Pritchard, 1980). Another factor which contributes to the extent of tissue distribution of 2,4-D is its ability to bind to serum proteins (Erne, 1966a; Fang and Lindstrom, 1980; Orberg, 1980). Peak tissue levels in rats have been reported anywhere from 10 minutes to 8 hours depending on the dose administered (0.4 to 240 mg/kg body weight) (Khanna and Fang, 1966; Pelletier et aI., 1989). Following exposure, 2,4-D has been detected in liver, kidney, and lung of a variety of animal species (Clark et aI., 1975; Erne, 1966a). Levels in brain were reported to account for only a very small percentage of the exposure dose (Erne, 1966a; Tyynela et aI., 1990); however, at levels of intoxication (i.e., 300 mg/kg body weight, which is well above the level of renal saturation), levels in brain and cerebrospinal fluid of rats were increased relative to plasma levels (Elo and Ylitalo, 1977,1979; Tyynela et aI., 1990). At these high dose levels, the organic acid transport system responsible for the efflux of 2,4-D out of the brain is inhibited (Kim et aI., 1983; Pritchard, 1980; Tyynela et aI., 1990; Ylitalo et aI., 1990). In addition, vascular damage has been reported in rats administered extremely high doses of 2,4-D (i.e., more than 300 mg/kg body weight) (Elo et aI., 1988), which may facilitate an increased influx of 2,4-D through the compromised bloodlbrain barrier (Elo et al., 1988; Hervonen et al., 1982; Tyynela et aI., 1990). Saturation of plasma protein binding also may contribute slightly to the increased brain:blood ratio of 2,4-D reported in rats at these exposure levels (Tyynela et aI., 1990; Ylitalo et aI., 1990). 2,4-D also has been reported to pass the placental barrier in mice, rats, and pigs, and has been detected in the uterus, placenta, fetus, and intrauterine fluid of exposed animals (Erne, 1966a; Fedorova and Belova, 1974; Lindquist and Ullberg, 1971) but was rapidly eliminated (Lindquist and Ullberg, 1971). 72.6.3 PHARMACOKINETICS Depending on the chemical form of 2,4-D and the animal species tested, plasma half-lives following oral exposure of
72.6 Toxicological Studies
1627
100 mg/kg body weight range from 3.5 to 12 hours (Erne, 1966a). Lower doses (i.e., 3 mg/kg body weight) in rats showed half-lives of 0.5 to 0.8 hours (Khanna and Fang, 1966), indicating that clearance rates are highly dependent on dose. In human studies, plasma clearance of orally administered 2,4-D was found to follow first-order kinetics with urinary excretion half-lives ranging from 10.2 to 28.4 hours (Sauerhoff et aI., 1977), which is consistent with the findings (urinary excretion half-life = 18 hours) from a forestry worker who exhibited the highest amount of 2,4-D excretion among two groups exposed over a period of 11 or 18 days (Frank et aI., 1985). The pharmacokinetics of 2,4-D following dermal absorption is apparently different from that following the oral route (Pelletier et aI., 1989). Plasma levels tend to reach a plateau and decline more rapidly following oral exposure. In addition, plasma clearance has been reported to follow biphasic kinetics beginning 8 hours post-dosing, with half-lives for various tissues ranging from 0.6 to 2.3 hours for the first phase and 25.7 to 29 hours for the second phase. Furthermore, cumulative urinary excretion of2,4-D increases more slowly following dermal rather than oral exposure.
of the dose was excreted within 48 hours, whereas at higher doses (i.e., ::::60 mg/kg body weight) the percentage of dose excreted within 24 hours decreased linearly with increasing dose (Khanna and Fang, 1966). In rats, ::::90% of oral doses of 30 mg/kg body weight or less were excreted in the urine within 24 hours (Khanna and Fang, 1966; Knopp and Schiller, 1992; Pelletier et aI., 1989). Similarly, in humans administered an oral dose of 5 mg 2,4-D/kg body weight, 77% of the dose was excreted within 96 hours (Kohli et aI., 1974) and 87 to 100% of the dose was excreted in the urine over 6 days (Sauerhoff et aI., 1977). 2,4-D is predominately excreted by the kidney using an active transport system. Saturation of renal clearance appears to occur at 50 to 60 mg/kg body weight (Gorzinski et aI., 1987; Khanna and Fang, 1966) based on kidney concentrations and urinary excretion rates. Another significant route of excretion in occupationally exposed workers is perspiration (Sell et aI., 1982). Following a 2-hour exposure, 2,4-D was detected in T-shirt extracts (i.e., measure of perspiration) for 2 weeks and in urine for 5 days. 2,4-D also has been reported to be excreted in the milk of lactating rats exposed to 2,4-D (Fedorova and Belova, 1974).
72.6.4 METABOLISM
72.6.6 ANIMAL STUDIES
Once absorbed into body fluids and tissues, the salts and esters of 2,4-D undergo acid and/or enzymatic hydrolyzation to form 2,4-D acid. In laboratory animals and humans following oral exposures, the presence of acid-hydrolyzable conjugates has been reported at 0 to 27% of the administered 2,4-D (Erne, 1966b; Grunow and Bohme, 1974; Kohli et al., 1974; Sauerhoff et aI., 1977). The available data indicate that 2,4-D is not metabolized to reactive intermediates and is excreted predominately as the parent compound.
72.6.6.1 Acute Toxicity
72.6.5 EXCRETION
Numerous acute toxicological tests have been conducted on the various forms of 2,4-D as summarized in Table 72.1. Overall, oral exposure to 2,4-D shows moderate to low toxicity, whereas dermal and inhalation toxicity are low. Dermal irritation in rabbits was considered slight for the acid form of 2,4-D and minimal for the salt and ester forms. Reported eye irritation in rabbits, on the other hand, is severe for the acid and salt forms, but minimal for the ester.
72.6.6.2 Subchronic Toxicity
Regardless of the route of exposure, 2,4-D is predominately excreted in the urine (Erne, 1966a; Feldman and Maibach, 1974; Khanna and Fang, 1966; Knopp and Schiller, 1992; Moody et aI., 1990, 1991; Pelletier et aI., 1989). The rate of urinary excretion is inversely proportional to dose. For example, at oral doses of 3 to 30 mg/kg body weight given to rats, 93 to 96%
Further to the acute toxicity data, numerous subchronic studies have been conducted on a variety of 2,4-D forms, by different exposure routes, and in various animal species. The subchronic studies conducted range from 3-week dermal studies in rabbits to 13-week dietary studies in dogs and rodents. The results of these studies are summarized in Table 72.2. Overall, at doses
Table 72.1 Acute Toxicitya Involving Various Chemical Forms of2,4-D Oral LD50 (mglkg bw/d) Form of 2,4-D
Rat
Mouse
Dog
Dermal LD50 (mg/kg bw/d)
Inhalation LD50 (mglliter)
Guinea pig
Chicken
Rat
Rabbit
Rat 1.79
Acid
639-980
312-434
25-250
397-553
358-817
na
1400->2000
Salt
863->2000
na
na
na
na
>2000
>2000
>3.8
Ester
440-982
na
na
na
na
na
1829->2000
>4.6
bw = body weight. na = not available. aCondensed from Munro et al. (1992).
1628
CHAPTER 72
Phenoxy Herbicides (2,4-D)
Table 72.2 Summaryaa of Subchronic Studies on 2,4-D Tested in Various Animal Species
Chemical species
Route
Dose (mg/kg bw/d)
NOAELb
[in acid equivalents]
(mg/kg bw/d)
Reference
13-week: Rat 2,4-D (100% pure)
diet
0, 15,60, 100, 15O
15 (females only)
Gorzinski et af. (198Ia)
2,4-D (97.5% pure)
diet
0, 15,60, 100, 150
15 (females only)
Gorzinski et af. (1981b)
2,4-D (97.5% pure)
diet
0,1,5,15,45
2,4-D (96.1 % pure)
diet
0, 1, 15, 100, 300
IS
Charles et al. (1 996b)
2,4-D ethylhexyl ester
diet
0, 1, 15, 100,300
15
Charles et af. (1 996b)
2,4-D dimethylamine salt
diet
0, 1, 15, 100,300
15
Charles et af. (1996b)
2,4-D butoxyethyl ester
diet
0,1,15,100,300
15
Szabo and Rachunek (1991)
2,4-D triisopropanolamine salt
diet
0, 1, 15, 100,300
IS
Yano et af. (1991b)
2,4-D isopropylamine salt
diet
0, 1, 15, 100, 300
15
Yano et af. (1991a)
intraperitoneal
0, 100,150
1 (males only)
Serota (1983a)
12-week: Rat 2,4-D sodium salt
Lukowicz-Ratajczak and Krechniak (1988)
13-week: Mouse 2,4-D acid
diet
0,5,15,45,90
2,4-D acid
diet
0, 1, 15, 100,300
oral intubation
50,100,200
Kuntz et af. (1990)
diet
100
Lundgren et al. (1987)
2,4-D acid
gelatin capsule
0,0.3, 1,3, 10
ITF (1990)
2,4-D acid
diet
0, 0.5, 1.0, 3.75, 7.5
Charles et af. (1996c)
2,4-D dimethylamine salt
diet
0, 1.0,3.75,7.5
Charles et af. (1996c)
2,4-D 2-ethylhexyl ester
diet
0, 1.0,3.75,7.5
Charles et al. (1 996c)
2,4-D acid
dermal
0, 10, 100, 1000
2,4-D dimethylamine salt
dermal
0, 10, 100, 300
2,4-D 2-ethylhexyl ester
dermal
0, 10, 100, 1000
2,4-D triisopropanolamine salt
dermal
0,55,193,553
553
Mizell et af. (1 990b)
2,4-D isopropylamine salt
dermal
0,39,98,275
275
Mizell et af. (1990a)
2,4-D butoxyethyl ester
dermal
0,32,96,321
321
Mizell et af. (1989)
Serota (1983b) 15
Schulze (1991)
14-day: Mouse 2,4-D acid
4-day: Mouse 2,4-D acid
13-week: Dog
21-day: Rabbit 1000 10 10
Schulze (1990a) Schulze (1 990b) Schulze (1990c)
from Munro et af. (1992). bNOAEL = no-observed-adverse-effect level.
a Adapted
above the threshold of saturation for renal clearance, the key target organs in rats appear to be primarily the kidney and, to some extent, the thyroid. The changes reported in the rat kidney were the loss of epithelial cells in the proximal tubule brush border; the changes in the thyroid were follicular cell hypertrophy in association with a reduction in serum thyroxine levels. These changes were consistent over all forms of 2,4-D tested, with a reported no-observed-adverse-effect level (NOAEL) of 15 mg/kg body weight/day (Charles et aI., 1996b; Szabo and Rachunek, 1991; Yano et aI., 1991a, 1991b). Some of these findings were reported at lower doses in older rat studies us-
ing the acid form of 2,4-D (Gorzinski et aI., 1981a, 1981b); however, the histological effects were not statistically significant at 15 mg/kg body weight/day, which is consistent with the more recent studies. In another older study (Serota, 1983a), other minor histological effects in the kidney were reported in male rats at 5 mg/kg body weight/day and in female rats at 1 mg/kg body weight/day. These effects were not reported in any other of the subchronic studies. Although some thyroid changes have been reported at 15 mg/kg body weight/day in rats (Serota, 1983a), these changes were considered incidental and do not affect the conclusion that the subchronic NOAEL
72.6 Toxicological Studies
for 2,4-D is 15 mg/kg body weight/day (Munro et aI., 1992). Similar results with respect to the kidney have been reported in 13-week dietary mouse studies (Serota, 1983b; Schulze, 1991). In a 13-week dog study in which 2,4-D was administered by gelatin capsules, kidney effects consisting of reduced cytoplasmic eosinophilia of the epithelial cells lining some convoluted tubules were reported at lower doses, resulting in a NOAEL of 1 mg/kg body weight/day (lTF, 1990). Thyroid changes were not reported in the mouse or dog. 72.6.6.3 Reproductive and Developmental Toxicity Several multigenerational and developmental animal studies have been conducted to assess the potential of 2,4-D to affect reproduction and the developing fetus, and are summarized in Table 72.3. In general, the results of the available studies indicate that 2,4-D is not teratogenic and does not affect reproduction except at maternally toxic doses or those saturating the threshold for renal clearance (i.e., ~50 mg/kg body weight/day). At doses above the maximum tolerated dose (MTD), some developmental effects have been reported in test animals (i.e., decreased fetal weight gain, increased incidence of lumbar ribs and wavy ribs, and delayed ossification of bone). The only teratogenic effect (i.e., cleft palate) reported was in mice, but occurred only at maternally toxic doses. The potential testicular and ovarian toxicity of 2,4-D has been extensively evaluated in a recent series of subchronic and chronic studies in rats. In rats fed 0, 1, 15, 100, or 300 mg/kg body weight/day of either 2,4-D acid, 2,4-D dimethylamine salt, or 2,4-D 2-ethylhexyl ester for 90 days, only minimal effects were noted in testes at the top dose of 300 mg/kg body weight/day (Charles et aI., 1996b). These effects, consisting of decreased testeslbody weight ratios accompanied by slight histological evidence of atrophy, occurred only at a dose which exceeded the MTD. The NOAEL for testicular effects was 100 mg/kg body weight/day, while the overall NOAEL for the subchronic studies was 15 mg/kg body weight/day based primarily on minor effects in the kidney. No 2,4-D-induced toxicity was reported in the ovaries at any dose. In a subsequent chronic toxicity/oncogenicity study conducted in rats with 2,4-D acid at doses of 0,5, 75, or 150 mg/kg body weight/day, no treatmentrelated effects were reported in testes or ovaries at any dose level (Charles et aI., 1996a). The findings from these studies were consistent with observations from a series of earlier 90-day (Gorzinski et aI., 1987; Serota, 1983a, b; Serrone et aI., 1991; Szabo and Rachunek, 1991; Yano et aI., 1991a, 1991b) and chronic studies (Charles et aI., 1996a; Serota, 1986) conducted with 2,4-D acid and a variety of salt and ester derivatives. The testicular and ovarian toxicity of 2,4-D acid and its dimethylamine salt and 2-ethylhexyl ester also have been examined in a recent series of subchronic and chronic studies in beagle dogs. Dogs administered either 0, 1,3.75, or 7.5 mg/kg body weight/day of 2,4-D acid, 2,4-D dimethylamine salt, or 2,4-D 2-ethylhexyl ester for 13 weeks exhibited decreased testes weights at the two highest dose levels (Charles et aI.,
1629
1996c). The toxicological significance of these findings is uncertain since the organ weight changes were not accompanied by any corroborative histological changes. In addition, both of the two high-dose group animals exhibited body weight gain depressions of approximately 30 to 85%. No 2,4-D-induced effects were seen in ovaries. A NOAEL of 1.0 mg/kg body weight/day was established from these studies based on effects in the kidneys. In a follow-up I-year chronic study conducted at dietary doses of 0, 1,5, or 7.5 mg/kg body weight/day of 2,4-D acid, no testes or ovary alterations were reported (Charles et aI., 1996c). The findings from these dog studies were consistent with earlier 13-week (lTF, 1990) and 2-year dog studies (Hansen et aI., 1971). The generallack of 2,4-D-associated testicular toxicity is entirely consistent with the failure of 2,4-D to induce changes in reproductive performance. 72.6.6.4 Immunotoxicity Several subchronic and chronic toxicity studies have provided no evidence from hematological, clinical chemistry, or histopathologic evaluations that 2,4-D is likely to induce immune system dysfunction (Charles et aI., 1996a, 1996b, 1996c; Szabo and Rachunek, 1991; Yano et aI., 1991a, 1991b). Studies that were conducted specifically to examine the possible impact of 2,4-D on various immune system functional parameters have not provided definitive results (Blakley, 1986; Blakley and Blakley, 1986; Blakley and Schiefer, 1986; Zhamsaranova et aI., 1987). These studies are difficult to interpret in that the results (1) were inconsistent when evaluated by different routes of exposure, or by comparison of findings from acute and subchronic test regimes; (2) often not reproducible; and (3) not accompanied by adequate descriptions of the normal range of immune parameter values in the test animal populations (Munro et aI., 1992). Dennal exposure to 2,4-D acid, salts, or esters also has not been associated with delayed contact hypersensitivity in guinea pigs (Carreon et aI., 1983; Carreon and Rao, 1985; Gargus, 1986; Jeffrey, 1986; Jeffrey and Rao, 1986; Schultz et aI., 1990). 72.6.6.5 Neurotoxicity Overall, it may be concluded that 2,4-D has little, if any, potential to induce adverse effects in the nervous system at doses that do not cause overt systemic toxicity or that do not saturate processes involved with tissue clearance and renal excretion. No lesions or overt clinical signs of central nervous system toxicity were observed in any of the subchronic toxicity studies in rats (Charles et aI., 1996b; Szabo and Rachunek, 1991; Yano et aI., 1991a, 1991b) or mice (Schulze, 1991), at doses up to 300 to 560 mg/kg body weight/day. In a chronic rat study designed specifically to investigate the impact of 2,4-D on the nervous system, several neurological parameters were assessed, of which only forelimb grip strength was altered, to a minimal degree, at the highest dose tested, 150 mg/kg body weight/day (Jeffries et aI., 1994). Other animal studies have yielded electromyogram results considered indicative of skeletal muscle myotonia following administration of
1630
CHAPTER 72
Phenoxy Herbicides (2,4-D)
Table 72.3 Summary of Developmental and Reproductive Toxicity Studies on 2,4-D Tested in Various Animal Species NOAELa
Exposure Route
duration
Dose (mg/kg bw/d)
2,4-D triisopropanolamine salt
gavage
GDb 7-19
0, 10, 30, or 7SC
IO d ; 75 e
Liberacki et al. (1994),
2,4-D isopropylamine salt
gavage
GD7-19
0, 10, 30, or 7SC
IO d ; 75'
Liberacki et al. (1994),
2,4-D butoxyethyl ester
gavage
GD 7-19
0, 10, 30, or 75 C
IQd; 75'
Liberacki et al. (1994),
2,4-D acid
gavage
GD 6-18
0, 10, 30, or 90
3Qd; 90e
Hoberman (1990),
2,4-D dimethylamine salt
gavage
GD 6-18
0, 10,30, or 90c
30d ; 90e
Martin (1991),
2,4-D ethylhexyl ester
gavage
GD 6-18
0, 10,30, or 7SC
30d ; 75 e
Martin (1992a),
2,4-D diethanolamine
gavage
GD 6-19
0, 10.2, 20.3, or 40,6C
1O.2d ; 40.6e
oral
GD 6-15
0, 12.5, 25, 50, 75, or 87.5 c 87.5 c
Chemical species
(mg/kg bw/d)
Reference
Developmental: Rabbit
Charles et al. (l996a) Charles et al. (l996a) Charles et at. (1996a) Charles et al. (l996a) Charles et al. (l996a) Charles et al. (l996a)
Developmental: Rat
2,4-D isooctyl ester 2,4-D propylene glycol butyl ether
oral
GD 6-15
0, 12.5, 25, 50, 75, or
2,4-D acid
oral
GD6-15
0, 12.5,25,50,75, or 87.5
2,4-D isooctyl ester
oral
GD5-8
0, or 87.SC
2,4-D propylene glycol butyl ether
oral
GD5-8
0, or 87.5'
2,4-D isoocty1 ester
oral
GD8-11
0,50, or 87.5' 87.5 c
87.5 d ; 25 e 87.5 d ; 25 e
Schwetz et al. (1971)
87.5 d ; 25' 87.5 d .e
Schwetz et al. (1971)
87.5 d.e 87.5 d ; <50e
Schwetz et al. (1971)
87.5 d ,e
Schwetz et al. (1971) Khera and McKinley (1972)
Schwetz et al. (1971) Schwetz et al. (1971) Schwetz et al. (1971)
2,4-D isooctyl ester
oral
GD 12-15
0,50, or
2,4-D isoocty1 ester
oral
GD 6-15
0,50, or 150
2,4-D butyl ester
oral
GD 6-15
0,50, or 150
15Qd; 50e 15Qd; 50e
0,50, or 150
150d ; 50e
Khera and McKinley (1972)
0, lOO, or 300
300d ; 50e
Khera and McKinley (1972)
0,50, or 100
100d ; 50e
Khera and McKinley (1972)
0, 25, 50, or 100
IOQd; 50e 15Qd; 50e
Khera and McKinley (1972)
87.5 d ; 25' 87.5 d ; 25'
Unger et al. (1981) Chemoff et al. (1990)
2,4-D butoxyethynol 2,4-D dimethylamine salt (49.5%) 2,4-D acid
oral oral oral
GD6-15 GD6-15 GD 6-15
2,4-D acid
oral
GD 6-15
2,4-D acid
oral
GD 6-15
0,25,50, 100, or 150 0,6.25, 12.5,25, or 87.SC
Khera and McKinley (1972)
Khera and McKinley (1972)
2,4-D propylene glycol butyl ether
oral
GD 6-15
2,4-D isooctyl ester
oral
GD 6-15
0,6.25, 12.5,25, or 87.5'
2,4-D acid
gavage
GD 6-15
0, or 115
2,4-D ethylhexyl ester
gavage
GD6-15
0, 10, 30, or 90C
<115 d,e IO d ,30e
2,4-D dimethylamine salt
gavage
GD 6-15
0, 12.5,50, or 100c
12.5 d ; 50e
Lochry (1990),
2,4-D acid
gavage
GD 6-15
0, 8, 25, or 75
75 d,e
Nemec et al. (1983),
Unger et al. (1981) Martin (1992b), Charles et at. (1996a) Charles et al. (1996a) Charles et al. (1996a)
Developmental: Mouse
2,4-D acid
oral
GD7-15
0, or 0.56 mMlkg bw
<0.56d ; 0.56e
Courtney (1977)
2,4-D acid
oral
GD 11-14
0, or 0.80 mMlkg bw
<0.8Qd,e
Courtney (1977)
2,4-D acid
oral
GD 12-15
0, or 1 mMlkg bw
Courtney (1977)
2,4-D acid
subcutaneous
GD 12-15
0, or I mMlkg bw
Courtney (1977)
2,4-D isopropyl ester
oral
GD 7-15
0, or 0.56 mM/kg bw
Courtney (1977)
2,4-D n-butyl ester
oral
GD 7-15
0, or 0.56 mMlkg bw
<0.56d ,e 0.56d ,e
2,4-D n-butyl ester
oral
GD 12-15
0, or 1 mMlkg bw
Courtney (1977)
2,4-D isooctyl ester
oral
GD 7-15
0, or 0.56 mMlkg bw
Courtney (1977)
2,4-D propylene glycol butyl ether
oral
GD7-15
0, or 0.56 mMlkg bw
<0.56d,e <0.56d,e
Courtney (1977)
Courtney (1977) (continues)
72.6 Toxicological Studies
1631
Table 72.3 (continued) NOAELa
Exposure
(mglkg bw/d)
Reference
0, or I IlL'Wkg bw
Id;
Courtney (1977)
0, or 87.5
87.5 d ,e
Kavlock et al. (1987)
0, or 87.5
87.5 d ,e
Kavlock etal. (1987)
OD 8-12
0, or 87.5
87.5 d ,e
Kavlock et al. (1987)
subcutaneous
OD 6-14
0, or 100
<100d ,e
Bionetics (1968)
2,4-0 acid
subcutaneous
006-14
0, or 98
<98d ,e
Bionetics (1968)
2,4-D acid
subcutaneous
OD 6-14
0, or 215
2ls d ,e
Bionetics (1968)
2,4-D acid
subcutaneous
OD 6-14
0, or 50
sod,e
Bionetics (1968)
2,4-D acid
oral
OD 6-14
0, or 10O
<100d ,e
Bionetics (1968)
2,4-D
oral
OD 6-10
0,20,40,60,orIOO
Collins and Williams (1971)
2,4-D
oral
OD 6-10
0,40,60,orl00
100e 100e
2,4-0
oral
006-10
0,40,60, or 100
40
Collins and Williams (1971)
diet
2-generation
0, 5, 20, or 80
2od;5'
Rodwell (1984)
Chemical species
Route
duration
2,4-D propylene glycol butyl ether
oral
OD 12-15
2,4-D acid
oral
OD 8-12
2,4-D propylene glycol butyl ether
oral
OD 8-12
2,4-D isooctyl ester
oral
2,4-D acid
Dose (mg/kg bw/d)
Developmental: Hamster Collins and Williams (1971)
Reproductive Toxicity: Rat 2,4-D acid
aNOAEL = no-observed-adverse-effect level. = gestational days. c2,4-D acid equivalents. dMatemal. hOD
epetal.
high doses of2,4-D (50 to 100 mg/kg body weight/day) (Arnold et aL, 1991; Beasley et aL, 1991; Elo and MacDonald, 1989; Kwiecinski, 1981; Steiss et aL, 1987; Toyoshima et aL, 1985), In a subchronic toxicity study on 2,4-D 2-ethylhexyl ester, some of the high-dose animals were reported to show clinical signs that could possibly be related to myotonia (e.g., hunched posture, languid behavior, ataxia) (Charles et al., 1996b). Myotonia induced by high levels of exposure to 2,4-D does not appear to be the result of toxicological action upon the central nervous system (Buslovich and Pichugin, 1983), but appears to be due to effects mediated at the junction of skeletal muscle nerves and muscle tissue. The biochemical mechanism involved in the induction of myotonia in experimental animals is not well understood; however, according to Rudel and Senges (1972), alteration of chloride ion conductance in muscle fibers appears to be involved. Because the development of myotonia in animals exposed to high doses of 2,4-D was not accompanied by any pathological effects, and because the reporting of myotonia is restricted to dose levels greater than the threshold for saturation of renal tubular secretion, the effects reported in animal studies are not considered to be indicative of a potential of 2,4-D to induce peripheral polyneuropathy in humans. In an independent review of the literature, it was concluded that exposure to 2,4-D does not produce polyneuropathy in humans nor does polyneuropathy occur in several animal species exposed to high levels of 2,4-D (Mattsson and Eisenbrandt, 1990). Using tests ofneurobehavioral parameters (such as the Functional Observational Battery), decreased activity levels, behav-
ioral changes, and motor skill abnormalities have been reported in rats at doses greater than 60 mg/kg body weight/day and in rabbits at doses of approximately 30 mg/kg body weight/day (Breslin et aL, 1991; de Duffard et aL, 1990b; Duffard et aL, 1995; Hoberman, 1990; Jeffries et aL, 1994; Liberacki et aL, 1991; Martin, 1991; Mattsson et aL, 1994, 1997; Oliveira and Palermo-Neto, 1993; Rodwell, 1991; Schulze and Dougherty, 1988; Zablotny et aL, 1991). In acute studies conducted using beagle dogs, clinical signs of central nervous system depression and/or abnormalities in electroencephalograms were only reported at doses of 175 mg/kg or greater (Arnold et al., 1991). Rats exposed to high doses of 2,4-D n-butyl ester were reported to display alterations in neurotransmitter concentrations in the brain (de Duffard et aL, 1990a; Elo and MacDonald, 1989; Oliveira and Palermo-Neto, 1993). These neurochemical alterations have been hypothesized to result from compromise of the blood-brain barrier by high doses of 2,4-D (Elo et aL, 1988; Tyynela et aI., 1986). These authors reported that doses of 2,4-D greater than 150 mg/kg body weight resulted in extravasation of albumin in various areas of the brain. Several investigators have reported accumulation of 2,4-D in the brain or cerebrospinal fluid, following administration of high doses (40 to 300 mg/kg body weight) of 2,4-D (Elo and Ylitalo, 1977, 1979; Kim et aL, 1988; Oliveira and PalermoNeto, 1993; Tyynela et aL, 1990). Kim et aL (1988) suggested that the increased accumulation of 2,4-D in the brain at high doses was likely not the result of increased permeability of the blood-brain barrier since the entry of the organic solute,
1632
CHAPTER 72
Phenoxy Herbicides (2,4-D)
2-deoxyglucose, into rabbit brain was unaffected by 2,4-D pretreatment. Instead, it has been hypothesized that reduced elimination of 2,4-D from the brain via the choroid plexus through competitive inhibition of the organic acid transport pathway was likely responsible for the increased accumulation of 2,4-D (Kim et aI., 1988; Ylitalo et aI., 1990). The organic acid transport pathway normally actively eliminates acidic metabolites from the brain through the blood-brain barrier. At doses below the capacity of normal renal clearance, there is no evidence in experimental animals to indicate that 2,4-D can have an impact on the nervous system. In fact, no clinically observable adverse effects on the nervous system have been observed in animals at doses below 10 to 30 mglkg body weight, even in long-term studies. 72.6.6.6 Chronic Toxicity and Carcinogenicity Several long-term bioassays have been conducted in rats, mice, and dogs (Arkhipov and Kozlova, 1974; Charles et aI., 1996a, 1996c; Innes et aI., 1969; Hansen et aI., 1971; Serota et aI., 1986, 1987). There has been no evidence to suggest that 2,4-D acts as a carcinogen in any of these species. Rats In one older 2-year rat feeding study (Serota et al., 1986), an increase in the incidence of brain astrocytomas was reported in male rats only at the highest dose tested of 45 mglkg body weight/day; however, the biological characteristics of the tumors were not consistent with chemical carcinogenesis. Moreover, based on the lack of decreased latency, the lack of increased multiplicity, the lack of increased severity, the lack of preneoplastic or target organ effects, the restriction of tumor development to one species and sex, the intergroup variability exhibited among historical controls, the lack of a plausible mechanism of tumorigenesis, the low exposure of the brain to 2,4-D compared to other tissues, and the fact that these tumors have not been reproduced in subsequent studies, it is unlikely that the increased incidence of brain astrocytomas reported by Serota et al. (1986) was related to 2,4-D treatment. In another older study (Hansen et aI., 1971) in which rats were fed up to 62.5 mg 2,4-Dlkg body weight/day for 2 years, an overall increase in the number of randomly distributed tumors was reported to be statistically significant for male rats. As discussed by Munro et al. (1992), this study was not considered to provide any evidence that 2,4-D is carcinogenic in the rat since it did not meet the requirements of Good Laboratory Practice (GLP) standards, the dose groups were fairly small, the maximum tolerated dose (MTD) was not achieved, and the microscopic examination was not comprehensive. In a third feeding study (Arkhipov and Kozlova, 1974), rats were fed 10% of the reported LDso (details of dosing not reported) with no significant increase in tumor incidence. More recently, a 2-year GLP-compliant study in which rats were fed 5 to 150 mg 2,4-Dlkg body weight/day was completed without any evidence of carcinogenicity (Charles et aI., 1996a). In particular, there was no increased incidence of brain astrocytomas even at the MTD. Noncancer endpoints reported in the
animals were very similar to those reported in the subchronic studies (Charles et aI., 1996b), and a NOAEL of 5 mglkg body weight/day was established based on increased thyroid weight. Mice ~o evidence of carcinogenicity has been reported in three long-term mouse studies (Charles et aI., 1996a; Innes et aI., 1969; Serota et aI., 1987). In the first study (lnnes et aI., 1969), mice were orally administered one of three esters of 2,4-D (i.e., isopropyl, butyl, or isooctyl ester) at a dose of 46.4 mglkg body weight/day for 18 months. No increase in tumor incidence was reported. In the second study (Serota et aI., 1987), mice were fed 1 to 45 mg 2,4-Dlkg body weight/day for 106 weeks with no evidence of carcinogenicity or other treatmentrelated effects. In male mice administered the highest two doses (15 and 45 mglkg body weight/day), an increase in cytoplasmic homogeneity in the renal tubular epithelium was reported. In the third and most recent GLP-compliant study (Charles et aI., 1996a), female and male mice were fed 5 to 300 and 5 to 125 mg 2,4-Dlkg body weight/day, respectively, for 2 years without any evidence of tumorigenesis. Noncancer effects were limited to slight depression of red blood cell parameters, minor organ weight changes, and histopathological renal effects at the top two doses in both sexes. Dogs Similar to the results from the rodent studies, there has been no evidence to suggest that 2,4-D has carcinogenic potential in dogs (Charles et aI., 1996c; Hansen et aI., 1971). The results of the long-term studies in dogs support the results reported in subchronic studies. In the older study by Hansen et al. (1971), small groups of beagle dogs were fed 2,4-D in the diet at concentrations reaching 500 ppm over a 2-year period. Following gross and microscopic examinations of several tissues and organs, no lesions related to 2,4-D treatment were reported. The more recent study by Charles et al. (1996c) examined the effects of feeding 0, 1, 5, or 7.5 mg 2,4-Dlkg body weight/day to beagle dogs for a period of 52 weeks. The reported effects included body weight gain reduction in females, notably at the highest dose, some serum chemistry alterations (i.e., increased urea nitrogen, creatinine, cholesterol, and alanine aminotransferase activity) in the two highest dose groups, and some histopathological alterations (i.e., perivascular chronic active inflammation of the liver and an increase of pigment in tubular epithelium in both sexes in the two highest dose groups, and pigment in the sinusoidal lining cells in the females of the two highest dose groups). Overall, 2,4-D administration was well tolerated and produced no effects on clinical signs, hematology, urinalysis, or gross necropsy. A NOAEL of I mglkg body weight/day was suggested by the authors. 72.6.7 GENOTOXICITY Numerous in vitro and in vivo genotoxicity studies have been conducted with 2,4-D. Overall, the results indicate that 2,4-D has very little genotoxic potential. This conclusion has been
72.7 Studies in Humans reached in previous reviews of 2,4-D (CCT, 1987; EPA, 1997; Munro et al., 1992) and is consistent with metabolism studies which have indicated that 2,4-D does not metabolize to reactive intermediates. With very few exceptions, bacterial mutagenicity tests using Salmonella typhimurium and Escherichia coli have produced negative results with 2,4-D (Anderson and Styles, 1978; Charles et aI., 1996a; Ercegovich and Rashid, 1977; Kappas, 1988; Kappas and Markaki, 1988; Kappas et aI., 1984; MerschSundermann et aI., 1989; Rashid, 1979; Rashid and Mumma, 1986; Rashid et aI., 1984; Simmon et aI., 1977; Soler-Niedziela et al., 1988; Styles, 1973; Waters et al., 1980). In yeast cells, mitotic gene conversion and recombination has been reported, but was highly dependent on pH and occurred only at pH 4.3 (Simmon et aI., 1977; Waters et aI., 1980; Zetterberg, 1978; Zetterberg et al., 1977). Negative or weakly positive results were reported in unscheduled DNA repair and sister chromatid exchange (SCE) assays with mammalian cell systems (Charles et aI., 1996a; Clausen et aI., 1990; Galloway et aI., 1987; Jacobi and Witte, 1991; Styles, 1977; Waters et aI., 1980). For the most part, the weakly positive results occurred only in conjunction with cytotoxicity (Clausen et aI., 1990; Korte and Jalal, 1982). Similar to the in vitro studies, the majority of in vivo studies with animals, using the most accepted and validated procedures, have produced negative results (Munro et aI., 1992). Some studies with occupationally exposed individuals provided marginally positive results in lymphocytes but could not be directly correlated with 2,4-D exposure due to various confounding factors (e.g., age, sex, race, lifestyle habits, etc.) (Crossen et al., 1978; Kaye et aI., 1985; Yoder et aI., 1973). Several other studies have shown that 2,4-D exposure has no effect on chromosomal aberration or SCE frequency (Charles et aI., 1996a, b; Hogstedt and Westerlund, 1980; Linnainmaa, 1983a, 1983b, 1984; Mulcahy, 1980; Mustonen et aI., 1986, 1989).
72.7 STUDIES IN HUMANS There has been some concern over a possible aSSOCIation between exposure to 2,4-D and the development of cancer, specifically non-Hodgkin's lymphoma, Hodgkin's disease, and soft tissue sarcoma. In a critical review of over 90 epidemiology studies relating to occupational exposures to various herbicides including 2,4-D, Munro et al. (1992) concluded that the epidemiological findings are inconsistent, the evidence for a causal association between 2,4-D and cancer is weak, and that no association between 2,4-D and human cancer has been convincingly documented. In overview, the majority of the studies were conducted with farmers, forestry workers, and other similar groups of potential users of herbicides. In most of the studies, there were methodological shortcomings in conducting exposure assessments specifically related to 2,4-D. Moreover, the majority of the studies involved occupational exposures to a wide variety
1633
of chemical, physical, and biological agents including phenoxy herbicides, and it was difficult to discern specific exposure to 2,4-D. Without specific information regarding exposure to 2,4D and with the contribution of other confounding factors, the establishment of a dose-response relationship is difficult to ascertain. In cohort studies, exposure to 2,4-D could be reasonably assumed; however, no conclusive evidence was reported to show an association between 2,4-D and cancer. The only positive correlations were reported in three large cohort studies (Saracci et aI., 1991; Wigle et aI., 1990; Wiklund and Holm, 1986; Wiklund et aI., 1987), in which exposures were primarily to 2,4,5-T or were mixed with other herbicides, and none of the reported effects were consistent among the studies. Although there has been a persistent hypothesis that exposure to 2,4-D may be associated with an increased incidence of nonHodgkin's lymphoma (Hoar et al., 1986; Zahm et aI., 1990), the analysis by Munro et al. (1992) concluded that the results of these case-control studies did not strongly support the hypothesis based on weaknesses in the methodology employed and lack of control for other possible risk factors for non-Hodgkin's lymphoma (e.g., viruses and immune system modulation). Other reviews have been conducted since the work by Munro et al. (1992). A SAB/SAP Special Joint Committee from the D.S. Environmental Protection Agency (EPA, 1994) reviewed the available data on 2,4-D and concluded that " ... the data are not sufficient to conclude that there is a cause and effect relationship between the exposure to 2,4-D and non-Hodgkin's lymphoma" and 2,4-D still remains classified as a Group D (not classifiable as to human carcinogenicity) (EPA, 1997). Similarly, the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and the WHO Expert Group on Pesticide Residues reviewed the data on 2,4-D and stated that the results of the available epidemiology studies are inconsistent and that any reported associations are weak (Rowland, 1997). The National Cancer Institute of Canada also convened an Ad Hoc Panel on Pesticides and Cancer which concluded that "it was not aware of any definitive evidence to suggest that synthetic pesticides contribute significantly to overall cancer mortality" (Ritter, 1997). A few additional studies have been published since the above reviews (Becher et aI., 1996; Bums et aI., 2001; Fleming et aI., 1997; Zahm, 1997). Three of these are mortality studies in factory workers (Becher et al., 1996; Bums et aI., 2001) and lawn care workers (Zahm, 1997), and one is a retrospective cohort study in pesticide applicators (Fleming et aI., 1997). As with previous epidemiology studies, the findings were inconsistent and did not provide any conclusive evidence of increased cancer risk associated with exposure to 2,4-D. In fact, Fleming et al. (1997) reported that in a cohort of 33, 669 pesticide applicators, there were no confirmed cases of soft tissue sarcoma or nonHodgkins lymphoma and Bums et al. (2001) concluded there was "no evidence of causal association between exposure to 2,4-D and mortality due to all causes and malignant neoplasm" and "no significant risk due to NHL was found." In addition, Acquavella et al. (1998) conducted a meta-analysis of 37 stud-
1634
CHAPTER 72
Phenoxy Herbicides (2,4-D)
ies in farmers and concluded that farmers did not have elevated rates of several cancers, with the exception of lip cancer. In a preliminary study with a small group of farmers, Faustini et al. (1996) reported that 2,4-D exposure affected some immunological variables; however, the data were highly variable (i.e., large standard deviations) and the group tested was very small (n = 10). There has been some discussion in the literature linking immunosuppressive agents with an increased incidence of non-Hodgkin's lymphoma (Hardell et al., 1998; Hardell and Axelson, 1998), but there has been no clear or consistent evidence in humans indicating that 2,4-D affects the immune system. This is well supported by animal studies.
72.8 SUMMARY 2,4-D is the most common of the phenoxy herbicides and is one of the best-studied agricultural chemicals. It is primarily used as a herbicide in agriculture, forestry, and lawn care practices, and is effective against a wide variety of broadleaf plants. Occupational exposure to 2,4-D is mainly through dermal contact but can also occur, to a lesser extent, via ingestion and inhalation. Studies in humans have shown exposures to be extremely low even in all occupational groups (i.e., <40 J.Lg/kg body weight/day). The extensive database of metabolic, toxicological, and epidemiological studies on 2,4-D has provided no evidence that 2,4-D poses any health risk to humans when used according to label directions.
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Rabbits," Unpublished Report K-008876-016. Dow Chemical Company, Midland, MI. Liberacki, A. B., Zablotny, C. I., Yano, B. L., and Breslin, W. J. (1994). Developmental toxicity studies on a series of 2,4-D salts and esters in rabbits. Toxicologist 14(1), 162. Lindquist, N. G., and Ullberg, S. (1971). Distribution of the herbicides 2,4,5T and 2,4,-D in pregnant mice. Accumulation in the yolk sac epithelium. Experientia 27(12), 1439-1441. Linnainmaa, K (1983a). Sister chromatid exchanges among workers occupationally exposed to phenoxy acid herbicides 2,4-D and MCPA. Teratog. Carcinog. Mutagen. 3(3), 269-279. Linnainmaa, K (1983b). Nonmutagenicity of phenoxy acid herbicides 2,4dichlorophenoxyacetic acid and 4-methyl-2-chlorophenoxyacetic acid. In "Chlorinated Dioxins and Dibenzofurans in the Total Environment" (Choudhary et aI., eds.), p. 385. Butterworth Publishers, Woburn, MA Linnainmaa, K (1984). Lochry, E. A. (1990). "Developmental Toxicity (Embryo-Fetal Toxicity and Teratogenic Potential) Study of 2,4-D Dimethylamine Salt (2,4-D-DMA) Administered Orally via Gavage to Crl:CD®BR VAFlPlus® Presumed Pregnant Rats." Protocol Number: 320-001. Argus Research Laboratories, Inc., Horsham, PA. Lukowicz-Ratajczak, J., and Krechniak, J. (1988). Effects of sodium 2,4dichlorophenoxy acetate on renal function in the rat. Bull. Environ. Contam. Toxicol. 41, 815-821. Lundgren, B., Meijer, J., and Depierre, J. W. (1987). Induction of cytosolic and microsomal epoxide hydrolases and proliferation of peroxisomes and mitochondria in mouse liver after dietary exposure to p-chlorophenoxyacetic acid, 2,4-dichlorophenoxyacetic acid and 2,4,5-trichlorophenoxyacetic acid. Biochem. Pharmacol. 36(6), 815-822. Martin, T. (1991). "Developmental Toxicity (Embryo-Fetal Toxicity and Teratogenic Potential) Study of 2,4-D Dimethylamine Salt (2,4-D-DMA) Administered Orally via Stomach Tube to New Zealand White Rabbits." Protocol Number: 320-004. Argus Research Laboratories, Inc., Horsham, PA Martin, T. (1992a). "Developmental Toxicity (Embryo-Fetal Toxicity and Teratogenic Potential) Study of 2,4-D 2-Ethylhexyl Ester (2,4-D Isooctyl Ester) Administered Orally (Stomach Tube) to New Zealand White Rabbits." Protocol Number: 320-006. Argus Research Laboratories, Inc., Horsham, PA. Martin, T. (1992b). "Developmental Toxicity (Embryo-Fetal Toxicity and Teratogenic Potential) Study of 2,4-D 2-Ethylhexyl Ester (2,4-D Isooctyl Ester) Administered Orally via Gavage to Crl:CD®BR VAFlPlus® Presumed Pregnant Rats." Protocol Number: 320-005. Argus Research Laboratories, Inc., Horsham, PA Mattsson, J. L., Charles, J. M., Yano, B. L., Cunny, H. c., Wilson, R. D., and Bus, J. S. (1997). Single-dose and chronic dietary neurotoxicity screening studies on 2,4-dichlorophenoxyacetic acid in rats. Fund. Appl. Toxicol. 40, 111-119. Mattsson, J. L., and Eisenbrandt, D. L. (1990). The improbable association between the herbicide 2,4-D and polyneuropathy. Biomed. Environ. Sci. 3, 43-51. Mattsson, J. L., McGuirk, R. J., and Yano, B. L. (1994). 2,4-D Acute Neurotoxicity Study in Fischer 344 Rats." Unpublished Report K-002372-006. Dow Chemical Company, Midland, MI. Mersch-Sundermann, v., Hofmeister, A., Muller, G., and Hof, H. (1989). Untersuchungen zur mutagenitat organischer mikrokontaminationen in der umwelt. Ill. Mitteilung: Die mutagenitat ausgewahlter herbizide und insektizide im SOS-chromotest = Examination of mutagenicity of organic microcontaminations of the environment. Ill. Communication: The mutagenicity of selected herbicides and insecticides with the SOS-chromotest. Zentralblatt fur Hygiene und Umweltmedizin (Int. 1. Hyg. Environ. Med.) 189, 135-146. Mizell, M. J., Atkin, K T., and Crissman, J. W. (1990a). "2,4-Dichlorophenoxyacetic Acid Butoxyethyl Ester: 21-Day Dermal Toxicity Study in New Zealand White Rabbits." K-007722-008. The Dow Chemical Company, Midland, MI.
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Mizell, M. J., Atkin, KT., Haut, K. T., and Stebbins, K E. (1990b). "2,4-D Isopropylamine Salt: 21-Day Dennal Toxicity Study in New Zealand White Rabbits." M-004725-004. The Dow Chemical Company, Midland, MI. Mizell, M. J., Atkin, K T., and Stebbins, K E. (1989). "2,4-D Triisopropanolamine Salt: 21-Day Dennal Toxicity Study in New Zealand White Rabbits." K-008866-004. The Dow Chemical Company, Midland, MI. Moody, RP., Franklin, C. A., Ritter, L., and Maibach, H. I. (1990). Dennal absorption of the phenoxy herbicides 2,4-D, 2,4-D amine, 2,4-D isooctyl, and 2,4,5-T in rabbits, rats, rhesus monkeys, and humans: A cross-species comparison. J. Toxicol. Environ. Health 29(3), 237-246. Moody et al. (1991). MuJcahy, M. T. (1980). Chromosome aberrations and "Agent Orange." Med. J. Austral. 2(10), 573-574. Mullison, W. R (1987). "Environmental Fate of Phenoxy Herbicides. Fate of Pesticides in the Environment." Publication 3320, pp. 121-131. California Agricultural Experiment Station. Munro, I. C., Carlo, G. L., Orr, J. C, Sund, KG., Wilson, R M., Kennepohl, E., Lynch, B. S., Jablinske, M., and Lee, N. L. (1992). A comprehensive, integrated review and evaluation of the scientific evidence relating to the safety of the herbicide 2,4-D. J. Am. Coli. Toxicol. 11(5),559-664. Mustonen, R, Kangras, J., Vuojolahti, P., and Linnainmaa, K (1986). Effect of phenoxyacetic acids on the induction of chromosome aberrations in vitro and in vivo. Mutagenesis 1,241-255. Mustonen, R (1989). Nemec, M. D., Tasker, E. J., Werchowski, K M., and Mercieca, M. D. (1983). "A Teratology Study in Fischer 344 Rats with 2,4-Dichlorophenoxyacetic Acid." Industry Task Force on 2,4-D Research Data. No. WIL-81135. Wil Research Laboratories, Inc., Ashland, OH. OJiveira, G., and Palenno-Neto, J. (1993). Effects of2,4-dichlorophenoxyacetic acid (2,4-D) on open-field behaviour and neurochemical parameters of rats. Pharmacol. Toxico!. 73(2), 79-85. Orberg, J. (1980). Effects oflow protein consumption on the renal clearance of 2,4-dichlorophenoxyacetic acid (2,4-D) in goats. Acta. Pharmacol. Toxicol. 46, 138-140. Pelletier, 0., Ritter, L., Caron, J., and Somers, D. (1989). Disposition of 2,4dichlorophenoxyacetic acid dimethylamine salt by Fischer 344 rats dosed orally and dennally. J. Toxicol. Environ. Health 28, 221-234. Pelletier, 0., et al. (1990). Pritchard, J. B. (1980). Accumulation of anionic pesticides by rabbit choroid plexus in vitro. J. Pharmacol. Exper. Therapeut. 212(2), 354-359. Rashid, K. A. (1979). The relationship between mutagenic and DNA-damaging activity of pesticides and their potential for carcinogenesis. Diss. Abstr. Int. 39, 4726-B (Abstr. No. 7909115). Rashid, K A., Babish, J. G., and Mumma, R O. (1984). Potential of 2,4-dichlorophenoxyacetic acid conjugates as promutagens in the salmonella/microsome mutagenicity test. J. Environ. Sci. Health B19(849), 689-701. Rashid, K A., and Mumma, R O. (1986). Screening pesticides for their ability to damage bacterial DNA. J. Environ. Sci. Health (Part B-Pestic. Food Contam. Agric. Wastes) B21(4), 319-334. Ritter, L. (1997). Report of a panel on the relationship between public exposure to pesticides and cancer. Cancer 80(10),2019-2033. RodweII, D. (1991). "Teratology Study in Rabbits with Diethanolamine Salt of 2,4-D Acid." Unpublished Report SLS 3229.13. Springborn Laboratories, Inc., OH. Rodwell, D. E. (1984). "A Dietary Two-Generation Reproduction Study in Fischer 344 Rats with 2,4-Dichlorophenoxyacetic Acid." Project No. WIL81137. Wil Research Laboratories, Inc., Ashland, OH. Rowland, J. C. (1997). 2,4-dichlorophenoxyacetic acid (2,4-D). In "FAO Panel of Experts on Pesticide Residues, and WHO Expert Group on Pesticides Residues in the Food, and Environment. Pesticide Residues in Food-1996: Toxicological Evaluations." WHO/PCS/97 1, 45-96. Food and Agriculture Organization of the United Nations (FAO), Rome, Italy. Rudel, R, and Senges, J. (1972). Experimental myotonia in mammalian skeletal muscle: Changes in membrane properties. Pflugers. Arch. 331, 324-334. (Cited in Bernard et aI., 1985.)
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Saracci, R., Kogevinas, M., Bertazzi, P. A, Demesquita, B. H. B., Coggon, D., Green, L. M., Kauppinen, T., Labbe, K A., Littorin, M., Lynge, E., Mathews, J. D., Neuberger, M., Osman, J., Pearce, N., and Winkelmann, R (1991). Cancer mortality in workers exposed to chlorophenoxy herbicides and chlorophenols. Lancet 338(8774), 1027-1032. Sauerhoff, M. W, Braun, W H., Blau, G. E., and Gehring, P. J. (1977). The fate of 2,4-dichlorophenoxyacetic acid (2,4-D) following oral administration to man. Toxicology 8, 3-11. Schultz, S. K, Brock, A. W, and Killeen, J. C. (1990). "Dennal Sensitization Study (Closed Patch Repeated Insult) in Guinea Pigs, Rabbits with Diethanolamine Salt of 2,4-D." Unpublished Report 90-165. Industry Task Force 11 on 2,4-D Research Data. Ricerca Inc., Painesville, OH. Schulze, G. E. (1990a). "21-Dennal Irritation and Dennal Toxicity Study in Rabbits with 2,4-Dichlorophenoxyacetic Acid." HLA Study No. 2184-109. Hazleton Laboratories America Inc., Vienna, VA. Schulze, G. E. (1990b). "2l-Dennal Irritation and Dennal Toxicity Study in Rabbits with Dimethylamine Salt of 2,4-Dichlorophenoxyacetic Acid." HLA Study No. 2184-111. Hazleton Laboratories America Inc., Vienna, VA. Schulze, G. E. (1990c). "21-Dennal Irritation and Dennal Toxicity Study in Rabbits with 2,4-Dichlorophenoxyacetic Acid-2-Ethylhexyl Ester." HLA Study No. 2184-110. Hazleton Laboratories America Inc., Vienna, VA. Schulze, G. E. (1991). "Final Report: Subchronic Toxicity Study in Mice with 2,4-Dichlorophenoxyacetic Acid." Industry Task Force 11 on 2,4-D Research Data. Hazleton Laboratories America, Inc., Vienna, VA. Schulze, G. E., and Dougherty, J. A. (1988). Neurobehavioral toxicity of 2,4-D-n-butyl ester (2,4-D ester): Tolerance and lack of cross-tolerance. Neurotoxicol. Teratol. 10,75-79. Schwetz, B. A, Sparschu, G. L., and Gehring, P. J. (1971). The effect of 2,4dichlorophenoxyacetic acid (2,4-D) and esters of 2,4-D on rat embryonal, foetal and neonatal growth and development. Fd. Cosmet. Toxicol. 9, 801817. Sell, C R, Maitlen, J. C., and Aller, W A. (1982). Perspiration as an important physiological pathway for the elimination of 2,4-dichlorophenoxyacetic acid from the human body. Am. Chem. Soc. Abstr. Pap. 183, PEST #74. Serota, D. G. (1983a). "Subchronic Toxicity Study in Rats with 2,4-D Acid." Unpublished Report 2184-102. Industry Task Force 11 on 2,4-D Research Data. Hazleton Laboratories America, Inc., Vienna, VA. Serota, D. G. (1983b). "Subchronic Toxicity Study in Mice with 2,4-D Acid." Unpublished Report 2184-100. Industry Task Force 11 on 2,4-D Research Data. Hazleton Laboratories America, Inc., Vienna, VA Serota, D. G. (1986). "Combined chronic toxicity and oncogenicity study in rats with 2,4-D acid." Unpublished Report 2184-103. Industry task Force 11 on 2,4-D Research Data. Hazleton Laboratories America, Inc., Vienna, VA. Serota, D. G., et al. (1986). "Combined Chronic Toxicity and Oncogenicity Study in Rats with 2,4-D Acid." Unpublished Report 2184-102. Hazleton Laboratories America, Inc., Vienna, VA Serota, D. G., et al. (1987). "Oncogenicity Study in Mice with 2,4-D Acid." Unpublished Report 2184-101. Hazleton Laboratories America, Inc., Vienna, VA. Serrone, D. M., Killeen, J. C, and Benz, G. (1991). "A Subchronic Toxicity Stndy in Rats with the Diethanolamine Salt of 2,4-Dichlorophenoxyacetic Acid." Unpublished Report 90-0186. Ricera Inc., Painesville, OH. Simmon, V. E, Kauhanen, K, and Tardiff, R. G. (1977). Mutagenic activity of chemicals identified in drinking water. In "Progress in Genetic Toxicology" (B. A. Bridges Scott and E H. Sodels, eds.), Vol. 2, pp. 249-258. ElsevierlNorth-Holland, Amsterdam. Soler-Niedziela, L., Ong, T., Nath, J., and Zeiger, E. (1988). Mutagenicity studies of dioxin and related compounds with the Salmonella arabinose resistant assay system. Toxic. Assess. 3(2), 137-145. Steiss, J. E., Braund, K. G., and Clark, E. G. (1987). Neuromuscular effects of acute 2,4-dichlorophenoxyacetic acid (2,4-D) exposure in dogs. J. Neurol. Sci. 78, 295-301. Styles, J. A. (1973). Cytotoxic effects of various pesticides in vivo and in vitro. Mutal. Res. 21(1), 50-51. Styles, J. A (1977). Mammalian cell transfonnation in vitro. Br. 1. Cancer 37, 931-936.
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Phenoxy Herbicides (2,4-D)
Szabo, J. R., and Rachunek, B. L. (1991). "2,4-D, Butoxyethyl Ester: 13-Week Dietary Toxicity Study in Fischer Rats." ID. DECO-TXT:K-007722-015. DowE1anco: Indianapolis, IN. Toyoshima, E., Mayre, R. E, Max, S. R., and Eccles, C. (1985). 2,4Dichlorophenoxyacetic acid (2,4-D) does not cause polyneuropathy in rats. J. Neurol. Sci. 70, 225-229. Tyynela, K, Elo, H. A., and Ylitalo, P. (1990). Distribution of three common chlorophenoxyacetic acid herbicides into the rat brain. Arch. Toxicol. 64(1), 61-65. Tyynela, K, Elo, H., Ylitalo, P., and Hervonen, H. (1986). The central nervous system toxicity of chlorophenoxyacetic acid herbicides. Arch. Toxicol. (SuppJ. 9), 355. Unger, T. M., Kliethennes, J., Van Goethem, D., and Short, R. D. (1981). "Teratology and Postnatal Studies in Rats of the Propylene Glycol Butyl Ether and Isooctyl Esters of 2,4-Dichlorophenoxyacetic Acid." PB81-191140. U.S. Environmental Protection Agency, Research Triangle Park, NC. Waters, G. D., Simmon, V. E, Mitchell, A. D., Jorgenson, T. A., and Valencia, R. (1980). An overview of short-tenn tests for the mutagenic and carcinogenic potential of pesticides. J. Environ. Sci. Health. BI5(6), 867-906. Wigle, D. T., et al. (1990). Mortality study of Canadian male farm operators: Non-Hodgkin's lymphoma mortality and agricultural practices in Saskatchewan. J. Nat. Cancer Inst. 82(7), 575-582. Wiklund, K, and Holm, L. E. (1986). Soft tissue sarcoma risk in Swedish agricultural and forestry workers. J. Nat. Cancer Inst. 76(2), 229-234. Wiklund, K, et al. (1987). Soft tissue sarcoma risk among agricultural and forestry workers in Sweden. Chemosphere 16(8/9), 2107-2110. Wolfe, W. H. (1983). The epidemiology and toxicology of Agent Orange. In "Proceedings of the 14th Conference on Environmental Toxicology," Dayton,OH. Woolson, E. A., Thomas, R. E, and Ensor, P. D. J. (1972). Survey of polychlorodibenzo-p-dioxin content in selected pesticides. J. Agr. Food Chem. 20(2), 351-354. World Health Organization (WHO) (1975). "Evaluations of Some Pesticides in Food," The Monographs, WHO Pesticide Series No. 4, pp. 159-183. World Health Organization, Geneva.
World Health Organization (WHO) (1984). "2,4-Dichlorophenoxyacetic Acid (2,4-D)." Environmental Health Criteria 29, pp. 1-15l. IPCS International Programme on Chemical Safety, United Nations Environment Programme, International Labour Organisation, and the World Health Organisation. Yano, B. L., Cos se, P. E, Atkin, L., and Corley, R. A. (199Ia). "2,4-D Isopropylamine Salt (2,4-D IPA): A 13-Week Dietary Toxicity Study in Fischer 344 Rats." ID. HET m-004725-006. Dow Elanco, Indianapolis, IN. Yano, B. L., Cosse, P. E, Markham, D. A., and Atkin, L. (l99Ib). "2,4-D Triisopropanolamine Salt (2,4-D IPA): A 13-Week Dietary Toxicity Study in Fischer 344 Rats." ID. K-008866-006. DowElanco, Indianapolis, IN. Yeary, R. A. (1986). Urinary excretion of2,4-D in commercial lawn specialists. Appl. Ind. Hyg. (1)3, 119-12l. Ylitalo, P., Narhi, U., and Elo, H. A. (1990). Increase in the acute toxicity and brain concentrations of chlorophenoxyacetic acids by probenicid in rats. Gen. Pharmacol. 21(5), 811-814. Yoder, J., Watson, M., and Benson, W. W. (1973). Lymphocyte chromosome analysis of agricultural workers during extensive occupational exposure to pesticides. Mutat. Res. 21, 335-340. Zablotny, C. L., Yano, B. L., and Breslin, W. J. (1991). "2,4-D Triisopropanolamine Salt (2,4-D TIPA): A 13-Week Dietary Toxicity Study in Fischer 344 Rats." Unpublished Report No. K-008866-006. Zahm, S. H., et al. (1990). A case-control study of non-Hodgkin's lymphoma and the herbicide 2,4-dichlorophenoxyacetic acid (2,4-D) in eastern Nebraska. Epidemiology 1(5), 349-356. Zahm, S. H. (1997). Mortality study of pesticide applicators and other employees of a lawn care service company. 1. Occup. Environ. Med. 39(11), 1055-1067. Zetterberg, G. (1978). Genetic effects of phenoxy acids on microorganisms. Ecol. Bull. 27, 193-204. Zetterberg, G., Busk, L., Elovson, R., Starec-Nordenhammar, I., and Ryttman, H. (1977). The influence of pH on the effects of 2,4-D (2,4-dichlorophenoxyacetic acid, Na salt) on the Saccharomyces cerevisiae and Salmonella typhimurium. Mutat. Res. 42, 3-18. Zhamsaranova, S. D., Lebedeva, S. N., and Lyashenko, V. A. (1987). The immunodepressive effects of the herbicide 2,4-D in mice. Gig. 1. Sanit. 5, 80--81.
CHAPTER
73 Dicamba Paul Harp Virginia Polytechnic State University
Chemical name: 3, 6-dichloro-2-methoxybenzoic acid or 3,6-dichloro--o-anisic acid.
Structure:
COOH
Cl'0 VCl
OCH3
73.1 SYNONYMS Common names are dicamba (ANSI, BSI, ISO, WSSA) and dianat (USSR). Code numbers include CAS 1918-00-9, CAS 1982-69-0 (for the sodium salt), CAS 2300-66-5 (dicamba dimethylammonium), SHA 029801, and Velsicol 58-CS-ll. Trade names include Banfel®, Banvel®, Banvel® 4S, Banvel® CST, Banvel® D, Banvel® S, Banvel® XG, Clarity® 4S, Metambane®, Veteran® lOG, and Veteran® CST. Combinations of dicamba and other herbicides are marketed in products such as Banlene Plus®, Banlene Solo®, Brushmaster®, Celebrity®, Diptyl®, Distinct®, Dock1ene®, Durtok® 540, Durtok® Amina 270, Fallowmaster®, Field Marshall®, Grassland Herbicide®, Marksman®, Northstar®, Novertex® gazons H, OpTill®, Resolve®, Selectone G®, Trimec® 992, Trimec® Bentgrass, Trimec® Brush Killer, Trimec® Classic, Trimec® Encore, Trimec® Plus, Trimec® S.I., Trimec® Southern, Trimonal®, Tri-Power®, Veteran® 720, Veteran® 2010, Wallop®, and Weedmaster®.
73.2 PHYSICAL AND CHEMICAL PROPERTIES Dicamba in pure form is an odorless white crystalline solid with a melting point of 114-116°C and a vapor pressure at 100°C of 3.75 x 10- 3 mm Hg. The chemical formula is CSH6Clz03 and the molecular weight is 221.04. Technical grade dicamba (80-90% purity) is a crystalline solid with a pale buff color. Dicamba is soluble in water and is resistant to hydrolysis and oxidation under normal environmental conditions. Handbook of Pesticide Toxicology Volume 2. Agents
73.3 FORMULATIONS AND USES Dicamba is a benzoic acid herbicide that acts by mimicking the effects of auxins (i.e., natural plant growth hormones), causing enhanced but uncontrolled growth rates, alterations in plant function homeostasis, and death. Liquid, pellet, and granule formulations are available that contain either the acid or one of several salts (diglycolamine, dimethylamine, dimethylammonium, isopropylamine, potassium, or sodium). As indicated in Section 73.1, dicamba is often combined with one or several other herbicidal agents including 2,4-D, 2,4-DP, atrazine, glyphosate, imazethapyr, ioxynil, MCPA, and mecoprop. It is used to control a wide spectrum of annual and perennial broadleaf weeds and is effective in both pre- and post-emergence applications. A primary agricultural use is weed reduction in grain/cereal crops. Industrial/commercial applications include maintenance of pastures, forest lands, fence rows, and transportation and utility rights-of-way.
73.4 TOXICOKINETICS Dicamba is readily absorbed following ingestion, but only minimal absorption occurs after dermal exposure. Following ingestion, dicamba appears to be rapidly and non selectively distributed to most organ systems. Most animal studies indicate greater than 90% of the original dose ultimately undergoes urinary excretion. Small amounts of dicamba are excreted in the feces and have been shown to represent both true elimination (indicated after subcutaneous administration) and excretion of unabsorbed herbicide (following oral exposure). Dicamba is excreted mostly in unmetabolized form but may also be conjugated with glucuronic acid or glycine. Elimination occurs rapidly, and there is no evidence of bioaccumulation in mammalian systems.
73.5 TOXICITY TO LABORATORY ANIMALS Dicamba has a low mammalian toxicity, with reported oral LDso values in rats ranging from 757 to 2629 mg/kg. Mouse
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Copyright © 2001 by Academic Press. All rights of reproduction in any fonn reserved.
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CHAPTER 73
Dicamba
oral LDso values are greater than 1000 mg/kg, and values ranging from 566 to 3000 mg/kg have been reported for rabbits and guinea pigs. Dermal LDso values in rabbits are greater than 2000mg/kg.
73.6 TOXICITY TO HUMANS Evaluation of the specific adverse effects of dicamba in humans has been confounded by the relatively small amount of data available from exposures exclusively to dicamba. Many of the documented exposures have been to products containing a combination of herbicidal agents. Signs of dicamba intoxication include vomiting, bradycardia, shortness of breath, cyanosis, depression, incontinence, muscular weakness/exhaustion subsequent to muscle spasms (myotonia), and death. Survivors usually recover quickly (within 2-3 days) with no long-term effects. Dicamba is a mildly corrosive sensitizing agent, and skin irritation or bums can follow dermal exposure. Eye contact can result in a temporary clouding of the cornea that may persist for 5 to 7 days. However, more severe or permanent ocular damage is possible and appropriate eye protection should be used, especially when working with concentrated solutions. A study of circulating cholinesterase activity in pesticide applicators has suggested an anticholinesterase action of dicamba (Potter et aI., 1993). Of the 14 herbicide appliers in the study that worked with dicamba, six had significant reductions (> 20%) in red blood cell (RBC) acety!cholinesterase (AChE) activity. The six workers applied an average seasonal volume of 191 gallons of dicamba, while the eight workers without significant AChE inhibition applied an average of only 17 gallons. None of the workers had reductions in plasma butyr!cholinesterase, however, even though in vitro experiments with commercial-grade dicamba showed significant inhibition of both RBC and plasma cholinesterase [ICso ~ 70 ppm (Potter et aI., 1993)]. Contamination of the commercial-grade dicamba with impurities capable of inhibiting cholinesterase could explain the apparent anticholinesterase activity of dicamba. Conversely, the structural similarity to a group of amphiphilic agents known to inhibit RBC AChE may indicate an ability of dicamba to alter the conformational state of the membranebound enzyme.
73.7 REPRODUCTIVE EFFECTS No adverse effects were measured in a three-generation study in rats treated with dicamba.
73.8 GENOTOXIC EFFECTS Dicamba has been examined for possible carcinogenic or mutagenic activity with somewhat contradictory and inconclusive results. In early studies, rats chronically exposed to dicamba
(25 mg/kg/day for 2 years) did not exhibit an increase in tumor formation (U.S. EPA, 1988). Perocco and co-workers (1990), however, reported a DNA-damaging potential of dicamba based on an enhanced unwinding rate of rat liver DNA following in vivo dicamba administration. Using in vitro experiments with human peripheral blood lymphocytes, dicamba induced unscheduled DNA synthesis and slightly increased the frequency of sister chromatid exchange. Conversely, several studies prior to the work of Perocco et al. (1990) reported negative results for mutagenicity using various non-mammalian-based assay systems. More recent work (Hrelia et aI., 1994) with mammalian tissue also failed to detect any clastogenic activity of dicamba, but the difference in results compared with Perocco et al. (1990) may reflect differences associated with exposure route (in vivo versus in vitro). Dicamba has been shown to induce peroxisomal enzymes in rat liver and cause transcriptional upregulation of the peroxisome proliferator-activated receptor (Espandiari et aI., 1995). Although tumorigenic effects of dicamba have not been reported, long-term exposure to peroxisome proliferators has been associated with increased hepatic tumor frequency in rodents. Further investigation also revealed activation of hepatic nuclear factor-K B, a transcription factor that may be involved in the hepatocarcinogenic action of certain peroxisome proliferators (Espandiari et aI., 1998). The exact implications of these findings are unclear and may represent actions of dicamba that require further study in order to fully understand any possible long-term consequences of chronic exposure to the herbicide.
73.9 TREATMENT OF POISONING Treatment is symptomatic and supportive. No specific antidote is available.
REFERENCES Espandiari, P., Ludewig, G., Glauert, H. P., and Robertson, L. W. (1998). Activation of hepatic NF-KB by the herbicide dicamba (2-methoxy-3,6dichlorobenzoic acid) in female and male rats. 1. Biaehem. Mal. Taxieal. 12,339-344. Espandiari, P., Thomas, V. A., Glauert, H. P., O'Brien, M., Noonan, D., and Robertson, L. W. (1995). The herbicide dicamba (2-methoxy-3,6dichlorobenzoic acid) is a peroxisome proliferator in rats. Fundam. Appl. Taxieol. 26, 85-90. Hrelia, P., Vigagui, F., Maffei, F., Morotti, M., Colacci, A., Perocco, P., Grilli, S., and Cantelli-Forti, G. (1994). Genetic safety evaluation of pesticides in different short-term tests. Mutat. Res. 321, 219-228. Perocco, P., Ancora, G., Rani, P., Valenti, A. M., Mazzullo, M., Colacci, A., and Grilli, S. (1990). Evaluation of genotoxic effects of the herbicide dicamba using in vivo and in vitro test systems. Enviran. Mal. Mutagen. 15, 131-135. Potter, W. T., Garry, V. F., Kelly, 1. T., Tarone, R., Griffith, l., and Nelson, R. L. (1993). Radiometric assay of red cell and plasma cholinesterase in pesticide appliers from Minnesota. Toxieol. Appl. Pharmacal. 119, 150-155. V.S. Environmental Protection Agency (1988). "Dicamba: Health Advisory." Office of Drinking Water, V.S. EPA, Washington, DC.
CHAPTER
74 Imidazolinones* Frederick G. Hess and Jane E. Harris BASF Corporation
Kimbedy Pendino Hoffman-La Roche, Inc.
Kathryn Ponnock Middlesex County Community College
74.1 IDENTITY, PROPERTIES, AND USES 74.1.1 CHEMICAL NAMES Imazapyr (Arsenal® herbicide): 2-(4-isopropyl-4-methyl-5oxo-2-imidazolin-2-yl) nicotinic acid. Imazamethabenz-methyl (Assert® herbicide): methyl 2-(4isopropy 1-4-methy l-5-oxo-2-imidazolin-2-yl)- p-toluate mixed with methyl 6-(4-isopropyl-4-methyl-5-oxo-2-imidazolin-2yl)-m-toluate (3 : 2). Imazapic (Cadre® herbicide): 2-(4-isopropyl-4-methyl-5oxo-2-imidazolin-2-yl)-5-methylnicotinic acid. Imazethapyr (Pursuit® herbicide): 5-ethyl-2-(4-isopropyl-4methyl-5-oxo-2-imidazolin-2-yl) nicotinic acid. Imazamox (Raptor® herbicide): 2-(4-isopropyl-4-methyl-5oxo-2-imidazolin-2-yl)-5-(methoxymethyl) nicotinic acid. Imazaquin (Scepter® herbicide): 2-(4-isopropyl-4-methyl-5oxo-2-imidazolin-2-yl)-3-quinoline carboxylic acid. 74.1.2 PHYSICAL AND CHEMICAL PROPERTIES
respective empirical formulas, molecular weights (range from 261-311), physical state (off-white to tan powders), melting points (range from 144-222°C), and vapor pressures « 1 x 10- 7 mm Hg) are very similar. One might expect that such close similarities in physical properties would be reflected in similarities of toxicological profiles, as discussed in Section 74.2. The very low vapor pressures of the imidazolinones indicate a low potential to volatilize. With the exception of imazamethabenz-methyl, which is an ester, the imidazolinones have pKa values (dissociation constants) forthe carboxylic acid group that range from 3.0 to 3.5. Based on these dissociation constants, both the water solubility and n-octanollwater partition coefficients are pH dependent. For example, if the localized pH is above approximately 3.5, the amount of the imidazolinone in ionized form increases, which consequently increases water solubility and decreases the n-octanollwater partition coefficients. In contrast, if the local pH is below approximately 3.0, the amount of the imidazolinone in ionized form decreases, which consequently decreases water solubility and increases the n-octanollwater partition coefficients. In addition, the imidazolinone herbicides demonstrate other dissociation constants at approximately 2 and 11, which, in turn, alter the specific water solubility and n-octanollwater partition coefficients.
74.1.3 STRUCTURE The physical and chemical properties of the six imidazolinone herbicides (technical products) are presented in Table 74.1. The *The summaries and evaluations contained in this chapter are, in most cases, based on unpublished proprietary data. A registration authority should not grant a registration on the basis of this information unless it has first received authorization for such use from the owner of the data or has received the data on which these summaries are based, either from the owner of the data or from a second party that has obtained permission from the owner of the data for this purpose. Handbook of Pesticide Toxicology Volume 2. Agents
The chemical structures of the respective imidazolinone herbicides bear a close resemblance to each other, as presented in Fig. 74.1. The active molecule of each compound has an identical imidazolinone ring structure with a carboxylic acid group, or specifically a carboxylic ester group for imazamethabenzmethyl, attached to respective backbone groups, such as a pyridine ring.
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Copyright © 2001 by Academic Press. All rights of reproduction in any fonn reserved.
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CHAPTER 74 Imidazolinones
Table 74.1 Comparison of the Physical and Chemical Properties of the Imidazolinone Herbicides Technical herbicide ImazamethabenzProperties
Imazapyr
methyl
Imazapic
Imazethapyr
Imazamox
Imazaquin
Empirical formula
C13 H IS N30 3 261
Cl6 H 20 N2 0 3 288
C14 H 17 N30 3 275
CISHl9 N 30 3
Molecular weight
289
CISHl9 N4 0 3 305
C17 H 17 N30 3 311
Physical state
White-to-tan
Off-white powder
Off-white-to-tan
Light-tan powder
Off-white powder
Light-tan powder
powder
powder
Melting point
169-173°C
144-153°C
204-206°C
177-179°C
166-167°C
219-222°C
Vapor pressure
<1 x 10- 7 mmHg
1.13 x 10- 8 mmHg
<1 x 10- 7 mmHg
<1 x 10-7 mmHg
<1 x 10- 7 mmHg
<2 x 10- 8 mmHg
at 60°C
at 25°C
at 60°C
Several of the compounds show very similar structureactivity relationships. For instance, imazapic, imazethapyr, imazamox, and imazaquin differ only by the respective sub-
irnazapyr (ARSENAL@herbicide)
(BAS 693 H)
at 60°C
at 60°C
at 45°C
stituent groups that are attached to the pyridine ring (Fig. 74.1). Imazapic has a methyl group, whereas imazethapyr has an ethyl group, imazamox has a methoxymethyl group, and imazaquin has a benzene ring fused to the pyridine ring. One might expect that such close resemblance in chemical structures would be reflected in similarities of toxicological profiles, as discussed in Section 74.2.
74.1.4 HISTORY AND USES irnazarnethabenz-rnethyl (ASSERT@herbicide)
(BAS 712 H)
imazapic (CADRE@herbicide)
(BAS 715 H)
irnazethapyr (PURSUrr@herbicide)
(BAS 685 H)
CH'~COOH o
imazamox (RAPTOR® herbicide)
I
'>-...~ N
~
(BAS 720 H)
HN
o
imazaquin (SCEPTER® herbicide)
(BAS 725 H)
®
Registered Trademark of BASF Corporation
Figure 74.1
Chemical structures of the imidazolinone herbicides.
The unique class of synthetic chemical compounds called the imidazolinone herbicides was discovered in the 1970s, with the first D.S. patent awarded in 1980 for imazamethabenzmethyl (Assert®). Other imidazolinones in the series, imazapyr (Arsenal®), imazapic (Cadre®), imazethapyr (Pursuit®), and imazaquin (Scepter®), received D.S. patents in 1989. More recently, imazamox (Raptor®) received a D.S. patent in 1994. The active ingredients of the imidazolinone herbicides are effective for selective postemergent weed control for the following crops: cereals, including wheat and barley (imazamethabenz-methyl); peanuts (imazapic); soybeans (imazamox, imazaquin, and imazethapyr); and other legumes, including peas, beans, and alfalfa (imazamox and imazethapyr). In addition, imazapyr is a broad-spectrum herbicide that is effective for noncrop uses for total vegetation control on industrial sites and railroad, highway, and utility rights-of-way and for forestry applications. Most recently, the imidazolinone herbicides have shown specificity for postemergent weed control on imidazolinonetolerant corn called Imi-Com® (imazethapyr and imazapyr), Smart™canola (imazethapyr and imazamox), imidazolinonetolerant rice (imazethapyr), imidazolinone-tolerant sugar beet (imazamox), and imidazolinone-tolerant wheat (imazamox). After application, each imidazolinone herbicide is taken up by the foliage and/or roots of the susceptible weed, and subsequently, becomes translocated throughout the plant. Susceptible weeds stop growing and competing with the specific crop shortly after translocation; the weeds die within several weeks postapplication.
74.2 Toxicity to Laboratory Animals
74.2 TOXICITY TO LABORATORY ANIMALS Following extensive testing in the required mammalian toxicity studies, the six imidazolinone herbicides demonstrate a low toxicological potential. The results of these studies with the respective technical products are presented in Table 74.2. The imidazolinone herbicides demonstrate this very low toxicity profile in mammals because their herbicidal activity relies on their mode of action, that is, the inhibition of the specific plant enzyme, acetohydroxyacid synthase (AHAS). AHAS is an important biosynthetic enzyme in the formation of three branched-chain aliphatic amino acids, namely, isoleucine, leucine, and valine, and is found in plant but not mammalian tissues. This inhibition of AHAS disrupts protein synthesis and subsequently interferes with DNA synthesis and cell growth, eventually leading to cell death in the specific weed, as described in Section 74.1.
74.2.1 BASIC FINDINGS 74.2.1.1 Acute Toxicity Studies The results of the acute toxicity tests indicate that the technical products of the imidazolinone herbicides are generally relatively nontoxic (D.S. EPA Toxicity Category IV) by the oral and inhalation routes of administration and only slightly toxic (Category Ill) by the dermal route, according to the respective LDso and LCso values presented in Table 74.2. Further, the technical products of this class of chemical compounds are either nonirritating or only slightly irritating in the rabbit primary skin irritation studies. The skin sensitization studies in guinea pigs, conducted according to the method of Buehler, demonstrate that the technical products of the six imidazolinone herbicides are nonsensitizers (Table 74.2). Results from the rabbit primary eye irritation studies with the technical products of the imidazolinone herbicides ranged from no irritation (imazaquin) to slightly irritating (imazamethabenzmethyl and imazamox) to moderately irritating (imazapic and imazethapyr), showing complete recovery by day 7 postdosing. Consequently, these five imidazolinone herbicides are classified into D.S. EPA Toxicity Category III or IV. Finally, the rabbit primary eye irritation study with imazapyr demonstrated irreversible irritation (Category I) based on two out of six animals with scattered opacities at day 21. Thus, the labels for imazapyr's formulated products recommend protective eyewear, which should mitigate any potential for eye irritation during mixing/loading and application.
74.2.1.2 Short-TermlSubchronic Toxicity Studies 21-Day Dermal Studies in the Rabbit All six imidazolinone herbicides showed no dermal or systemic toxicity following 21 days of dermal exposure (6 hours per day, 5 days per week) at the highest doses tested, supporting no observable effect levels (NOELs) at the highest doses tested. Specifically,
1643
these no-effect levels were 1000 mg/kg body weight/day for imazapic, imazethapyr, imazamox (28-day study), and imazaquin; and 400 and 200 mg/kg body weight/day for imazapyr and imazamethabenz-methyl, respectively.
90-Day Dietary Toxicity Studies in the Dog Subchronic (90-day) feeding studies were conducted in dogs with imazapyr, imazethapyr, and imazamox. For these subchronic (90-day) feeding studies, the imidazolinone herbicides showed no systemic toxicity at the highest concentrations tested. Specifically, the NOELs were 10,000 ppm for imazapyr and imazethapyr and 40,000 ppm for imazamox. These highest dietary concentrations are equivalent to approximately 250 mg/kg body weight/day for imazapyr and imazethapyr, and equivalent to an approximate daily intake value of 1370 mg/kg body weight/day for imazamox, as calculated from food consumption data. 90-Day Dietary Toxicity Studies in the Rat For the subchronic (90-day) feeding studies in rats, most of the imidazolinone herbicides showed no systemic toxicity when tested at very high dietary concentrations of 20,000 ppm (imazapyr, imazapic, and imazamox) or 10,000 ppm (imazethapyr and imazaquin). Moreover, these highest dietary concentrations are equivalent to approximate daily intake values of 1740 mg/kg body weight/day (imazapyr), 1625 mg/kg body weight/day (imazapic), 1660 mg/kg body weight/day (imazamox), 820 mg/kg body weight/day (imazethapyr), and 830 mg/kg body weight/day (imazaquin), as calculated from food consumption data. Only imazamethabenz-methyl induced several mild treatment-related effects following 90 days of dietary exposure at the two highest concentrations tested, 5000 and 10,000 ppm. Specifically, slight but consistent decreases in mean body weight occurred for males and females at 5000 and 10,000 ppm, as compared to controls. These reductions in body weights resulted in decreases in overall body weight gain of 8% and 6% for males at 5000 and 10,000 ppm, respectively. For females at both 5000 and 10,000 ppm, overall body weights were decreased by 5%, as compared to controls. No effects on mean body weight or body weight gain were noted for males or females at 1000 ppm, the lowest dietary concentration tested. In addition, statistically significant increased relative (to body weight) liver weights were observed for males at 10,000 ppm, as compared to controls. Increased incidences of hepatocellular hypertrophy were noted for male rats at 5000 ppm (14/19) and 10,000 ppm (20120), as compared to controls (0120). The microscopic change of hepatocellular hypertrophy may represent an adaptive response in the liver, that is, induction of microsomal enzymes, which is generally not considered to be a toxic effect (Popp and Cattley, 1991). It is known that this morphological change and associated enzyme induction are reversible following withdrawal of chemical treatment (Popp and Cattley, 1991). In conclusion, for the 90-day dietary toxicity study in rats with imazamethabenz-methyl, the no-effect level was the lowest concentration tested, 1000 ppm (equivalent to approx-
Table 74.2 Comparison of Mammalian Toxicity Data and Genotoxicity Data for the Imidazolinone Herbicides Technical herbicide ImazamethabenzStudy
Imazapyr
methyl
Imazapic
Imazethapyr
Imazamox
Imazaquin
>5000 mg/kg
>5000mg/kg
>5000 mg/kg
>5000mg/kg
>5000mg/kg
>5000mg/kg
Acute toxicity Acute oral toxicity (rat) LDSO
Acute dermal
...... ~
0\
(relatively nontoxic)
(relatively nontoxic)
(relatively nontoxic)
(relatively nontoxic)
(relatively nontoxic)
(relatively nontoxic)
(Category IV)
(Category IV)
(Category IV)
(Category IV)
(Category IV)
(Category IV)
>2000mglkg (slightly toxic)
>4000 mglkg (slightly toxic)
>2000 mg/kg
>2000 mg/kg
>2000 mg/kg
toxicity
(slightly toxic)
(slightly toxic)
(slightly toxic)
(rabbit) LDso
(Category III)
(Category Ill)
(Category Ill)
Acute inhalation
> 1.3 mgll
> 1.4 mgll
>4.8 mgll
(Category II1) >3.3 mg/l
(Category III) >6.3 mgll
>2000 mg/kg (slightly toxic) (Category Ill) >5.7 mgll
toxicity
(slightly toxic)
(slightly toxic)
(relatively nontoxic)
(relatively nontoxic)
(relatively nontoxic)
(relatively nontoxic)
(rat) LCSO
(Category Ill)
(Category III)
(Category IV)
(Category IV)
(Category IV)
(Category IV)
(analytical) Primary dermal irritation
Slight irritation (Category Ill)
Nonirritating (Category IV)
Slight irritation
Slight irritation
(Category IV)
(Category IV)
Moderate irritation
Moderate irritation
(Category III)
(Category Ill)
Slight irritation (Category IV)
Slight irritation (Category Ill)
(rabbit) Primary eye irritation
Irreversible irritation (Category I)
Slight irritation (Category Ill)
Slight irritation (Category III)
Nonirritating (Category IV)
(rabbit) Dermal sensitization
Nonsensitizer
Nonsensitizer
Nonsensitizer
Nonsensitizer
N onsensitizer
Nonsensitizer
(guinea pig) (Buehler method) (continues)
Table 74.2 (continued)
Technical herbicide ImazamethabenzStudy
Imazapyr
methyl
Imazapic
Imazethapyr
Imazamox
Imazaquin
400 mg/kg b.w.lday
200 mglkg b. w.lday (HDT)
1000 mglkg b.w.lday (HOT)
1000 mglkg b.w.lday (HOT)
1000 mglkg b.w.lday (HOT)
1000 mg/kg b.w./day (HOT)
Short-termJ subchronic toxicity 2l-day dermal (Rabbit)
(HDT)
(HDT)
(HOT)
(28-day dermal)
No observable effect level 90-day dietary 1-0
0'1
,J;o.
VI
(dog)
10,000 ppm (HeT)
10,000 ppm (HeT)
(~250
(;:::;250 mg/kg b.w./day)
mglkg b.w./day)
40,000 ppm (HCT) (1370 mglkg b.w./day) (based on Fe data)
No observable effect level 90-day dietary (rat) No observable
20,000 ppm (HCT)
1000 ppm
20,000 ppm (HCT)
10,000 ppm (HCT)
20,000 ppm (HCT)
10,000 ppm (HCT)
(1740 mglkg b.w.lday)
(87.5 mglkg b.w.lday)
(1625 mg/kg b.w./day)
(820 mglkg b.w.lday)
(1660 mg/kg b.w.lday)
(830 mglkg b.w./day)
(based on FC data)
(based on FC data)
(based on FC data)
(based on FC data)
(based on FC data)
(based on FC data)
effect level Chronic toxicity I-year dietary (dog)
10,000 ppm (HCT) (;:::;250 mg/kg b.w.lday)
1000 ppm (~25
mg/kg b.w./day)
No observable adverse effect level
1000 ppm (~25
mg/kg b.w.lday)
40,000 ppm (HCT) (1165 mg/kg b.w.lday)
looOppm (~25
mg/kg b.w./day)
= 5000ppm
No observable effect level
(135 mglkgb.w.lday)
(based on FC data)
(based on FC data) (continues)
Table 74.2 (continued)
Technical herbicide ImazamethabenzImazapyr
Study
methyl
Imazapic
Imazethapyr
Imazamox
Imazaquin
No observable
7000 ppm (HCT)
5000 ppm (~750 mglkg b.w.lday)
7000 ppm (HeT)
1000 ppm (~150 mg/kg b.w./day)
Chronic toxicityl oncogenecity 18-month dietary (mouse) Systemic toxicity No observable
10,000 ppm (HCT) (;::::;1500 mg/kg b.w./day)
Adverse effect level = 525 ppm
effect level
(~79
Oncogenicity
""""
~ ~ ~
No observable
10,000 ppm (HeT) (~1500
(1200 mg/kg b.w.lday) (based on FC data)
(based on FC data)
mg/kg b.w./day)
2100 ppm (HeT)
mglkg b.w./day)
(1135 mglkg b.w./day)
(~315
mglkg b.w./day)
7000 ppm (HCT) (1135 mg/kg b. w.lday)
10,000 ppm (HeT) (;::::;1500 mglkg b.w./day)
(1200 mg/kg b.w./day)
4000 ppm (HeT) (;::::;600 mglkg b.w./day)
(based on Fe data)
(based on FC data)
effect level
7000 ppm (HeT)
2-year dietary (rat) Systemic toxicity No observable effect level
10,000 ppm (HeT)
20,000 ppm (HeT)
250ppm
(500 mg/kg b. w./day -
(~12.5
mg/kg b.w.lday)
males; 640 mglkg b.w./day -
(1030 mg/kg b.w./day) (based on Fe data)
10,000 ppm (HeT) (;::::;500 mg/kg b.w.lday)
20,000 ppm (HeT) (1165 mg/kg b.w./day)
10,000 ppm (HeT) (;::::;500 mg/kg b.w./day)
(based on Fe data)
females)
(based on Fe data) Oncogenicity No observable effect level
10,000 ppm (HeT) (500 mg/kg b.w.lday males; 640 mglkg
4000 ppm (HeT) (~200
mg/kg b.w.lday)
20,000 ppm (HeT) (lO30 mglkg b.w./day) (based on Fe data)
10,000 ppm (HCT) (;::::;500 mglkg b.w./day)
20,000 ppm (HeT) (1165 mglkg b.w.lday)
lO,OOO ppm (HCT) (;::::; 500 mglkg b.w./day)
(based on Fe data)
b.w.lday - females) (based on Fe data) (continues)
Table 74.2 (continued)
Technical herbicide ImazamethabenzStudy
Imazapyr
methyl
Imazapic
Imazethapyr
Imazamox
Imazaquin
400 mglkg b. w.lday
500 mg/kg b. w./day
500 mglkg b. w./day
300 mglkg b. w./day
300 mg/kg b. w.lday
250 mglkg b. w.lday
500 mglkg b.w.lday
700 mglkg b.w./day
1000 mglkg b.w./day
900 mglkg b.w./day
500 mg/kg b. w./day
Developmental and reproductive toxicity Teratology (rabbit) Maternal No observable
(HDT)
effect level Developmental No observable
400 mglkg b. w.lday
(HDT)
(HDT)
(HDT)
(HDT)
(HDT)
effect level Teratology (rat) Maternal
300 mglkg b. w.lday
No observable
adverse effect level
effect level
b.w.lday Developmental No observable
1000 mglkg b.w.lday (HDT)
375 mg/kg b. w./day
500 mg/kg b. w./day
500 mglkg b.w./day
1000 mg/kg b.w.lday
1125 mg/kg b.w.lday
1000 mglkg b.w.lday
500 mglkg b. w.lday
= 1000 mglkg
"""" .f;;;o.
Q\
-...1
No observable
(HOT)
1000 mg/kg b.w.lday (HDT)
1000 mglkg b.w.lday
10,000 ppm (HCT)
4000 ppm (HCT)
(HDT)
(HDT)
(HDT)
effect level Reproduction (multigeneration) (rat)
(800 mglkg b.w.lday -
Reproductive
males; 980 mg/kg
Toxicity
b.w.lday - females)
No observable
(~320
mglkg b.w.lday)
20,000 ppm (HCT) (1600 mg/kg b.w.lday)
10,000 ppm (HCT) (~800
mg/kg b.w.lday)
20,000 ppm (HCT) (1640 mglkg b.w.lday
10,000 ppm (HCT) (~800
mg/kg b.w.lday)
(based on FC data)
(based on FC data)
(based on FC data)
effect level Genotoxicity [Gene mutations (Ames and
Not mutagenic
Not mutagenic
Not mutagenic
Not mutagenic
Not mutagenic
or genotoxic
or genotoxic
or genotoxic
or genotoxic
or genotoxic
mammalian cell); in vitro structural chromosomal aberrations; in vivo cytogenetics (mouse micronucleus assay)] Categories are EPA Toxicity Categories. b.w., body weight; HDT, highest dose tested; HCT, highest concentration tested;
~,
approximately equal to; FC, food consumption.
Not mutagenic or genotoxic
1648
CHAPTER 74
ImidazoIinones
imately 87.5 mg/kg body weight/day, as based on actual food consumption data). 74.2.1.3 Chronic Toxicity Studies I-Year Dietary Toxicity Study in the Dog For the I-year chronic dog studies, imazapyr and imazamox demonstrated no treatment-related effects following dietary exposure at the highest concentration tested (HCT), namely: 10,000 ppm (equivalent to approximately 250 mg/kg body weight/day) for imazapyr and 40,000 ppm (approximately 1165 mg/kg body weight/day, as based on food consumption data) for imazamox (see Table 74.2). For the other imidazolinone herbicides, imazethapyr, imazapic, imazaquin, and imazamethabenzmethyl, slight treatment-related effects were observed in the respective I-year chronic dog studies. Specifically, for the I-year chronic feeding study in dogs with imazethapyr using dietary concentrations of 1000, 5000, and 10,000 ppm, the only treatment-related effects were indicative of a slight anemia, that is, decreased red cell parameters [statistically significant decreases in hematocrit, hemoglobin, red blood cell (RBC) count, mean corpuscular hemoglobin (MCH), mean corpuscular volume (MCV), and mean corpuscular hemoglobin concentration (MCHC)], which were observed at weeks 26 and 52 for females at the mid-concentration (5000 ppm) and the high concentration (10,000 ppm). For this chronic dog study with imazethapyr, no treatment-related histopathological lesions were observed at any dietary concentration, including the highest concentration tested (l 0,000 ppm). The no observable effect level for the study was lOOO ppm (equivalent to approximately 25 mg/kg/day). Similar effects indicative of anemia were observed in the I-year chronic dietary dog study with imazapic. In this study, the anemia occurred in both sexes and at higher dietary concentrations of 20,000 and 40,000 ppm. At the highest concentration tested (40,000 ppm), decreased red blood cell parameters (statistically significant decreases in hematocrit, hemoglobin, RBC count, MCH, MCV, and MCHC) were observed at weeks 5, 6, 13, 26, and 52 for males and females with accompanying statistically significant increased numbers of normoblasts and reticulocytes, as compared to controls. In addition, in the 40,000 ppm treatment group, increased incidences of erythropoiesis were observed microscopically in the bone marrow (5 of 6 males, 6 of 6 females) and spleen (4 of 6 males, 2 of 6 females), as compared to controls (0 of 6 males, 0 of 6 females for both bone marrow and spleen). For male and female dogs at 20,000 ppm, only transient statistically significant decreases in red blood cell parameters were observed at sporadic time points throughout the study. At the terminal sacrifice, generally slightly increased incidences of erythropoiesis were diagnosed microscopically in the bone marrow of both sexes (2 of 6 males, 1 of 6 females) and spleen of males (1 of 6 males, 0 of 6 females), as compared to controls (0 of 6 males, 0 of 6 females). For male and female dogs at 5000 ppm, no treatment-related effects indicative of anemia were observed. Further, for most dogs (4 of 5 males and 5 of 6 females) at the 40,000-ppm concentration in the chronic dog study
with imazapic, slight-to-moderate skeletal muscle myopathy was diagnosed microscopically at the terminal sacrifice of 52 weeks, which was preceded by transient increases (beginning at week 5) in the blood serum of the following enzymes contained in skeletal muscle: creatine kinase, aspartate aminotransferase, and lactate dehydrogenase. In addition, one male dog at 40,000 ppm, that died during the study (week 33) with nontreatment-related bronchopneumonia, showed moderate-tomarked skeletal muscle degeneration on microscopic examination. In contrast, only a limited presence of skeletal myopathy of minimal severity was diagnosed at 20,000 ppm (5 of 6 males and 2 of 6 females) and at 5000 ppm (5 of 6 males and 1 of 6 females). The skeletal myopathy observed at both 5000 and 20,000 ppm was not considered to be adverse because the limited presence of minimal skeletal myopathy at both concentrations was evidenced by only a few fibers out of hundreds evaluated per section per animal. Furthermore, these focal myopathies of minimal severity were not consistently diagnosed in all skeletal muscle sites examined per dog (i.e., vastus and abdominal muscles, diaphragm, and esophagus). Moreover, none of the dogs on study, including the 6 males and 6 females at 40,000 ppm, showed any clinical observations during the I-year study to indicate any muscle dysfunction. Finally, at the 5000and 20,OOO-ppm dietary concentrations, no statistically or biologically significant elevations occurred during the course of the I-year study in serum enzymes that are normally increased with skeletal muscle myopathy (i.e., creatine kinase, aspartate aminotransferase, or lactate dehydrogenase). For the reasons given previously, the minimal myopathy diagnosed histologically at 5000 and 20,000 ppm was not considered to impair or adversely affect the functional capacity of the affected skeletal muscles. Thus, based on the slight anemia observed at the midconcentration (20,000 ppm), the lowest dietary concentration (5000 ppm) was regarded as the no observable adverse effect level (NOAEL), equivalent to 135 mg/kg body weight/day, as calculated from actual food consumption data. Similar skeletal muscle myopathies were observed for another imidazolinone herbicide, imazaquin, having similar structure-activity relationships to imazapic and imazethapyr, as mentioned in Section 74.1. For the I-year chronic feeding study in dogs with imazaquin, the dietary concentration of 5000 ppm (highest concentration tested) induced slight anemia as evidenced by bone marrow hyperplasia (erythropoiesis), observed microscopically in 2 of 8 males and 4 of 8 females, and slight skeletal myopathy, observed microscopically in 7 of 8 males and 3 of 8 females. In addition, decreased red blood cell parameters indicative of anemia (statistically significant decreases in hematocrit, hemoglobin, RBC count, MCH, MCV, and MCHC) were observed at weeks 13, 26, and 52 for males and females. Additionally, at 5000 ppm, clinical chemistry parameters indicative of slight skeletal myopathy included statistically significant increased mean serum enzyme levels of creatine kinase, aspartate aminotransferase, and lactate dehydrogenase at weeks 13, 26, and/or 52 for males and females. The NOEL for systemic toxicity for this chronic dog study with imazaquin
74.2 Toxicity to Laboratory Animals was the mid-concentration of 1000 ppm (equivalent to approximately 25 mg/kg body weight/day), which is the same NOEL in the I-year dog study with imazethapyr. Finally, for the I-year chronic toxicity study in dogs with imazamethabenz-methyl, males and females were fed dietary concentrations of 0, 250, 1000, or 4000 ppm. Body weights at the highest concentration were consistently, but not statistically, lower than the controls for males throughout the study and for females from weeks 1-28. This reduction in body weight resulted in an 11 % decrease in overall body weight gain for males at 4000 ppm. There were no treatment-related effects on hematology or clinical chemistry parameters and no treatmentrelated gross necropsy or microscopic findings. Slight but statistically significant increases in absolute and relative (to body weight) liver weights were noted for females at 4000 ppm, as compared to controls. However, these slight increases in absolute and relative (to body weight) liver weights for females at 4000 ppm are considered to be equivocal because of the absence of statistically significant increases in relative (to brain) liver weights for females at 4000 ppm and the lack of treatmentrelated histopathological findings in the liver of males or females at 4000 ppm. The absence of hepatocellular hypertrophy in the liver of dogs at 4000 ppm suggests that the small increase in liver weight may not be treatment related. Based on the slight decreases in mean body weights and an 11 % decrease in overall body weight gain for males at 4000 ppm, the no observable effect level was 1000 ppm (equivalent to approximately 25 mg/kg body weight/day).
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74.2.1.4 Chronic Toxicity/Oncogenicity Studies
considered to be adverse because no correlating histopathological changes were noted in the adrenal, thyroid, or parathyroid glands at any dietary concentration in the study. Based on decreased mean body weights and body weight gain for females at 2100 ppm, the no observable adverse effect level (NOAEL) for systemic toxicity was 525 ppm (equivalent to approximately 79 mg/kg body weight/day). Imazethapyr was tested in the mouse for 18 months at dietary concentrations of 0, 1000, 5000, and 10,000 ppm. The only treatment-related effect occurred at the highest concentration tested (10,000 ppm); that is, decreased overall mean body weight gain was noted in both sexes (14% for males and 24% for females) at 10,000 ppm, as compared to controls. Therefore, the NOEL for systemic toxicity was 5000 ppm (equivalent to approximately 750 mg/kg body weight/day). Lastly, imazaquin was tested in the mouse for 18 months at dietary concentrations of 0, 250, 1000, and 4000 ppm. The only treatment-related effect occurred at the highest concentration tested (4000 ppm); that is, decreased mean body weights were noted at 4000 ppm, as compared to controls. Specifically, mean body weights of males at 4000 ppm were statistically significantly decreased during the first 12 weeks of the study, resulting in an 8% decrease in body weight gain for weeks 1-12, as compared to control males. In addition, mean body weights of females at 4000 ppm were statistically significantly lower than controls throughout the study, resulting in a decrease in overall mean body weight gain of 15%, as compared to control females. Therefore, the NOEL for systemic toxicity was 1000 ppm (equivalent to approximately 150 mg/kg body weight/day).
18-Month Chronic Toxicity/Oncogenicity Studies in the Mouse For the I8-month mouse feeding studies, all six of the imidazolinone herbicides showed no evidence of potential oncogenicity at the highest dietary concentrations of technical materials tested, namely, 10,000 ppm (imazapyr and imazethapyr), 7000 ppm (imazapic and imazamox), 4000 ppm (imazaquin), and 2100 ppm (imazamethabenz-methyl). Furthermore, three of the imidazolinone herbicides showed no systemic toxicity at the highest dietary concentrations tested, namely, imazapyr (at 10,000 ppm) and both imazapic and imazamox (at 7000 ppm). The other three imidazolinones showed mild systemic toxicity in the I8-month mouse studies. Specifically, imazamethabenz-methyl was tested in the mouse at dietary concentrations of 0, 130,525, and 2100 ppm. Mean body weights for females at 2100 ppm were statistically significantly lower than the controls during the first 8 weeks of the study, which resulted in a 7% decrease in overall body weight gain. At terminal sacrifice, mean absolute and relative (to body weight) thyroid/parathyroid weights were slightly but statistically significantly increased for females at 525 ppm and for both males and females at 2100 ppm, as compared to controls. In addition, mean absolute and relative (to body weight) adrenal gland weights were slightly but statistically significantly increased for males at 2100 ppm. However, these slight organ weight changes are not
2-Year Chronic Toxicity/Oncogenicity Studies in the Rat In chronic studies performed at high dietary concentrations for 2 years in the rat, all six of the imidazolinone herbicides showed no evidence of potential oncogenicity/carcinogenicity. Specifically, no treatment-related increased incidences of benign or malignant tumors were induced by the imidazolinones at the highest concentrations tested, namely, 20,000 ppm (imazapic and imazamox), 10,000 ppm (imazapyr, imazethapyr, and imazaquin), and 4000 ppm (imazamethabenz-methyl). Furthermore, five of the six imidazolinone herbicides showed no systemic toxicity at the highest dietary concentrations tested, namely, imazapic and imazamox (at 20,000 ppm) and imazapyr, imazethapyr, and imazaquin (at 10,000 ppm). Only imazamethabenz-methyl showed mild systemic toxicity in the 2-year rat feeding study. Specifically, imazamethabenz-methyl was tested in the rat at dietary concentrations of 0, 250, 1000, and 4000 ppm. Mean body weights were slightly but consistently decreased for males and females at 4000 ppm, and for females at 1000 ppm, as compared to controls. For males at 4000 ppm a decreased overall mean body weight gain of 6% was noted, as compared to controls. For females at 1000 and 4000 ppm, respective decreased overall mean body weight gains of 6% and 11 % were noted, as compared to controls. There were no treatment-related microscopic lesions with the exception of a statistically significant increased incidence of
1650
CHAPTER 74 Imidazolinones
thymic epithelial hyperplasia in females at 4000 ppm (21153), as compared to controls (7/49). However, for this 2-year (lifetime) rat study with imazamethabenz-methyl, the thymic hyperplasia did not progress to a neoplastic lesion nor has such a progression been described in the literature. Therefore, for this reason, this hyperplastic change in females at 4000 ppm appears to have only an equivocal toxicologic significance. In conclusion, based on the results from this study, the NOEL for systemic toxicity was 250 ppm (equivalent to approximately 12.5 mg/kg body weight/day). For imazamethabenz-methyl, the increased incidence of hepatocellular hypertrophy in the liver of male rats at both 5000 and 10,000 ppm, which was noted at termination of the 90-day dietary toxicity study, was not observed at either the 12-month interim sacrifice or the terminal sacrifice of the 2-year chronic rat study at dietary concentrations up to and including 4000 ppm. The absence of this microscopic finding indicates that adaptation probably occurred in the liver following prolonged exposure. In conclusion, the collective results from the oncogenicity studies in both the rat and the mouse indicate a lack of oncogenic!carcinogenic potential for all six imidazolinone herbicides. 74.2.2 ABSORPTION, DISTRIBUTION, METABOLISM, AND EXCRETION
In the rat metabolism studies with five of the six imidazolinone herbicides, only minimal metabolism is demonstrated. Following single oral gavage doses of imazapyr, imazapic, imazethapyr, imazamox, or imazaquin, these imidazolinone herbicides are rapidly absorbed and excreted, as evidenced by the presence of greater than 70% of unchanged parent compound in the urine within 24-48 h. Although imazamethabenzmethyl is also rapidly absorbed in the rat, greater than 60% becomes metabolized via hydrolysis of the ester to imazamethabenz acid, which is rapidly excreted in the urine within 24 h. The presence of a significant amount of unchanged parent or metabolite in the urine within 24-48 h, indicates rapid absorption of the imidazolinone herbicides from the gastrointestinal tract following a single oral gavage dose, as well as a low potential for bioaccumulation of parent compound or acid metabolite in mammalian tissues. 74.2.3 EFFECTS ON ORGANS AND TISSUES
The only consistent treatment-related effects that were observed in the extensive toxicological profile of the imidazolinone herbicides were slight-to-moderate skeletal myopathy and/or slight anemia in dogs, occurring in the I-year dietary toxicity studies with three structurally similar imidazolinones (imazapic, imazaquin, and imazethapyr). Specifically, both of these treatment-related effects were seen in male and female dogs treated with imazapic. Slight anemia at 20,000 ppm (mid-concentration) and 40,000 ppm (highest concentration
tested) was noted by decreases in hematological parameters and by histopathological examination, whereas slight-tomoderate skeletal myopathy at 40,000 ppm was determined by clinical chemistry evaluation and histopathological examination. Similarly, both of these treatment-related effects were seen in male and female dogs treated with imazaquin. Slight anemia at 5000 ppm (highest concentration tested) was noted by decreases in hematological parameters and by histopathological examination, whereas slight skeletal myopathy at 5000 ppm was determined by clinical chemistry evaluation and histopathological examination. In contrast, only slight anemia was obscrved by hematological parametcrs in fcmale dogs with imazethapyr at both 5000 ppm and I 0,000 ppm (highest concentration tested). There was no evidence of potential carcinogenicity in gross necropsy observations or in microscopic examinations of the full battery of tissues in the rat or mouse for any of the imidazolinone herbicides. 74.2.4 EFFECTS ON REPRODUCTION
Results from the reproductive and developmental toxicity studies indicate that the six imidazolinone herbicides are not reproductive toxicants, developmental toxicants, or teratogens (Table 74.2). Specifically, in the multigeneration reproductive toxicity studies conducted in rats, the imidazolinone herbicides did not affect reproductive performance, nor was there evidence of any significant prenatal or postnatal effects. All six reproduction studies support reproductive NOELs at the highest concentrations tested, namely, 20,000 ppm (imazapic and imazamox), 10,000 ppm (imazapyr, imazethapyr, and imazaquin), or 4000 ppm (imazamethabenz-methyl). These highest dietary concentrations are equivalent to approximate daily intake values of 1600 mg/kg body weight/day (imazapic), 1640 mg/kg body weight/day (imazamox), and 800 mg/kg body weight/day for males and 980 mg/kg body weight/day for females (imazapyr), as calculated from food consumption data. The preceding doses are equivalent to approximately 800 mg/kg body weight/day (imazethapyr and imazaquin) or 320 mg/kg body weight/day (imazamethabenz-methyl). Further, the teratology studies, which evaluated potential developmental toxicity of the six imidazolinone herbicides in rabbits and rats, revealed no evidence of developmental toxicity or teratogenic effects for fetuses of either species. All six imidazolinones, as tested in rabbits and rats, showed developmental NOELs equal to or higher than the maternal NOELINOAELs. For the rabbit teratology studies, the developmental NOELs were 1000 mg/kg body weight/day, the highest dose tested (HDT) for imazethapyr; 900 mg/kg body weight/day (HDT) for imazamox; 700 mg/kg body weight/day (HDT) for imazapic; 500 mg/kg body weight/day for imazamethabenz-methyl, 500 mg/kg body weight/day (HDT) for imazaquin; and 400 mg/kg body weight/day (HDT) for imazapyr. For these rabbit studies, the maternal NOELs were either the same dose level as the developmental no-effect level (e.g., 500 mg/kg
74.3 Toxicity to Humans body weight/day for imazamethabenz-methyl; 400 mg/kg body weight/day for imazapyr) or lower dose levels of 500 mg/kg body weight/day for imazapic, 300 mg/kg body weight/day for imazethapyr and imazamox, or 250 mg/kg body weight/day for imazaquin. Importantly, for all six imidazolinone herbicides, no treatment-related teratogenic effects were observed in the rabbit fetuses. For the rat teratology studies, the developmental NOELs were 1000 mg/kg body weight/day, the HDT for imazapyr, imazamethabenz-methyl, imazapic, and imazamox; 500 mg/kg body weight/day (imazaquin); and 375 mg/kg body weight/day (imazethapyr). For these rat studies, the maternal NOEL! NOAELs were either the same dose level as the developmental no effect level (e.g., 1000 mg/kg body weight/day for imazamethabenz-methyl and imazapic; 500 mg/kg body weight/day for imazaquin) or lower dose levels of 500 mg/kg body weight/day for imazamox, 375 mg/kg body weight/day for imazethapyr or 300 mg/kg body weight/day for imazapyr. Importantly, for all six imidazolinone herbicides, no treatment-related teratogenic effects were observed in the rat fetuses. In conclusion, based on the results given previously for the multigeneration reproduction studies and the rabbit and rat teratology studies, the imidazolinone herbicides demonstrate a lack of reproductive toxicity and are neither selective developmental toxicants nor teratogens in either the rabbit or the rat.
74.2.5 PATHOLOGY The only consistent treatment-related effects that were observed microscopically in the extensive toxicological profile of the imidazolinone herbicides were slight anemia and slight-tomoderate skeletal myopathy in dogs, occurring in the I-year dietary toxicity studies with two structurally similar imidazolinones (imazapic and imazaquin). Specifically, both of these treatment-related effects were seen in male and female dogs treated with imazapic. Slight anemia was noted by histopathological examination (erythropoiesis in the bone marrow and spleen) at 20,000 ppm (mid-concentration) and 40,000 ppm (highest concentration tested). Slight-to-moderate skeletal myopathy was also diagnosed by histopathological examination (skeletal muscle degeneration) at 40,000 ppm (terminal sacrifice). Similarly, both of these treatment-related effects were seen in male and female dogs treated with imazaquin. Slight anemia (erythropoiesis in the bone marrow) was diagnosed at 5000 ppm (highest concentration tested). In addition, slight skeletal myopathy (skeletal muscle degeneration) was observed at 5000 ppm by histopathological examination. There was no evidence of potential carcinogenicity for any of the imidazolinone herbicides from evaluations of the gross necropsy observations or the microscopic findings of the full battery of tissues in the rat or mouse.
1651
74.2.6 GENOTOXICITY STUDIES As presented in Table 74.2, based on the battery of in vitro and in vivo assays, the imidazolinone herbicides show a lack of potential genotoxic activity. This series of tests comprised the genotoxicity testing data requirements for all three categories [i.e., gene mutations (Ames and mammalian cell), in vitro structural chromosomal aberrations, and in vivo abnormal cytogenetics such as detected in the mouse micronucleus assay using bone marrow cells].
74.3 TOXICITY TO HUMANS 74.3.1 USE EXPERIENCE There are no incident reports attributable to the active ingredients for workers involved in the manufacturing process of the imidazolinone herbicides or for workers involved in mixing/loading/applying the end-use products of the imidazolinone herbicides for crop or noncrop uses. Results from the acute dermal and oral toxicity data, as cited in Section 74.2, indicate that the imidazolinone herbicides do not pose any acute dermal or dietary risks. Furthermore, because of their relatively low toxicity to mammals, the imidazolinone herbicides demonstrated relatively high no observable effect levels for potential systemic toxicity in the long-term studies, as cited in Section 74.2. In the absence of genotoxic, carcinogenic, reproductive, or teratogenic effects, a safety factor of lOO (I 0 x for interspecies differences, IO x for intraspecies differences) is appropriate to calculate the acceptable daily intake (ADI) for chronic human exposure. Applying a low (lOO-fold) safety factor to these relatively high systemic NOELs from the long-term studies with the imidazolinone herbicides results in ADIs that are relatively high. These relatively high ADIs suggest that this class of compounds does not pose a concern for chronic dietary exposure to humans. 74.3.2 TREATMENT OF POISONING There are no known cases of accidental or deliberate poisonings in humans. Because of their low toxicity profile and rapid excretion rate (see preceding discussion), the development of any physiological antidotes for the imidazolinone herbicides appears unnecessary. For formulated products, it is advisable to consult the specific material safety data sheet (MSDS) for emergency and first-aid procedures.
REFERENCES Popp, J. A., and Cattley, R. C. (1991). Hepatobiliary system. In "Handbook of Toxicologic Pathology" (W. M. Haschek and C. G. Rousseaux, eds.), pp. 279-314. Academic Press, San Diego.
CHAPTER
75 Toxicology of Triazolopyrimidine Herbicides Thomas R. Hanley, Jr., and Richard Billington Dow AgroSciences, LLC
75.1 INTRODUCTION
75.2 CLORANSULAM-METHYL
The triazolopyrimidines are herbicides used for the preemergent and postemergent control of broadleaf weeds in a variety of crops. The general structure of this class is a substituted triazolopyrimidine connected to a substituted phenyl ring through a sulfonamide bridge as presented in Fig. 75.1. The substituents of the various members of this class are presented in Table 75.1. The mode of action is through inhibition of acetolactate synthase (ALS) in plants, though the mechanism appears to be different from that of sulfonylureas. Acetolactate synthase (EC 4.13.18), also known as acetohydroxyacid synthase, is a key enzyme in the synthesis of the branched-chain aliphatic amino acids leucine, isoleucine, and valine. Inhibition of this enzyme in plants results in cessation of cell growth and division, leading to the death of susceptible plants. However, this enzyme is lacking in humans and other animals, which accounts for the low mammalian toxicity of these materials. Extensive toxicological testing has been conducted with these compounds according to standard test guidelines as published by the U.S. Environmental Protection Agency (EPA), the Organization for Economic Cooperation and Development (OECD), the European Union (EU) and the Japanese Ministry of Agriculture, Forestry and Fisheries (JMAFF) to determine potential health effects. The results of these studies have been evaluated in conjunction with use rates and exposure and residue data to provide comprehensive risk assessments. In general, these materials have low acute toxicity, have low chronic toxicity, and have been negative in tests for mutagenicity. The kidneys and, to a lesser extent, the liver are the primary organs affected by repeated exposure, and in most cases the histologic changes represent adaptive responses. Absorption is rapid, as is excretion, with no evidence of accumulation, and these materials appear, for the most part, to be metabolically stable in mammals. The mammalian toxicity of these materials is reviewed in this chapter. Handbook of Pesticide Toxicology Volume 2. Agents
75.2.1 IDENTITY, PROPERTIES, AND USES Chemical Name The chemical name for cloransulam-methyl is N -(2-carboxymethyl-6-chloropheny1)-5-ethoxy-7-fiuoro(1,2,4 )triazolo( 1,5c )pyrimidine-2-sulfonamide. Structure
See Fig. 75.1 and Table 75.1.
Synonyms Cloransulam-methyl is also known as XR-565, or XDE-565, and is sold as FirstRate® herbicide in the United States, and as PACTO® and SUPRA® herbicides in South America. (All trade names used in this chapter are registered trademarks of Dow AgroSciences, LLC.) The Chemical Abstract Service (CAS) registry number is 147150-35-4. Physical and Chemical Properties The empirical formula for cloransulam-methyl is ClsH13ClFNsOsS, with a molecular weight of 429.8. It is a solid at room temperature, with a low vapor pressure (3 x 10- 16 mm Hg at 25°C). The water solubility is pH dependent, with values of 2.96 mg/l at pH 5, 184 mg/l at pH 7, and 3430 mgll at pH 9 (20°C). The log Kow is estimated at 3.7; the pKa is 4.81. Uses Cloransulam-methyl is used as a soil-applied or incorporated preemergence or postemergence broadleaf herbicide in soy beans at maximum label rates of 44 g per hectare soil applied and 18 g per hectare postemergence.
Figure 75.1
1653
Generic structure of the triazolopyrimidine herbicides. Copyright © ZOO 1 by Academic Press. An rights of reproduction in any fonn reserved.
1654
CHAPTER 75
Triazolopyrimidine Herbicides Table 75.1 Substituents of Triazolopyrimidine Sulfonamide Herbicides 2
RI
R2
R3
R!
RS
R6
H
(l,Sc) Clorasulam-methyl
N
C
H
C02CH3
F
N
C
H
C02CH3 Cl
Cl
Diclosulam
Cl
OCH2CH3
F
H
Florasulam
N
C
H
F
F
OCH3
H
F
Flumetsulam
C
N
H
F
F
H
CH3
Metosulam
C
N
CH3
Cl
Cl
OCH3
OCH3
(I,Sa)
75.2.2 TOXICITY TO LABORATORY ANIMALS Acute Exposure The acute toxicity of cloransulam-methyl was low. The acute oral LDso in the rat was greater than 5000 mg/kg in both males and females and the dermal LDso in the rabbit was greater than 2000 mg/kg. The 4-hr inhalation LCso in the rat was greater than 3.77 mg per liter of air, which was the highest attainable respirable aerosol concentration. Cloransulam-methyl produced no indications of dermal irritation in rabbits or sensitization in the guinea pig, and only slight transient eye irritation in the rabbit following acute exposure (EPA, 1997a, b). Repeated Exposure Cloransulam-methyl was evaluated in subacute and subchronic dietary studies in rats, mice, and dogs. The primary target organs identified in these studies were the kidneys (rat and mouse), the liver (rat, mouse, and dog), and thyroid (rat). In the Fischer 344 (F344) rat, dosages of 100-1000 mg/kg/ day for 2 weeks produced slight decreases in red blood cell parameters and urine specific gravity in males, and slightly increased cecal and liver weights in females at 1000 mg/kg/day. The no-ob served-effect level (NOEL) was 500 mg/kg/day. Dosages of 100-1000 mg/kg/day for 13 weeks produced treatment-related kidney changes consisting of very slight to moderate hypertrophy of collecting tubule epithelial cells and/or slight vacuolation of the renal proximal tubular epithelium consistent with fatty changes in all dosage groups. Decreased body weight gain and feed consumption, very slight hepatocellular vacuolation, and slight thyroid follicular hypertrophy were also seen at 500 and 1000 mg/kg/day (Haut et aI., 1991, 1992a; Stebbins and Haut, 1994). Dosages of 100, 500, or 1000 mg/kg/day fed to B6C3Fl mice for 2 weeks produced slight hepatocellular hypertrophy at 500 mg/kg/day and above in males, and at 1000 mg/kg/day in females. The NOEL was 100 mg/kg/day. Dosages ranging from 50 to 1000 mg/kg/day given for 13 weeks produced slight centrilobular and midzonal hepatocellular hypertrophy at 100 mg/kg/day and above in males, and at 500 mg/kg/day and above in females. Electron microscopy characterized the hypertrophy as an increase in rough endoplasmic reticulum (RER) with a decrease in cytoplasmic glycogen content. Kidney effects in mice
consisted of decreased vacuolation of the renal tubules, consistent with decreased cytoplasmic lipid, accompanied by lower kidney weights at 500 mg/kg/day and higher. The subchronic lowest observed effect level (LOEL) and NOEL values in mice were 100 and 50 mg/kg/day, respectively (Haut et aI., 1992b; Stebbins and Haut, 1993). Cloransulam-methyl, when fed to dogs for 2 weeks at dosages of 500 mg/kg/day or higher, produced hepatic inflammation, degeneration, and necrosis. No effects were seen at dosages of 200 mg/kg/day or lower. In a subchronic study, dogs exhibited a taste aversion to this material at dosages of 200 mg/kg/day and above, which resulted in a combination of impaired nutritional status and toxicity of the material. A dosage of 40 mg/kg/day resulted in lower body weight gains. Histologic examination did not identify a target organ, though a subsequent chronic study in dogs identified the liver as the primary target organ. Based on decreased body weights, a subchronic NOEL was not established in dogs (Stebbins et aI., 1996; Szabo et aI., 1992). In a 21-day repeated dermal application study in rabbits, cloransulam-methyl at dosages of 100, 500, or 1000 mg/kg/day produced slight anemia in female rabbits at the highest dosage. Male rabbits were unaffected at 1000 mg/kg/day and the NOEL in females was 500 mg/kg/day (Gilbert and Yano, 1995a). Chronic Toxicity and Carcinogenicity The chronic toxicity of cloransulam-methyl has been evaluated in rats, mice, and dogs. In a 2-year study, Fischer 344 rats were fed cloransulammethyl at dosages of 10-325 mg/kg/day. Body weight gain was decreased at the highest dosage. Treatment-related histologic effects were limited to the kidneys and thyroid. Hypertrophy of a popUlation of renal collecting duct epithelial cells identified as a-intercalated cells was reported in males and females fed 325 mg/kg/day. (A similar histologic change noted in rats, mice, and dogs following exposure to florasulam will be discussed later in this chapter.) Vacuolation of the proximal tubules (consistent with fatty changes) in males fed cloransulam-methyl at 325 mg/kg/day, and females fed 75 or 325 mg/kg/day, and an increase in the incidence of mineralization of the renal pelvis in males fed 75 or 325 mg/kg/day also were present. Thyroid changes were confined to the highdosage males (325 mg/kg/day) and consisted of hyperplasia and
75.2 Cloransulam-Methyl hypertrophy of follicular epithelium. The NOEL in this study was 10 mg/kg/day (Jeffries et aI., 1995a). B6C3Fl mice were fed diets containing cloransulam-methyl at dosages of 10-1000 mg/kg/day for 2 years. As was seen in subchronic studies, the liver was the primary target organ in mice, with effects also noted in the kidneys. Increases in liver weights in males at ~100 mg/kg/day and females at 1000 mg/kg/day, and centrilobular hypertrophy in males at ~ 100 mg/kg/day were the only treatment-related effects noted in the liver. Kidney weights were decreased in males at 1000 mg/kg/day and females at ~100 mg/kg/day. In the kidneys, depletion of the normal epithelial cytoplasmic vacuoles, and decreases in the incidence of renal mineralization and renal tubular degeneration were noted in males at ~ 100 mg/kg/day. All of these histologic changes were interpreted to be either incidental or adaptive-physiologic responses to the test material rather than adverse effects. The NOEL in mice following chronic exposure was 10 mg/kg/day (Jeffries et aI., 1995b). There was no evidence of a tumorigenic or carcinogenic response in either mice or rats following long-term exposure. In a I-year chronic toxicity study, beagle dogs were fed dosages of 5-50 mg/kg/day. The only treatment-related effects were in the liver and consisted of a slight-to-moderate increase in accumulation of pigment in Kuppfer cells and hepatocytes, and slight centrilobular and midzonal hepatocellular hypertrophy at ~1O mg/kg/day, with changes in hepatic-related serum chemistry parameters at 50 mg/kg/day (Szabo and Davis, 1994). The U.S. EPA considered 10 mg/kg/day the NOEL in this study (EPA, 1997a). Mutagenicity In a battery of genotoxicity tests, cloransulammethyl showed no evidence of mutagenic potential. These tests included a bacterial reverse mutation assay (Ames test), an in vitro cytogenetic assay in Chinese hamster ovary cells (CHOIHGPRT assay), an in vitro chromosomal aberration assay in rat lymphocytes, and an in vivo cytogenetic assay in mouse bone marrow cells (EPA, 1997a, b).
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effects on reproductive performance or neonatal survival were seen even at the high dosage of 500 mg/kg/day. In a developmental toxicity study in SD rats, gavage dosages of up to 1000 mg/kg/day (limit test) on gestation days 6-15 produced no maternal or developmental toxicity. In a developmental toxicity study in New Zealand White rabbits administered gavage dosages of 0, 30,100, or 300 mg/kg/day on gestation days 7-19, maternal weight gain and feed consumption were affected only at 300 mg/kg/day. No adverse embryonal or fetal effects were noted at any dose level (Vedula et aI., 1992; Zablotny et aI., 1993, 1994). Absorption, Distribution, Metabolism, and Excretion Metabolism studies were conducted with 14C-radiolabeled cloransulam-methyl in the F344 rat using dose levels of 5 or 1000 mg/kg. At 5 mg/kg, over 90% of either a single dose or repeated (15 days) doses was absorbed. At 1000 mg/kg, only 28-30% of a single dose was absorbed. Urinary elimination was rapid in both cases with half-lives of approximately 6-9 hr. A higher percentage of the 5-mg/kg dose was excreted in the urine by females (68-80%) than by males (4050%) and these sex-dependent differences in disposition of the 5-mg/kg dose were attributed to more efficient elimination of unchanged cloransulam-methyl in the female versus male kidney. Analyses of urine and fecal extracts indicated that parent cloransulam-methyl accounted for the majority of the excreted radiolabled material. The only metabolite present at amounts greater than 5% was identified as the 4-0H phenyl derivative of cloransulam-methyl. Other minor metabolites included a hydroxylation of the pyrimidine ring, though the position of hydroxylation was not identified, and an N-acetyl cysteine conjugate of the parent material. Due to rapid elimination, cloransulam-methyl has little potential to accumulate upon repeated administration (Domoradzki et aI., 1995; Nolan et aI., 1995). 75.2.3 TOXICITY TO HUMANS
Neurotoxicity The neurotoxic potential of cloransulammethyl was evaluated in specialized studies. No neurotoxicologic effects were observed in rats following acute gavage exposure of up to 2000 mg/kg (highest dose tested). A complete battery of neurologic tests including functional observations (handheld and open field observations, grip strength, and landing foot splay), motor activity, and detailed neurohistopathology was conducted following 13-week exposure via the diet to dosages up to 1000 mg/kg/day. No treatment-related neurotoxic effects were observed in any of these measures (Shankar et aI., 1993; Spencer et aI., 1995). Reproductive Toxicity Cloransulam-methyl had no effect on reproduction or fetal development. In a multigeneration reproduction study in Sprague-Dawley (SD) rats, dosages of 100 mg/kg/day and above produced kidney and thyroid effects in the adults consistent with effects seen in subchronic and chronic studies. The NOEL for parental animals was 10 mg/kg/day. No
No data are available on accidental human exposures. However, the risk to humans from exposure to cloransu1am-methyl following normal use patterns is low. No detectable residues were found either in soy beans or, in most cases, in soybean forage or hay at a limit of detection of 0.005 ppm, and accumulation is unlikely based on plant and animal data. Tolerance levels of 0.02 ppm in soybean, 0.1 ppm for soybean forage, and 0.2 ppm for soybean hay have been established (EPA, 1997a). Using conservative estimates which assume 100% of crops contain the tolerance limit, and a reference dose (RID) of 0.10 mg/kg/day (based on the NOEL from the chronic dog study), the calculated maximum potential average daily dose from all sources indicate use of less than 0.22% of the RID in the subgroup with the highest aggregate exposure (nonnursing infants). The EPA estimated the margin of exposure (MOE) for occupational exposure to cloransulam-methyl to be between 2500 and 14,000, based on the use of a NOEL of 10 mg/kg/day (EPA, 1997a).
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75.3 DICLOSULAM
Acute Exposure The acute toxicity of diclosulam was low. The acute oral LDso in the rat was greater than 5000 mg/kg, the dermal LDso in the rabbit was greater than 2000 mg/kg, and the 4-hr inhalation LCso in the rat was greater than 5.04 mg/! of air. Diclosulam produced no indications of dermal irritation in rabbits or sensitization in the guinea pig, and only very slight transient eye irritation in the rabbit following acute exposure (EPA,1998).
above, secondary to slightly lower feed consumption in these animals. Slight decreases in red blood cell parameters were noted at 100 mg/kg/day and above. The NOEL from this study was 50 mg/kg/day (Stewart et aI., 1992a; Szabo and Davis, 1993a). Dietary exposure of B6C3Fl mice to dosages of 100-1000 mg/kg/day for 2 weeks resulted in slightly decreased kidney weights in both males and females at the high dosage (1000 mg/kg/day), and slightly decreased hepatocellular vacuolation (consistent with decreased glycogen content) in females at 500 mg/kg/day and above. The NOEL was 500 mg/kg/day in males and 100 mg/kg/day in females. Dosages of 1001000 mg/kg/day were given to B6C3Fl mice for l3 weeks. Significant body weight effects were seen in males at 1000 mg/kg/day and in females at 500 and 1000 mg/kg/day, and slight-to-moderate hepatocellular hypertrophy was the primary histopathologic change noted at 500 and 1000 mg/kg/day in males and females, respectively. Kidney weights were lower in males and females at 500 mg/kg/day and above, but there were no correlative changes in clinical chemistry or histopathologic parameters. The NOEL for subchronic exposure in the mouse was 100 mg/kg/day (Grandjean and Szabo, 1993; Stewart et aI., 1992b). Beagle dogs were given diclosulam at dosages of 50-500 mg/kg/day for 2 weeks. Dosages of 250 mg/kg/day and above were unpalatable and resulted in severely decreased weight gain or actual weight loss, degenerative changes in the kidneys, and hepatocellular necrosis. A dosage of 50 mg/kg/day produced microfocal hepatocellular necrosis in males, but no effects in females. In a subchronic study, dogs were given dosages of 0, 5,25, or 100 mg/kg/day for l3 weeks. Slight, diffuse centrilobular hepatocellular hypertrophy was observed at 25 mg/kg/day. Higher dosages proved to be unpalatable, with secondary toxicity associated with inanition superimposed on the effects of diclosulam on the liver. The subchronic NOEL for the dog was 5 mg/kg/day (Swaim and Szabo, 1992; Szabo and Rachunek, 1992). In a 21-day repeated dermal application study in rabbits, no dermal or systemic effects were seen at 1000 mg/kg/day, the highest dosage tested (Redmond and Kociba, 1996).
Repeated Exposure The primary target organs identified in dietary toxicity studies were the kidneys (rat) and the liver (rat, mouse, and dog). In the F344 rat, dosages of 500 and 1000 mg/kg/day for 2 weeks resulted in increased liver weights and enlarged ceca in males with no histopathologic changes. The NOEL was 100 mg/kg/day in males and 1000 mg/kg/day in females. Rats were given dietary dosages of 50-1000 mg/kg/day for 13 weeks. At 500 and 1000 mg/kg/day, body weights were decreased, and kidney and liver weights were increased. Very slight-tomoderate treatment-related hepatocellular hypertrophy was observed in males at 100 mg/kg/day and above, and in females at 1000 mg/kg/day. Kidney changes characterized as slightly to moderately decreased intracellular protein in the proximal tubule epithelium were seen in male rats at 500 mg/kg/day and
Chronic Toxicity and Carcinogenicity Chronic studies in rodents with diclosulam produced adaptive changes in the kidney as the primary effect. In a 2-year study in Fischer 344 rats, decreased body weight and weight gain were observed at 400 mg/kg/day, along with changes in hematology, clinical chemistry, and urinalysis parameters associated with the decreased body weight. Histologically, a slight alteration in tubular morphology, mostly within the cortico medullary junction, was observed in the kidneys at 100 and 400 mg/kg/day. This subtle change in the cytologic character and architecture was considered a slight alteration of the normal physiologic state, rather than a pathologic effect indicative of a toxic injury. No effects were noted in rats at 5 mg/kg/day (Minnema, 1996a). Chronic dietary exposure of B6C3Fl mice to dosages of 50-500 mg diclosulam/kg/day for two years produced no
75.3.1 IDENTITY, PROPERTIES, AND USES
Chemical Name The chemical name for diclosulam is N -(2,6-dichlorophenyl)-5-ethoxy-7 -fiuoro-(1,2,4)triazolo(1 ,5c) pyrimidine-2-sulfonamide. Structure
See Fig. 75.1 and Table 75.1.
Synonyms Diclosulam is also known as XR-564 or XDE564 and is sold primarily in the United States and South America under the trade names Strongarm®, SPIDER®, and CROSSER® herbicides. The CAS registry number is 14570121-9. Physical and Chemical Properties The empirical formula of diclosulam is C13HlOClzFNs03S, with a molecular weight of 406.2. Diclosulam is a solid, with a burnt vanilla odor, though the vaporpressure is low (5 x 1O- 1s mm Hg at 25°C). The water solubility is pH dependent and increases with increasing pH, from 117 mg/l at pH 5 to 4290 mg/l at pH 9, and the log Kow values are -0.047 at pH 7 and -0.448 at pH 9. Uses Diclosulam is a soil-applied, preplanting broadleaf herbicide for use in soy beans and peanuts at maximum label rates 35 and 26 g per hectare, respectively. 75.3.2 TOXICITY TO LABORATORY
ANIMALS
75.4 Florasulam
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treatment-related effects on survival, body weights, feed consumption, or clinical observations. The primary histologic change noted in male mice was a reduced vacuolation of the kidney tubular epithelium at all dose levels at the interim and terminal sacrifices, which correlated with decreased absolute and relative kidney weights. In female mice, minimal focal dilation with hyperplasia of the lining epithelium of renal cortical tubules was seen at 100 mg/kg/day and above. In males, this same focal dilation was seen spontaneously across all groups, including controls. There appeared to be no biologic or toxicologic significance to these microscopic changes. The no-observed-adverse-effect level (NOAEL) in mice following chronic exposure was 50 mg/kg/day (Minnema, 1996b). There was no evidence of tumorigenicity or carcinogenicity in either mice or rats. In beagle dogs fed dosages of 2-25 mg diclosulam/kg/day for 1 year, only slight elevations in mean alkaline phosphatase and creatinine levels in dogs given 25 mg/kg/day were observed. These slight elevations, however, were considered reflective of the normal variability in this species, and 25 mg/kg/day was the NOEL (Walker, 1996).
by both males and females. At 500 mg/kg, only 15-20% of a single dose was absorbed. Urinary elimination was rapid in both cases with half-lives of approximately 7-12 hr. A higher percentage of the 5-mg/kg dose was excreted in the urine by females (62-68%) than by males (39--43%), with the remainder of the absorbed dose eliminated in the feces. At 500 mg/kg, the majority of the administered dose (82-85%) was found in the feces, with only 6-12% eliminated via the urine in both males and females. Within 72 hr, less than 3% of the dose remained in the tissues and carcass in all dose groups. The primary urinary and fecal excretion products were identified as unchanged diclosu1am and an OH-phenyl oxidation product. In addition, the N-acetyl cysteine conjugate of diclosulam, and the S-oxide of the N-acetyl cysteine conjugate were excreted in the urine of males and females, whereas the sulfate and/or glucuronide conjugate of the OH-phenyl metabolite was seen only in the urine of male rats. Based on rapid elimination, diclosulam has little potential to accumulate upon repeated administration (Stewart et aI., 1996).
Mutagenicity In a battery of genotoxicity tests, diclosulam showed no evidence of mutagenicity. These tests included a bacterial reverse mutation assay (Ames test), an in vitro cytogenetic assay in Chinese hamster ovary cells (CHOIHGPRT assay), an in vitro chromosomal aberration assay in rat lymphocytes, and an in vivo cytogenetic assay in mouse bone marrow (EPA, 1998).
Risk assessments using conservative assumption indicate high margins of safety with diclosulam. Residue studies indicated no detectable residues at a limit of detection of 0.003 ppm, and no likelihood for accumulation. A tolerance level of 0.02 ppm, based on a limit of quantitation of 0.01 ppm, and a reference dose of 0.05 mg/kg/day based on the lowest NOEL (5 mg/kg/day from the chronic rat study) have been proposed (EPA, 1998). Calculation of a maximum potential average daily dose assuming 100% of proposed crops with residues equal to the tolerance level indicates theoretical exposure to only 0.1 % of the RID in the population with the highest potential exposure (nonnursing infants under 1 year old). The MOE for occupational exposure to diclosulam, calculated using exposure estimates from the U.S. EPA Pesticide Handlers Exposure Database (PHED), is estimated to be greater than 1000 based on the NOEL from the chronic dog study and assuming 100% absorption.
Neurotoxicity No neurotoxicologic effects were noted in rats following acute gavage exposure to up to 2000 mg/kg (highest dose tested) or in a complete battery of neurologic tests including detailed histopathologic examination following 1 year of exposure via the diet to dosages up to 400 mg/kg/day (Mattsson et al., 1996; Minnema, 1996c). Reproductive Toxicity Treatment with diclosulam had no effect on reproduction or fetal development. In a multigeneration reproduction study in Sprague-Dawley rats at dietary dosages up to 1000 mg/kg/day, no indications of parental or reproductive toxicity were seen. Gavage dosages of up to 1000 mg/kg/day to pregnant Sprague-Dawley rats on gestation days 6-15 produced no maternal or developmental toxicity. In New Zealand White rabbits, no developmental effects were noted even at gavage dosages up to 650 mg/kg/day on gestation days 7-19, which severely affected maternal feed consumption and weight gain. The maternal NOEL in rabbits was 65 mg/kg/day, whereas the developmental NOEL was 650 mg/kg/day (Morseth, 1994; Zablotny, 1996; Zablotny et aI., 1996). Absorption, Distribution, Metabolism, and Excretion Metabolism studies conducted with 14C-diclosulam in the F344 rat using dose levels of 5 or 500 mg/kg revealed that approximately 80% of a single or repeated (15 days) low doses was absorbed
75.3.3 TOXICITY TO HUMANS
75.4 FLORASULAM 75.4.1 IDENTITY, PROPERTIES, AND USES Chemical Name The chemical name for florasulam is N -(2,6-difluoropheny~- 5-methoxy-8-fluoro-(l ,2,4)triazolo(l ,5c) pyrimidine-2-sulfonamide. Structure
See Fig. 75.1 and Table 75.1.
Synonyms Florasulam is also known as XR-570, XDE-570, or DE-570 and is sold either alone or in combination under the registered trade names PRIMUS®, DERBY®, KANTOR®, and MUSTANG® herbicides. The CAS registry number is 14570123-1.
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Physical and Chemical Properties Florasulam is a light colored solid with an empirical formula of C12H8F3NS03S and a molecular weight of 359.3. It has a low vapor pressure (7.5 x 10- 8 mm Hg at 25°C) and decomposes at 193.5-230.5°C. The solubility increases with increasing pH, ranging from 84 mg/l at pH 5 to 9400 mg/l at pH 9. Florasulam is highly soluble in acetone (123 g/l) and acetonitrile (72 g/l), but substantially less soluble in octanol (0.18 g/l) and xylene (0.23 g/l). It has a pKa of 4.54 and log Kow values ranging from 1.00 at pH 4 to -2.06 at pH 10.
Uses Florasulam is a highly effective postemergence broadleaf herbicide for use in cereals, grassland, and turf. Maximum label use rate for the various crops range from 5 to 109 per hectare. 75.4.2 TOXICITY TO LABORATORY ANIMALS Acute Exposure Florasulam was essentially nonhazardous by the oral, dermal, and inhalation routes, was nonirritating to skin and eyes, and did not induce delayed contact hypersensitivity in either a modified Buehler test or a Magnusson and Kligman maximization study. The oral LDso was greater than 6000 mg/kg in the rat and 5000 mg/kg/day in the mouse, the dermal LDso in the rabbit was greater than 2000 mg/kg, and the 4-hr inhalation LCso in the rat exceeded 5 mg/l (Brooks, 1997; Clements and Cieszlak, 1995; Gilbert, 1995a, b, c, d; Gilbert and Yano, 1995b; Johnson, 1996). Repeated Exposure In dietary studies of 2- to 13-week duration, the kidney was identified as a target organ in rats, mice, and dogs, whereas the liver was a target organ in dogs. In F344 rats, subacute exposure to dosages of 500 mg/kg/day and above was associated with karyomegaly and anisokaryocytosis in proximal tubular epithelial cells in males and females, and tubular degeneration with regeneration in females. Individual proximal tubular cell necrosis was also seen in both sexes at 1000 mg/kg/day. The NOEL was 100 mg/kg/day. Subchronic studies were conducted in F344 and Sprague-Dawley rats at dosages up to 1000 mg/kg/day. Dosages of 500 or 1000 mg/kg/day produced necrosis with regeneration in descending proximal tubules and a marginally increased incidence of degeneration with regeneration of renal tubules in females. Papillary mineralization (tubular debris) and papillary necrosis were reported at the highest dosages (::0:800 mg/kg/day). Other highdosage effects included acidic urine (males only), increased kidney weight, perineal soiling, reduced body weight gain and feed consumption (due, at least in part, to reduced palatability of the diet), and reduced red blood cell indices. With the exception of mineralized debris in renal papillae and degeneration and regeneration of cortical tubules, all effects partially or completely resolved by the end of a 4-week recovery period. The subchronic NOEL in rats was 100 m6/kg/day (Liberacki et aI., 1996; Redmond and Johnson, 1996a; Szabo and Davis, 1993b).
In B6C3Fl mice, subacute exposure to dosages up to 1000 mg/kg/day was without effect (Szabo and Davis, 1992). The only response to subchronic exposure to dosages up to 1000 mg/kg/day was hypertrophy of renal collecting duct epithelial cells at 500 mg/kg/day and above. The subchronic NOEL in mice was 100 mg/kg/day (Redmond and Johnson, 1996b). In Beagle dogs, subacute exposure to a nominal dosage of 450 mg/kg/day was associated with reduced body weight gain and reduced feed consumption (due, at least in part, to reduced palatability of the diet). Hepatic changes characterized by increased liver weight and bile duct hyperplasia in males and females, and bile stasis and hepatocellular necrosis in males were observed in this group. At 150 mg/kg/day, effects were limited to increased liver weight and bile duct hyperplasia. Serum alkaline phosphatase (AP) activity, probably of hepatic origin, was elevated at all dosages, including the low dosage of 50 mg/kg/day. The increase in serum AP at the low dosage was without histopathological correlate. Therefore, 50 mg/kg/day was considered a subacute NOAEL. Liver effects were not exacerbated by an extended treatment period, but renal hypertrophy (not seen after 4 weeks of exposure) similar to that reported in rats and mice was evident in dogs after subchronic exposure to 50 mg/kg/day (Sullivan and Cronin-Singleton, 1995; Sullivan and Singleton, 1995). Repeated dermal exposure to dosages up to 1000 mg/kg/day for 4 weeks produced only transient dermal irritation in rats during the last week of treatment, with no systemic effects (Scortichini and Kociba, 1997). Chronic Toxicity and Carcinogenicity In chronic dietary studies (1- to 2-year duration), hypertrophy of a population of renal collecting duct cells identified as a-intercalated cells remained the most sensitive morphological effect in all species. Hypertrophy was present at 50 mg/kg/day in dogs, 125 and 250 mg/kg/day and above in male and female rats, respectively, and 500 mg/kg/day and above in mice. In F344 rats, 2-year dietary exposure to dosages between 10 and 500 mg/kg/day identified the kidney as the only target organ. At the high dosage (250 mg/kg/day in females and 500 mg/kg/day in males), very slight to slight hypertrophy of renal collecting duct epithelial cells was evident after 1 year. After 2 years, this change had progressed to a moderate degree in some males. Other effects at this dosage included reduced body weight gain, reduced urinary pH, perineal soiling, reduced red blood cell indices, and renal changes similar to effects seen following subchronic exposure to high dose levels. At the next dosage (125 mg/kg/day in females and 250 mg/kg/day in males), very slight to slight hypertrophy of renal collecting duct epithelial cells was evident in males after I year, and in males and females after 2 years. Reduced body weight gain, renal papillary mineralization (males), decreases in spontaneous chronic renal disease, reduced urinary pH, and perineal soiling were also seen in these animals. No effects occurred at 10 mg/kg/day and there was no treatment-related tumorigenicity or carcinogenicity (John son et aI., 1997).
75.4 Florasulam In B6C3F1 mice, 2-year dietary exposure to dosages up to 1000 mg/kg/day identified the kidney as the only target organ with hypertrophy of intercalated cells, decreased renal epithelial cell cytoplasmic lipid-like microvacuoles, and a decreased incidence of spontaneous chronic renal disease at 1000 mg/kg/day. Reduced body weights accompanied by minor changes in serum cholesterol and triglycerides also were present at 1000 mg/kg/day. At 500 mg/kg/day, very slight hypertrophy of renal collecting duct epithelial cells occurred in most males and females along with decreases in cytoplasmic lipid-like microvacuoles and spontaneous chronic renal disease (females only) of renal tubules. No effects occurred at 50 mg/kg/day, and there was no treatment-related tumorigenic or carcinogenic response at any dosage (Quast et aI., 1997). In Beagle dogs, I-year dietary exposure to florasulam revealed kidneys, liver, and adrenal glands as target organs. In the subchronic study, renal hypertrophy and modest elevations of serum alkaline phosphatase and liver weight were the only treatment-related effects seen at 100 mg/kg/day. However, treatment beyond 13 weeks resulted in reduced food consumption and body weight gain in some animals, and significant elevations in serum enzyme activities associated with liver toxicity. The original high dosage of 100 mg/kg/day was therefore reduced to 50 mg/kg/day on Day 105. Thereafter, food consumption and body weight gain improved and red blood cell indices in females and serum transaminases in both sexes returned to normal. Serum alkaline phosphatase remained elevated to the end of the study. Slight hypertrophy of renal collecting duct epithelial cells and slight vacuolization of the zona reticularis and zona fasciculata of adrenal glands were detected histologically in this high dosage group. The fatty change in the adrenals of dogs represented a slight exacerbation of a spontaneous lesion, not associated with inflammation, necrosis, or clinical chemistry changes, and was considered of uncertain toxicological importance. No histopathological lesions were evident in the liver. The NOEL was 5 mg/kg/day (Stebbins and Haut, 1997). Histological and ultrastructural evaluation of the affected renal collecting duct cells characterized the hypertrophy as a mitochondrial proliferation of a-intercalated cells, which functionally are involved in acid-base regulation and contain high levels of H+ -ATPase and H+K+ -ATPase in the apical membrane (Brown et aI., 1988; Garg, 1991; Hamm and HeringSmith, 1993; Madsen and Tisher, 1986; Stokes, 1993; Verlander et aI., 1991). Hypertrophy of intercalated cells has been reported as a physiological response to several factors affecting acid-base homeostasis, including respiratory acidosis, metabolic acidosis, hypokalemia, and altered serum adrenal mineralocorticoid levels (Ahn et aI., 1996a, b; DeFronzo, 1980; Eiam-ong et aI., 1994; Hansen et aI., 1980; Madsen et aI., 1991; Tsuruoka and Schwartz, 1996a, b; Verlander et aI., 1994; Weiner and Wingo, 1997; Wingo and Cain, 1993). However, none of these factors was found to be adversely affected by florasulam (Weiner, 1997). The lack of any adverse sequelae associated with this change suggests that it is an adaptive rather than an adverse response to florasulam.
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Mutagenicity In a battery of genotoxicity tests, florasulam showed no evidence of mutagenic potential. These tests included an in vitro bacterial reverse mutation assay (Ames test), an in vitro cytogenetic assay in Chinese hamster ovary cells (CHO/HGPRT assay), an in vitro chromosomal aberration assay in rat lymphocytes, and an in vivo cytogenetic assay in mouse bone marrow cells (Lawlor, 1995; Lick et aI., 1995; Linscombe et aI., 1995a, b). Neurotoxicity Acute gavage and chronic (I-year dietary) neurotoxicity studies in Fischer 344 rats revealed only nonspecific findings. In both acute and chronic neurotoxicity studies with florasulam, perineal urine staining at the highest dosages was the only treatment-related effect. No other effects were seen following an extensive battery of neurologic tests and neurohistopathological examinations (Mattsson and McGuirk, 1997; Shankar and 10hnson, 1996). Reproductive Toxicity In developmental tOXICIty studies, there were no adverse effects on intrauterine development or prenatal survival in rats or rabbits administered gavage dosages as high as 600-750 mg/kg/day. Maternal effects on survival, feed consumption, and/or weight gains occurred at these high dosages. In SD rats, the embryo-fetal NOEL was 750 mg/kg/day, whereas the maternal NOEL was 250 mg/kg/day. In New Zealand White rabbits, the NOEL for both maternal and embryo-fetal effects was 500 mg/kg/day. In a two-generation dietary reproduction study in SD rats at dosages of 10-500 mg/kg/day, parental effects (weight gain, feed consumption, renal changes) were seen only at the highest dosage, with no effects on any reproductive parameter. Transient decreases in neonatal body weights, secondary to decreases in maternal feed consumption, were seen at 500 mg/kg/day. The parental NOEL was 100 mg/kg/day, whereas the NOEL for reproductive effects was 500 mg/kg/day (Liberacki and Camey, 1997; Liberacki et aI., 1997; Zablotny and Camey, 1997). Absorption, Distribution, Metabolism, and Excretion In metabolism studies in F344 rats, single oral doses of 10-500 mg I4C-florasulam per kilogram were readily and extensively absorbed (>90% of a lO-mg/kg dose within 24 hr) and rapidly eliminated (plasma tI/2 = 8-10 hr) primarily in the urine (>85% of administered dose). The feces contained small amounts of the administered radioactivity (5-17%) depending on dose. More than 75% of the I4C activity in urine was found to be unchanged florasulam. Two minor metabo1ites, identified as a free and a conjugated (sulfated) hydroxyphenyl derivative of florasulam, were found. Feces contained unchanged florasulam and the free hydroxyphenyl-derivative. There was no evidence of hydrolysis of the sulfonamide bridge based on the metabolites found in the urine and feces. The rapid elimination of florasulam from tissues indicated no potential to accumulate upon repeated administration (Dryzga et aI., 1996; Hansen, 1997). Absorption following in vivo dermal exposure of rats to a concentrated suspension formulation containing I4C-florasulam was minimal (mean <0.5% over 72 hr). Results
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obtained from in vitro studies were similar to those obtained from the in vivo study (Bounds, 1997; Perkins and Billington, 1998).
75.4.3 TOXICITY TO HUMANS Risk assessment calculations for the general population and for pesticide handlers indicate a low-risk estimate. Residue studies have indicated no detectable levels in cereal gains at the limit of quantitation. Maximum residue limits (MRLs) of 0.01 ppm in grains, and 0.05 ppm whole plants and straw based on the limit of detection, and an acceptable daily intake (AD!) of 0.05 mg/kg/day on the basis of a chronic NOEL of 5 mg/kg/day in dogs have been proposed. The theoretical dietary intake of florasulam from all routes has been estimated to account for <0.5% of the ADI even in infants, the most susceptible subpopulation. Margins of exposure of > 2000 have been calculated for pesticide handlers using the PHED and an NOEL of 5 mg/kg/day.
75.5 FLUMETSULAM 75.5.1 IDENTITY, PROPERTIES, AND USES Chemical Name The chemical name for flumetsulam is N -(2,6-difluoropheny1)-5-methy1-( 1,2,4 )triazolo( 1,5-a)pyrirnidine-2-sulfonamide. Structure
See Fig. 75.1 and Table 75.1.
Synonyms Flumetsulam (also known as XRD-498) is the generic name for this material which is sold globally as BROADSTRIKE@, PYTHON@, PRESIDE@, and SCORPION@ herbicides. The CAS registry number is 98967-40-9. Physical and Chemical Properties The empirical formula of flumetsulam is C12H9F2NS02S, with a molecular weight of 325.3. Flumetsulam is a light colored powder at room temperature, with a melting point of 252.9°C, a vapor pressure of 2.8 x 1O- 1S mm Hg at 25°C, a Kow of 1.62 at pH 3.44, and a pKa of 4.60. Flumetsulam is soluble in water at 5.65 g/l at pH 7 and 25°C, but solubility decreases with decreasing pH, and it is less soluble in organic solvents. Uses Flumetsulam is a broad-spectrum, season-long herbicide used in the control of broadleaf weeds in soybeans, corn, and other major crops. Flumetsulam is applied as a soilincorporated preplanting, preemergence, or postemergence herbicide depending on the formulation, at a maximum use rate of 80 g per hectare.
75.5.2 TOXICITY TO LABORATORY ANIMALS Acute Exposure Flumetsulam has been examined in a complete battery of toxicologic studies to evaluate potential mammalian toxicity. Flumetsulam was relatively nontoxic following acute exposure, with an acute oral LDso greater than 5000 mg/kg, a dermal LDso greater than 2000 mg/kg, an acute 4-hr inhalation LCso above the highest attainable aerosol concentration of 1.2 mg/l of air, and only slight, transient eye irritation. No signs of dermal irritation were observed in rabbits following acute exposure, nor was there any evidence of dermal sensitization in guinea pigs (EPA, 1993). Repeated Exposure Subchronic toxicity studies in rats, mice, and dogs have indicated a low degree of toxicity following repeated oral exposure. In rats, dietary exposure to concentrations of up to 5% (approximately 6000 mg/kg/day) for 2-4 weeks identified the kidney as the primary target organ. Effects in the kidneys consisted of focal necrosis and inflammation of the papilIa(e), and tubular epithelial cell degeneration and regeneration, with secondary effects on urinalysis parameters at the highest dosage. The only effect reported at 1000 mg/kg/day was cecal enlargement in males. However, the ceca were normal histologically, and 1000 mg/kg/day was considered the NOAEL following 2-4 weeks of exposure. Rats fed diets containing flumetsulam at dosages of 250-2500 mg/kg/day for 13 weeks exhibited dose-dependent changes similar to those seen after 4 weeks of exposure. The NOAEL in rats following subchronic exposure was 25 mg/kg/day (Yano et aI., 1987, 1988; Zempel et aI., 1988). In B6C3Fl mice fed flumetsulam for 2 weeks, decreased kidney weights were reported in males at dietary concentrations of 1.5 and 3.0% and in females given 3.0%, which corresponded to dosages >3500 mg/kg/day. The NOEL in mice was 0.5% (approximately 1150-1365 mg/kg/day). B6C3Fl mice given dosages of 100-5000 mg/kg/day for 13 weeks displayed only a minimal increase in centrolobular-to-midzonal hepatocellular eosinophilia at the highest dosage, and decreased vacuolation of renal proximal tubular epithelium which is of doubtful toxicologic significance. The NOEL for mice was 1000 mg/kg/day, and 5000 mg/kg/day was considered a NOAEL (Bond et aI., 1987; Stott et aI., 1986). In both rats and mice, the increases in the size and weight of the cecum, observed only at high dosages and unassociated with any histologic changes, were considered adaptive in nature, most likely secondary to the effects of flumetsulam on the microenvironment within the cecum. Beagle dogs were fed flumetsulam at nominal dosages of 100-1000 mg/kg/day (males) or 1500 or 2500 mg/kg/day (females) for 2 weeks. In females, degeneration and regeneration of the renal tubular epithelial cells, and lymphocytic infiltration of hepatic sinusoids were reported. The NOEL in dogs was approximately 800 mg/kg/day (nominally 1000 mg/kg/day). In dogs given a dosage of 1000 mg/kg/day for 13 weeks, degenerative microscopic changes in the renal papilla, slight biliary stasis, and hepatocellular necrosis were observed. At 500
75.6 Metosulam mg/kg/day, slight renal papillary degeneration was noted microscopically in males, and increases in serum AP and globulin levels and decreased serum albumin levels, with no histopathologic correlates, were reported in both males and females. A dosage of 500 mg/kg/day was considered the NOAEL in females (Cosse et aI., 1989). Repeated dermal exposure to dosages of 2:: 100 mg/kg/day for 21 days produced very slight epidermal hyperplasia, but no indications of any systemic effects (Stebbins et aI., 1990). Chronic Toxicity and Carcinogenicity Flumetsulam was fed to F344 rats and B6C3Fl mice for 2 years at dosages of 100-1000 mg/kg/day. No treatment-related adverse effects were noted in mice (Bond et aI., 1991). In rats, atrophy of the renal papilla(e) with secondary hyperplasia and/or mineralization of the pelvic epithelium were noted in males given 1000 mg/kg/day, but not in females, and the NOEL was 500 mg/kg/day (Stott et aI., 1991). As was noted following subchronic exposure, cecal enlargement with no accompanying histopathologic changes was observed in both rats and mice following chronic exposure. There was no evidence of a tumorigenic or carcinogenic response in either rats or mice at dosages up to 1000 mg/kg/day. The dog appeared to be the most sensitive species to longterm exposure to f1umetsulam. Administration of dosages of 500 mg/kg/day in the diet for 1 year produced inflammatory and atrophic changes in the kidney, accompanied by calculi in females. At 100 mg/kg/day, only increased alkaline phosphatase activity and decreased serum albumin were reported, with no histologic changes in any organs. The chronic NOEL in dogs from this study was 20 mg/kg/day, whereas the NOAEL was 100 mg/kg/day (Yano et aI., 1991). Mutagenicity Flumetsulam was negative for mutagenic activity in an in vitro bacterial reverse mutation assay (Ames test), an in vitro cytogenetic assay in Chinese hamster ovary cells (CHOIHGPRT assay), an in vitro rat hepatocyte unscheduled DNA synthesis (UDS) assay, and an in vivo cytogenetic assay in mouse bone marrow cells (EPA, 1993). Reproductive Toxicity Flumetsulam did not affect development or reproduction in either rats or rabbits. No evidence of maternal toxicity, embryo-fetotoxicity or teratogenicity was observed in rats following exposure of pregnant females to 1000 mg/kg/day in the diet, though the weights of the ceca were increased, consistent with effects noted in previous dietary studies. No parental toxicity or alterations in reproductive performance occurred in rats given up to 1000 mg/kg/day over two generations. Gavage administration of f1umetsulam to pregnant rabbits at dosages of 500-700 mg/kg/day produced dose-related episodes of anorexia, with sequelae secondary to the altered nutritional status (deteriorated clinical condition, mortality, stomach erosions, etc.), but no embryo-fetotoxicity or teratogenicity accompanied these maternal effects. The maternal NOEL from this study was 100 mg/kg/day, whereas the NOEL for
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embryo-fetal development was 700 mg/kg/day (Hanley et aI., 1989; Zempel et aI., 1990; Zielke et aI., 1988). Absorption, Distribution, Metabolism, and Excretion Flumetsulam was rapidly, though incompletely, absorbed in mice and rats, with absorption half-lives of less than 1 hr following oral administration of doses of either 5 or 1000 mg/kg. Excretion was also rapid, with a urinary half-life of approximately 5-7 hr. Following oral administration of 14C-f1umetsulam, approximately 50-75% of the administered radiolabel was excreted in the urine primarily as unchanged parent material, though two minor «20% of urinary radiolabel) metabolites, believed to be conjugates of parent f1umetsulam, were found in the urine of mice. Approximately 20-35% of the dose was found in the feces, which represented apparently unabsorbed f1umetsulam (based on almost total elimination in the urine of an intravenous dose to rats) and tissue levels of 14C accounted for less than 1.5% of the administered dose. There were no differences in absorption, distribution or elimination based on sex, though slight differences were seen with increasing dose (Pottenger et aI., 1991; Timchalk et aI., 1988). 75.5.3 TOXICITY TO HUMANS Risk assessment indicates a low potential risk from normal use of f1umetsulam. Residue tolerances of 0.05 ppm have been set by the U.S. EPA for soybeans and for corn grain, fodder, and forage. However, no residues were detected in soybeans even after postemergent application of six times the maximum label rate. A reference dose of 1 mg/kg/day was established on the basis of a NOAEL of 100 mg/kg/day from a 1 year study in dogs. Dietary risk evaluation assuming 100% of crops are treated and residues are at the established tolerance levels indicates only 0.013% of the RID is used even by the highest exposed subgroup, nonnursing infants less than 1 year old (EPA, 1993). Based on the use rates, and the NOEL from the chronic dog study, the MOE for worker exposure is greater than 1000.
75.6 METOSULAM 75.6.1 IDENTITY, PROPERTIES, AND USES Chemical Name The chemical name for metosulam is N-(2,6-dichloro-3-methylphenyl)-5,7-dimethoxy-(1,2,4) triazolo( 1,5a)pyrimidine-2-sulfonamide. Structure
See Fig. 75.1 and Table 75.1.
Synonyms Metosulam is also known as methoxsulam, XRD511, XDE-511, and DE-511, and is sold either alone or in combination, under a variety of registered trade names including TACCO@, SANSAC@, ECLIPSE@, ATOL@, KOMPAL@, and SINAL@ herbicides. The CAS number is 139528-85-1.
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Physical and Chemical Properties Metosulam is a cream to tan colored powder with a Iow vapor pressure (7.5 x 10- IS mm Hg at 25°C). The empirical formula is CI4HI3CI2Ns04S, and the molecular weight is 418.3. The solubility of metosulam in water at 20°C and pH 7 is 700 mg/l. Given a pKa of 4.8, the solubility is pH dependent, with values of 100 mg/l at pH 5 and 5600 mg/l at pH 9 (at 20°C), and the log Kow is 2.12 at pH 5.
Uses Metosulam is a broad-spectrum, postemergence broadleaf herbicide intended for use in cereals, maize, pasture, alfalfa, and rice. Maximum label use rates for the various crops range from 5 to 30 g per hectare. 75.6.2 TOXICITY TO LABORATORY ANIMALS Acute Exposure The acute toxicity of metosulam was very low. The LDso values of metosulam given orally to Fischer 344 rats and CD-1 mice were both greater than 5000 mg/kg. No toxicity, including histopathological changes of eyes and kidneys, was evident in beagle dogs given one to five daily doses of 2000 mg/kg, by gelatine capsule. The dermal LDso in the rabbit was greater than 2000 mg/kg. The 4-hr inhalation LCso in the rat was greater than the highest attainable concentration of 1.9 mg/l of air. Metosulam, when applied to the intact skin of rabbits, produced no signs of irritation. Following instillation into rabbit eyes, slight conjunctival redness developed within 1 hr of treatment, but all treated eyes were normal within 1 day of treatment. There was no indication of contact sensitization in guinea pigs exposed to metosulam using either a Magnusson and Kligman maximization test or a modified Buehler topical patch method (UK Ministry of Agriculture, Fisheries, and Food, 1996). Repeated Exposure Repeated exposure toxicity studies were conducted with metosulam in rats, mice, dogs, rabbits, and monkeys. In rats, dietary administration of dosages of 5005000 mg/kg/day for 2 weeks to Sprague-Dawley rats resulted in lower body weights associated with unpalatability and the NOEL was 100 mg/kg/day. No significant effects were reported in Long Evans rats administered dosages of up to 2000 mg/kg/day for 2 weeks. Following subchronic exposure, the kidney was identified as the major target organ. In aB-week dietary study, the primary toxicological effects were renal alterations characterized as hypertrophy and nuclear pleomorphism of cells lining the proximal convoluted tubules at 100 mg/kg/day and above. After 4 weeks on control diet, hypertrophy of renal tubular cells had resolved, and nuclear pleomorphism was markedly decreased. The NOEL for subchronic dietary administration of metosulam in rats was 10 mg/kg/day. CD-1 mice administered metosulam in the diet at 100-5000 mg/kg/day for 2 weeks exhibited centrilobular hepatocellular necrosis and decreased vacuolation in the liver only at 2000 mg/kg/day and above. The NOEL was 1000 mg/kg/day. The only effect observed in a 13-week dietary study at dosages up
to 2000 mg/kg/day, was mild hepatocellular hypertrophy and the NOEL was 250 mg/kg/day. The kidney was identified as the most sensitive target organ in the dog (similar to the rat). In addition, ocular toxicity in the form of retinal damage unique to this species was also observed. Dietary administration of metosulam to Beagle dogs at dosages of 100-1000 mg/kg/day for 14 days proved unpalatable and produced dose-related decreases in feed consumption and body weights; retinal degeneration, necrosis, and detachment; and degeneration or focal necrosis of distal renal collecting tubules and collecting ducts. The NOEL in this study was 25 mg/kg/day. Metosulam was fed to dogs at dosages of 5, 25, and 50 mg/kg/day for 13 weeks. Clinical signs of blindness occurred as early as 6 weeks in all dogs administered 50 mg/kg/day. Microscopic examination of the eyes from these dogs showed retinal degeneration with detachment. Choroidal structures (tapetum lucidum, pigmented epithelium, and choroidal blood vessels) and other ocular structures were normal. Ocular tissues from dogs administered 5 mg/kg/day were normal. In the kidneys, very slight to moderate degeneration of the distal convoluted tubules and collecting ducts of dogs administered 25 mg/kg/day and above was reported, and the NOEL was 5 mg/kg/day. Male and female Cynomolgus monkeys exposed to oral dosages of 0 or 100 mg/kg/day for 6 weeks showed no renal or ocular toxicity after detailed examination which included an extensive histopathologic evaluation. Repeated dermal exposure of New Zealand White rabbits to dosages of up to 1000 mg/kg/day for 21 days produced no signs of dermal irritation or systemic effects (UK Ministry of Agriculture, Fisheries, and Food, 1996). Chronic Toxicity and Carcinogenicity Following chronic (2-yr) exposure in Sprague-Dawley rats at dosages of 5-100 mg/kg/day, the primary effects were confined to the kidneys, consistent with the findings following subchronic exposure, and the effects were more severe in male than in female rats. At 100 mg/kg/day, nuclear pleomorphism and hyperplasia of cells of the proximal tubules as well as basophilic adenomas and adenocarcinomas of the renal cortex were observed. At 30 mg/kglday, nuclear pleomorphism of proximal tubular cells was present and only a single renal cortical adenocarcinoma, which was within the historical control incidence for this tumor (Charles River Breeding Laboratories, 1987), was observed in this group. Short-term exposure studies demonstrated the presence of mitotic figures and nuclear pleomorphism in the renal cortex of male rats following as little as 1 week of dietary exposure to 100 mg metosulam/kg/day. Increased mitotic activity measured by BrdU incorporation correlated with the renal tubular epithelial changes noted histologically (UK Ministry of Agriculture, Fisheries, and Food, 1996). This suggested a nongenotoxic mechanism of repeated injury as described by Dietrich and Swenberg (1991) as the probable origin of the renal tumors in the chronic study with metosulam. In CD-l mice fed dose levels of metosulam of up to 1000 mg/kg/day for 18 months, there was no evidence of any increase
75.6 Metosulam
in tumor incidence and no effects were noted in any other parameter. Metosulam administered to Beagle dogs at dosages of 3-37.5 mg/kg/day for 12 months produced effects in the eyes and kidneys consistent with the findings of the subchronic study. At 37.5 mg/kg/day, variable retinal degeneration with detachment, beginning with diminished or absent pupillary light reflex, increased tapetal reflectivity, and progressive retinal deterioration, were observed. Degenerative lesions of the distal convoluted tubules and collecting ducts were also observed at 37.5 mg/kg/day. No effects were observed at lower levels, and the NOEL was 10 mg/kg/day (UK Ministry of Agriculture, Fisheries, and Food, 1996). The sensitivity of the dog eye to metosulam appears to be unique to this species. The pathology involved the loss of the photoreceptor layer and its nuclei, together with a collapse of the outer and inner plexiform layers and inner nuclear layer. No retinopathy was associated with the pigmented epithelial layer or in the tapetal cells. The pathologic changes were not consistent with either inherited retinal degeneration or nutritional deficiencies as an etiology. It is important to note that retinal pathologies were not detected with metosulam in any other species. Dosages of 300 mg/kg/day for 12 days in rabbits, and up to 1000 and 2000 mg/kg/day for 13 weeks in rats and mice, respectively, were not associated with any retinal changes. Exposure of mice to 1000 mg/kg/day for 18 months or rats to 100 mg/kg/day for 2 years likewise induced no retinal pathology. Significantly, dosages of 100 mg/kg/day for 6 weeks in the nonhuman primate, the species most closely resembling human ocular anatomy and physiology, also produced no evidence of retinal toxicity. Pharmacokinetic studies using radiolabeled metosulam indicated metosulam localized over the outer layer of the retina in the dog, but no selective localization was detected in the rat or mouse (see below) (UK Ministry of Agriculture, Fisheries, and Food, 1996). Mutagenicity A battery of mutagenicity tests which included an in vitro bacterial assay (Ames test), in vitro assays using mammalian cells (CHOIHGPRT, RLCAT, and UDS assay), and an in vivo mouse bone marrow micronucleus test were all negative (UK Ministry of Agriculture, Fisheries, and Food, 1996). Reproductive Toxicity Metosulam did not produce any adverse reproductive or developmental effects when tested in rats and rabbits. There were no effects on maternal or developmental parameters in a conventional teratogenicity study in the SD rat at dietary levels up to 1000 mg/kg/day. In New Zealand White rabbits, maternal effects were noted at oral gavage dosages of 100 or 300 mg/kg/day and the maternal NOEL in rabbits was 30 mg/kg/day, but there was no indication of developmental effects at 300 mg/kg/day. In a two-generation reproduction study in Sprague-Dawley rats at dosages of 5-100 mg/kg/day, renal toxicity was observed among the parental rats at 100 mg/kg/day consistent with the effects noted
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following chronic exposure, but reproductive performance was unaffected. The NOEL for parental toxicity from this study was 30 mg/kg/day, whereas the NOEL for reproductive effects was 100 mg/kg/day (UK Ministry of Agriculture, Fisheries, and Food, 1996). Absorption, Distribution, Metabolism, and Excretion The metabolic fate of 14C-metosulam in rats, mice, and dogs following single or multiple oral administrations was evaluated. 14C-Metosulam was absorbed rapidly (t1/ 2 < 1 hr) in all three species, though the extent of absorption was significantly higher in the rat (>70%) than in the dog and mouse ('"'-20%). The rate of 14C elimination in mice and rats was comparable (fI/2 = 5460 hr), whereas the elimination rate in dogs was slightly slower (t1/2 = 73 hr). In all three species, 14C-metosulam and metabolites were excreted in the urine. HPLC analysis of urine samples revealed extensive metabolism in both mice and rats, but much less pronounced metabolism in dogs. Analysis of 14C activity of the dog eyes indicated that this organ, a target for toxicity in the dog, exhibited an affinity for the radiotracer not seen in other species. Histoautoradiographic sections of dog eyes revealed radioactivity localized regionally over the outer layer of the retina, whereas analysis of tissues from rats and mice for 14C activity and histoautoradiography indicated a lack of selective affinity for any ocular tissues (Timchalk et at., 1996). The major metabolites were an oxidation product of the 3-methyl moiety of the phenyl ring and a demethylation of the 3-methoxy moiety of the pyrimidine ring. In studies performed in male rats with 14C-metosulam labelled in either the phenyl or pyrimidine ring, no evidence of cleavage of the sulfonamide bridge was seen (UK Ministry of Agriculture, Fisheries, and Food, 1996). In vitro dermal penetration studies using rat (SpragueDawley) and fresh human skin demonstrate that less than 1% of the applied metosulam actually penetrated the skin (UK Ministry of Agriculture, Fisheries, and Food, 1996).
75.6.3 TOXICITY TO HUMANS
Risk assessment calculations for the general population and for pesticide handlers indicate acceptable risk estimates. Residue studies have indicated no detectable levels in cereal grains at the limit of quantitation. Maximum residue limits of 0.1 ppm in grains based on the limit of quantitation, and an acceptable daily intake (AD!) of 0.01 mg/kg/day on the basis of the chronic NOEL of 5 mg/kg/day in rats and a conservative safety factor of 500 have been proposed. Using these values, the maximum theoretical dietary intake (MTDI) of metosulam from all routes of exposure has been estimated to account for <5% of the AD!. An acceptable operator exposure level (AOEL) of 0.04 mg/kg/day has been calculated using a NOEL of 10 mg/kg/day and a safety factor of 250 (UK Ministry of Agriculture, Fisheries, and Food, 1996).
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Triazolopyrimidine Herbicides
REFERENCES Ahn, K. Y., Park, K. Y., Kim, K. K., and Kone, B. C. (1996a). Chronic hypokalemia enhances expression of the H+K+ -ATPase ct2-subunit gene in renal medulla. Am. J. Physiol. 217, F314-321. Ahn, K. Y., Turner, P. B., Madsen, K. M., and Kone, B. C. (1996b). Effects of chronic hypokalemia on renal expression of the "gastric" H+K+ -ATPase ct-subunit gene. Am. 1. Physiol. 270, F557-566. Bond et al. (1987). Dow AgroSciences, LLC, unpublished data. Bond et al. (1991). Dow AgroSciences, LLC, unpublished data. Bounds (1997). Dow AgroSciences, LLC, unpublished data. Brooks (1997). Dow AgroSciences, LLC, unpublished data. Brown, D., Hirsch, S., and Gluck, S. (1988). Localization of a proton-pumping ATPase in rat kidney. 1. Clin. Invest. 82,2114-2126. Charles River Breeding Laboratories (1987). "Spontaneous Neoplastic Lesions in the Crl:CD BR Rat." Charles River Breeding Laboratories. Clements and Cieszlak (1995). Dow AgroSciences, LLC, unpublished data. Cosse et al. (1989). Dow AgroSciences, LLC, unpublished data. DeFronzo, R. A. (1980). Hyperkalemia and hyporeninemic hypoaldosteronism. Kidney Int. 17, 118-134. Dietrich, D. R., and Swenberg, J. A. (1991). Preneoplastic lesions in rodent kidney induced spontaneously or by non-genotoxic agents: Predictive nature and comparison to lesions induced by genotoxic carcinogens. Mutat. Res. 248, 239-269. Domoradzki et al. (1995). Dow AgroSciences, LLC, unpublished data. Dryzga et al. (1996). Dow AgroSciences, LLC, unpublished data. Eiam-ong, S., Laski, M. E., Kurtzman, N. A., and Sabatini, S. (1994). Effect of respiratory acidosis and respiratory alkalosis on renal transport enzymes. Am. J. Physiol. 267, F390-399. Environmental Protection Agency (EPA) (1993). Pesticide tolerance for flumetsulam. 40 CFR 180, Federal Register 58(207), 57966 (Thursday, October 28, 1993). Environmental Protection Agency (EPA) (l997a). Cloransulam-methyl: Pesticide fact sheet. OPPTS 7501C (available at http://www.epa.gov/opprdOOll factsheets/cloransu.html). Environmental Protection Agency (EPA) (l997b). Cloransulam-methyl: Pesticide tolerances. 40 CR 180, Federal Register 62(182), 49158 (Friday, September 19, 1997) (available at http://www.epa.gov/fedrgstrIEPA-PEST/ 1997/SeptemberIDay-19/p24939.htm). Environmental Protection Agency (EPA) (1998). Diclosulam: Notice of filing of pesticide petitions. 40 CR 180, Federal Register 63(224), 64484 (Friday, November 20, 1998) (available at http://www.epa.gov/fedrgstrIEPA-PEST/ I 998/NovemberlDay-20/p31 066.htm). Garg, L. C. (1991). Respective role ofH-ATPase and H-K-ATPase in ion transport in the kidney. J. Am. Soc. Nephrol. 2, 949-960. Gilbert (l995a). Dow AgroSciences, LLC, unpublished data. Gilbert (l995b). Dow AgroSciences, LLC, unpublished data. Gilbert (l995c). Dow AgroSciences, LLC, unpublished data. Gilbert (l995d). Dow AgroSciences, LLC, unpublished data. Gilbert and Yano (l995a). Dow AgroSciences, LLC, unpublished data. Gilbert and Yano (1995b). Dow AgroSciences, LLC, unpublished data. Grandjean and Szabo (1993). Dow AgroSciences, LLC, unpublished data. Hamm, L. L., and Hering-Smith, K. S. (1993). Acid-base transport in the collecting duct. Semin. Nephrol. 13,246-255. Hanley et al. (1989). Dow AgroSciences, LLC, unpublished data. Hansen (1997). Dow AgroSciences, LLC, unpublished data. Hansen, G. P., Tisher, C. c., and Robinson, R. R. (1980). Response of the collecting duct to disturbances of acid-base and potassium balance. Kidney Int. 17,326-337. Haut et al. (1991). Dow AgroSciences, LLC, unpublished data. Haut et al. (1992a). Dow AgroSciences, LLC, unpublished data. Haut et al. (l992b). Dow AgroSciences, LLC, unpublished data. Jeffries et al. (1995a). Dow AgroSciences, LLC, unpublished data. Jeffries et al. (l995b). Dow AgroSciences, LLC, unpublished data. Johnson (1996). Dow AgroSciences, LLC, unpublished data. Johnson et al. (1997). Dow AgroSciences, LLC, unpublished data. Lawlor (1995). Dow AgroSciences, LLC, unpublished data.
Liberacki and Camey (1997). Dow AgroSciences, LLC, unpublished data. Liberacki et al. (1996). Dow AgroSciences, LLC, unpublished data. Liberacki et al. (1997). Dow AgroSciences, LLC, unpublished data. Lick et al. (1995). Dow AgroSdences, LLC, unpublished data. Linscombe et al. (1995a). Dow AgroSciences, LLC, unpublished data. Linscombe et al. (1995b). Dow AgroSciences, LLC, unpublished data. Madsen, K. M., and Tisher, C. C. (1986). Structural-functional relationship along the distal nephron. Am. J. Physiol. 250, Fl-15. Madsen, K. M., Verlander, J. w., Kim, J., and Tisher, C. C. (1991). Morphological adaptation of the collecting duct to acid-base disturbances. Kidney Int. Suppl. 33, S57-63. Mattsson et al. (1996). Dow Dow AgroSciences, LLC, unpublished data. Mattsson and McGuirk (1997). Dow AgroSciences, LLC, unpublished data. Minnema (1996a). Dow AgroSciences, LLC, unpublished data. Minnema (l996b). Dow AgroSciences, LLC, unpublished data. Minnema (1996c). Dow AgroSciences, LLC, unpublished data. Morseth (1994). Dow AgroSciences, LLC, unpublished data. Nolan et al. (1995). Dow AgroSciences, LLC, unpublished data. Perkins, 1. M., and Billington, R. (1998). In vitro/in vivo correlation for skin absorption of a spray solution of the herbicide florasulam. In "Sixth lnt. Conf. on the Perspectives in Percutaneous Penetration," Leiden, Holland. Pottenger et al. (1991). Dow AgroSciences, LLC, unpublished data. Quast et al. (1997). Dow AgroSciences, LLC, unpublished data. Redmond and 10hnson (l996a). Dow AgroSciences, LLC, unpublished data. Redmond and 10hnson (1996b). Dow AgroSciences, LLC, unpublished data. Redmond and Kociba (1996). Dow AgroSciences, LLC, unpublished data. Scortichini and Kociba (1997). Dow AgroSciences, LLC, unpublished data. Shankar and Johnson (1996). Dow AgroSciences, LLC, unpublished data. Shankar et al. (1993). Dow AgroSciences, LLC, unpublished data. Spencer et al. (1995). Dow AgroSciences, LLC, unpublished data. Stebbins and Haut (1993). Dow AgroSciences, LLC, unpublished data. Stebbins and Haut (1994). Dow AgroSciences, LLC, unpublished data. Stebbins and Haut (1997). Dow AgroSciences, LLC, unpublished data. Stebbins et al. (1990). Dow AgroSciences, LLC, unpublished data. Stebbins et al. (1996). Dow AgroSciences, LLC, unpublished data. Stewart et al. (1992a). Dow AgroSciences, LLC, unpublished data. Stewart et al. (1992b). Dow AgroSciences, LLC, unpublished data. Stewart et al. (1996). Dow AgroSciences, LLC, unpublished data. Stokes, J. B. (1993). Ion transport by the collecting duct. Semin. Nephrol. 13, 202-212. Stott et al. (1986). Dow AgroSciences, LLC, unpubliahed data. Stott et al. (1991). Dow AgroSciences, LLC, unpubliahed data. Sullivan and Cronin-Singleton (1995). Dow AgroSciences, LLC, unpubliahed data. Sullivan and Singleton (1995). Dow AgroSciences, LLC, unpublished data. Swaim and Szabo (1992). Dow AgroSciences, LLC, unpublished data. Szabo and Davis (1992). Dow AgroSciences, LLC, unpublished data. Szabo and Davis (l993a). Dow AgroSciences, LLC, unpublished data. Szabo and Davis (1993b). Dow AgroSciences, LLC, unpublished data. Szabo and Davis (1994). Dow AgroSciences, LLC, unpublished data. Szabo and Rachunek (1992). Dow AgroSciences, LLC, unpublished data. Szabo et al. (1992). Dow AgroSciences, LLC, unpublished data. Timchalk et al. (1988). Dow AgroSciences, LLC, unpublished data. Timchalk, c., Dryzga, M. D., Johnson, K. A., Eddy, S. L., Freshour, N. L., Kropscott, B. E., and Nolan, R. J. (1996). Comparative pharmacokinetics of [14C]metosulam (N[2,6-dichloro-3-methylphenyl]-5,7-dimethoxy1,2,4-triazolo-[1,5a]-pyrimidine-2-sulfonamide) in rats, mice and dogs. J. Appl. Toxicol. 17,9-21. Tsuruoka, S., and Schwartz, G. J. (l996a). Adaptation ofrabbit cortical collecting duct HC0 transport to metabolic acidosis in vitro. J. Clin. Invest. 97, 1076-1084. Tsuruoka, S., and Schwartz, G. J. (1996b). Metabolic acidosis stimulates H+ secretion in the perfused rabbit outer collecting duct of the inner stripe. 1. Am. Soc. Nephrol. 7, 1262. UK Ministry of Agriculture, Fisheries, and Food (1996). "Evaluation of Fully Approved or Provisionally Approved Products. Evaluation on: Metosulam."
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Pesticide Safety Directorate, Issue 13, No. 148, Ministry of Agriculture, Fisheries, and Food. Vedula et al. (1992). Dow AgroSciences, LLC, unpublished data. Verlander, J. W., Madsen, K M., and Tisher, C. C. (1991). Structural and functional features of proton and bicarbonate transport in the rat collecting duct. Semin. Nephrol. 11,465-477. Verlander, J. w., Madsen, K M., Cannon, J. K, and Tisher, C. C. (1994). Activation of acid-secreting intercalated cells in rabbit collecting duct with ammonium chloride loading. Am. J. Physiol. 266, F633-645. Walker (1996). Dow AgroSciences, LLC, unpublished data. Weiner (1997). Dow AgroSciences, LLC, unpublished data. Weiner, 1. D., and Wingo, C. S. (1997). Hypokalemia---consequences, causes and correction. J. Am. Sac. Nephrol. 8, 1179-1188.
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Wingo, C. S., and Cain, B. D. (1993). The renal H-K-ATPase: Physiological significance and role in potassium homeostasis. Ann. Rev. Physiol. 55, 323347. Yano et al. (1987). Dow AgroSciences, LLC, unpublished data. Yano et al. (1988). Dow AgroSciences, LLC, unpublished data. Yano et at. (1991). Dow AgroSciences, LLC, unpublished data. Zablotny (1996). Dow AgroSciences, LLC, unpublished data. Zablotny and Carney (1997). Dow AgroSciences, LLC, unpublished data. Zablotny et al. (1993). Dow AgroSciences, LLC, unpublished data. Zablotny et al. (1994). Dow AgroSciences, LLC, unpublished data. Zablotny et al. (1996). Dow AgroSciences, LLC, unpublished data. Zempel et al. (1988). Dow AgroSciences, LLC, unpublished data. Zempel et al. (1990). Dow AgroSciences, LLC, unpublished data. Zielke et al. (1988). Dow AgroSciences, LLC, unpublished data.
CHAPTER
76 Inhibitors of Aromatic Acid Biosynthesis Donna Farmer Monsanto Company
76.1 INTRODUCTION
Structure
Glyphosate is a broad-spectrum, postemergent systemic herbicide with activity on essentially all annual and perennial plants. Glyphosate-based formulations are used worldwide in virtually every phase of agricultural, industrial, silvicultural, and residential weed control. Due to low solubility in water, glyphosate is typically formulated into commercial products in the form of a salt. Glyphosate is poorly absorbed both dermally and via oral exposure and it is not biotransformed. It has been shown that glyphosate does not bioaccumulate. Animal studies indicate that glyphosate is essentially nontoxic via acute oral and dermal exposure, and that glyphosate salts are nonirritating to the eyes and skin. Glyphosate does not produce dermal sensitization in guinea pigs. In repeated dose studies in laboratory animals, treatment-related effects included reduced body weight gain, increased liver weights, degenerative ocular lens changes, and microscopic liver changes but only at very high dose levels (approximately 2-3% of the diet). No treatment-related tumors have been found in multiple carcinogenicity studies. Glyphosate has consistently produced negative results in standard mutagenicity assays conducted according to international guidelines. Regulatory agencies and other scientific organizations have concluded that glyphosate is neither carcinogenic nor mutagenic. The U.S. Environmental Protection Agency (U.S. EPA) USEPA has classified glyphosate in Category E ("Evidence of Non-Carcinogenicity in Humans"). There is no evidence of developmental or reproductive effects resulting from glyphosate exposure. In humans, accidental exposure to glyphosate formulations may result in minor, transient ocular and dermal irritation, but serious effects have not been observed.
The structure of glyphosate is shown in Fig. 76.1.
Synonyms The common name glyphosate is in general use. Trade names include Roundup®, RoundupUltra®, RoundupPro®, Landmaster®, Rodeo®, Accord®, Spark®, Vision®, and Biactive®. The CAS registry number for the acid is 1071-83-6. Physical and Chemical Properties Glyphosate acid is typically referred to as the technical grade material and has the empirical formula C3HgN05P. It is a white, odorless, crystalline powder with a melting point of 184.5°C, a molecular weight of 169.1, and a specific gravity of 1.704. Glyphosate is not flammable, is not explosive and has a vapor pressure of 1.84 x 10-7 mm Hg at 45°C. Glyphosate is a relatively strong acid with a pH of 2 in 1% aqueous solution. The solubility of glyphosate in water is 1.2 wt% at 25°C and approximately 6 wt% at 100°C. It is slightly soluble in a few strong organic acids but relatively insoluble in most organic solvents. Because of its limited solubility in water, commercial herbicide formulations contain glyphosate in the form of a salt (i.e., isopropylamine, ammonium, phosphonium, etc.) (Franz et aI., 1997). History, Formulations, and Uses Gyphosate is a broadspectrum, nonselective, postemergent, systemic herbicide with activity on essentially all annual and perennial plants. The herbicidal properties of glyphosate were discovered by Monsanto in 1970, and the first commercial formulations were introduced in 1974 under the Roundup brand name. Today, glyphosatebased formulations are used in over 100 countries in virtually every phase of agricultural, industrial, silvicultural, and residential weed control, making it one of the most important weed-pest control tools ever introduced. Agricultural use of
76.2 GLYPHOSATE 76.2.1 IDENTITY, PROPERTIES, AND USES Chemical name cine.
Glyphosate is N-(phosphonomethyl) gly-
Handbook of Pesticide Toxicology Volume 2. Agents
Figure 76.1
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Glyphosate acid. Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
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glyphosate continues to expand. It has contributed significantly to the growing worldwide adoption of conservation and reduced tillage techniques as well as applications involving genetically modified plant varieties which can tolerate glyphosate treatment. 76.2.2 TOXICITY TO LABORATORY ANIMALS Standard toxicity studies have been performed with technical grade glyphosate (averaging 96% purity on a dry weight basis). The results have been summarized in the reregistration eligibility decision (RED) document issued by the United States Environmental Protection Agency (U.S. EPA, 1993), the World Health Organization (WHO, 1994), and Williams et al. (2000). These results demonstrate that glyphosate has very low acute toxicity and is not mutagenic, teratogenic, carcinogenic, or a reproductive toxicant. Acute Studies The oral LDso in rats is >5000 mg/kg and the dermal LDso in rabbits is >5000 mg/kg (WHO, 1994). An acute rat inhalation study has not been conducted with glyphosate technical because it is a nonvolatile solid material which would not generate a respirable vapor or particulate under circumstances of normal use. However, a 4-hour LCso for an aqueous solution of the isoproplyamine salt of glyphosate was shown to be greater than the highest attainable atmospheric concentration of 1.3 mg/l (Dudek, 1987). Irritation and Sensitization Studies Glyphosate technical produced mild skin irritation after a single 4-hour exposure. However, glyphosate did not produce dermal sensitization in guinea pigs (U.S. EPA, 1993). Glyphosate, when applied undiluted and without a wash, was severely irritating to the eyes (WHO, 1994). In contrast, the neutral pH isoproplyamine and monosodium salts are nonirritating to the eyes (Branch, 1981; Busch, 1987). Dose Studies Several subchronic and chronic toxicology studies have been conducted on glyphosate, and the results of these investigations have been reported by the U.S. EPA (1993), WHO (1994), and Williams et al. (2000). The major findings of these studies are summarized in this chapter. In a 3-month feeding study with Sprague-Dawley rats, the noobserved-effect level (NOEL) was 20,000 ppm (approximately 1445 mg/kg/day), the highest dose tested. Administration of glyphosate to CD-l mice for 3 months at dietary levels of 0, 5000, 10,000, and 50,000 ppm resulted in reduced body weight gains in high-dose animals. The NOEL was 10,000 ppm (approximately 2300 mg/kg/day). Glyphosate was applied to the shaven intact and abraded skin of New Zealand white rabbits for 6 hour per day, 5 days per week for 3 weeks at dose levels of 0, 100, 1000, and 5000 mg/kg/day. A slight degree of dermal irritation was observed at the site of application in the high-dose group. No adverse effects were
noted in the hematologic, biochemical, and histopathological evaluations. The systemic NOEL was considered to be 5000 mg/kg/day. Glyphosate was given to beagle dogs via oral capsule at dosages of 0, 20, 100, or 500 mg/kg/day for 1 year. No treatment-related effects were noted even at the highest dose tested; therefore, the NOEL was considered to be 500 mg/kg/day. Brahman-cross heifers received daily dosages of the isopropylamine salt of glyphosate via stomach tube for seven consecutive days at dosages of 0, 540, 830, 1290, and 2000 mg/kg/day. Mortality was observed only at the two highest doses. Other effects, including body weight loss, diarrhea, serum chemistry changes, and histopathological findings were observed at or above 830 mg/kg/day. Changes in several hemato logic parameters observed at 1290 mg/kg/day and above were considered secondary to fluid and blood volume alterations resulting from the diarrhea. The NOEL was considered to be 540 mg/kg/day. Three rodent bioassays were conducted with glyphosate. In the first of two long-term feeding studies conducted in Sprague-Dawley rats, glyphosate was administered in the diet at concentrations of 0, 60, 200, and 600 ppm for approximately 26 months. The NOEL was considered to be >600 ppm (32 mg/kg/day) because no tumors or other adverse effects related to treatment were noted at any dose level. In the second chronic study, rats were fed glyphosate in the diet at concentrations of 0, 2000, 8000, and 20,000 ppm for approximately 2 years. No tumors related to treatment were observed. The only effects considered related to treatment were observed in high-dose animals and included decreased body weight gain in females and degenerative ocular lens changes, increased liver weights, and elevated urine pH or specific gravity in males. The NOEL in this study was concluded to be 8000 ppm (409 mg/kg/day). Glyphosate was fed to CD-l mice in the diet at concentrations of 0, 100, 5000, or 30,000 ppm. No treatment-related tumors were observed. The NOEL in this study was concluded to be 5000 ppm (750 mg/kg/day) based upon reduced body weight gains in high-dose males and females and microscopic liver changes (central lobular hepatocyte hypertrophy and hepatocyte necrosis) in high-dose males. Absorption, Distribution, Metabolism, and Excretion Absorption of glyphosate across skin and gastrointestinal membranes is minimal. In vitro absorption of glyphosate through human skin was no more than 2% of applied dose (Wester et aI., 1991). Wester et al. (1991) also reported the in vivo dermal absorption of glyphosate in the rhesus monkey to be 2.2% at a high dose of 5400 I-Lg/cm 2 . The results of several studies show that there is rapid elimination, no biotransformation, and minimal tissue retention of glyphosate in various species, including mammals, birds, and fish (U.S. EPA, 1993; WHO, 1994). Greater than 90% of an orally administered dose of glyphosate is rapidly eliminated in 72 hours (National Toxicology Program, 1992). Typically, approximately 70% of the administered dose is eliminated in the feces, with the remainder eliminated in the urine. In all cases, less than 0.5% of the administered dose is found in the tissues and organs, demonstrating
76.2 Glyphosate
that glyphosate does not bioaccumulate in edible tissues. Studies of the metabolism of glyphosate in experimental animals (rats, rabbits, lactating goats, and chickens) indicate that it is not biotransformed, with essentially all the administered dose excreted as unchanged parent molecule (Bodden, 1988; Colvin and Miller, 1973; Ridley and Mirly, 1988). Genotoxicity Studies Glyphosate was negative in well-validated mutagenicity assays performed for regulatory purposes conducted according to international guidelines under good laboratory practices (U.S. EPA, 1993; WHO, 1994). These assays assessed a variety of end points both in vitro and in vivo and included the following: Salmonella typhimurium (Ames assay), Escherichia Coli WP-2 reverse mutation, rec-assay with Bacillus subtilis, CHOIHGPRT, in vivo mouse bone marrow micronucleus, and in vitro hepatocyte primary culture-DNA repair assay. Williams et al. (2000), in a review on glyphosate, employed a weight-of-evidence evaluation of the many genotoxicity assays including those submitted for regulatory purposes as well as others in the published scientific literature. It was concluded that glyphosate is neither mutagenic nor clastogenic. A limited number of studies in the literature have reported positive results regarding the genotoxic potential of glyphosate; review by these authors found that these assays used toxic dose levels, irrelevant routes of exposure, end points, and test systems, and/or deficient testing methodology. Carcinogenicity Studies Regulatory agencies and other scientific organizations have concluded that glyphosate is neither carcinogenic nor mutagenic. In June of 1991, the U.S. EPA following a thorough review of all toxicology data available concluded that glyphosate should be classified in Category E ("Evidence of Non-Carcinogenicity in Humans"). This classification was based upon the observation of no treatment-related tumors at any dose level with glyphosate tested up to the limit dose in rats and up to levels higher than the limit dose in mice, and upon the lack of evidence for mutagenicity with glyphosate (U.S. EPA, 1992). Mode of Action Glyphosate's mode of action has been previously described in detail (Franz et al., 1997). Glyphosate inhibits plant growth through competitive inhibition of the enzyme 5-enolpyruvoylshikimate 3-phosphate synthase (EPSPS). This enzyme plays a key role in the biosynthesis of the intermediate, chorismate, necessary for the synthesis of the essential amino acids phenylalanine, tyrosine, and tryptophan. This aromatic amino acid biosynthetic pathway (shikimic acid pathway) is found in plants as well as some fungi and bacteria but not in insects, birds, fish, mammals, and humans, thus providing a specific selective toxicity to plant species. Developmental and Reproduction Studies The reproductive and developmental toxicity data base has been evaluated by the U.S. EPA (1999). This assessment was conducted under the Federal Food, Drug, and Cosmetic Act, as amended by the Food Quality Protection Act (FQPA). The FQPA was enacted
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by the U.S. Congress in 1996 with provisions that support the governments focus on children's environmental health risks. It requires the U.S. EPA to more carefully consider the risks posed to infants and children by pesticide residues on food when setting acceptable residue levels and tolerances. The U.S. EPA is required to apply an additional 10-fold safety factor to ensure the protection of infants and children unless a determination can be made on the basis of reliable data that a lesser margin of safety is protective. As a result of their assessment, the EPA may also require additional testing to detect potential developmental neurotoxic effects. Regarding glyphosate, EPA has concluded that there is a complete toxicity data base and exposure data is complete or can be estimated based on data that reasonably accounts for potential exposures. It was concluded there is no indication that the developing fetus or neonate is more sensitive than adult animals. Consequently no developmental neurotoxicity studies were required. The EPA believes that reliable data support the use ofthe standard lOO-fold uncertainty factor and concluded that there is a reasonable certainty that no harm will result to infants and children from aggregate exposure to glyphosate residues. The studies supporting these conclusions are summarized next. Sprague-Dawley rats were dosed by gavage at doses of 0, 300, 1000, or 3500 mg/kg/day during days 6-19 of gestation. At 3500 mg/kg/day, the following signs of toxicity were observed: increased mortality (6 of 25 dams died) and other clinical signs of toxicity, decreased fetal weights, increased incidence of early resorptions, decreases in total number of implantations and the number of viable fetuses, and increased number of fetuses with reduced ossification of sternebrae. At the lower dose levels these effects were absent. There was no evidence of teratogenicity at any dose level. The NOEL for both maternal and developmental toxicity was 1000 mg/kg/day. In Dutch belted rabbits, glyphosate was tested at dose levels of 0, 75, 175, or 350 mg/kg/day from days 6 through 27 of gestation. The maternal NOEL was determined to be 175 mg/kg/day based on maternal signs of toxicity seen only at the highest dose tested. These effects included death (10 of 16 does died), diarrhea, and nasal discharge. Excessive maternal mortality resulted in an inadequate number of fetuses for evaluation at 350 mg/kg/day. Therefore, although no developmental toxicity was observed at any dose level, the developmental NOEL was considered to be 175 mg/kg/day. Glyphosate was administered to Sprague-Dawley rats in the diet at dosages of 3, 10, and 30 mg/kg/day for three successive generations (2 litters per generation). There were no treatment-related effects on mating, fertility, or other reproductive parameters. An equivocal increase was noted for the incidence of unilateral renal tubular dilation in male pups of the F3b generation in the high-dose group. The small increase in incidence was not considered related to treatment. This conclusion is supported by the absence of a similar effect in a more recent study which evaluated substantially more animals and used significantly higher dose levels (3% of the diet). In the more recent reproduction study, Sprague-Dawley rats were administered glyphosate in the diet at dosages of 0,
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2000, 10,000, and 30,000 ppm (equivalent to 0, 100,500, and 1500 mg/kg/day). There was no effect on the ability of treated rats to mate, conceive, carry, or deliver normal offspring. The systemic NOEL was 10,000 ppm based on soft stools and decreased body weights and pup weights during the second and third weeks of lactation. The reproductive NOEL was the highest dose tested, 30,000 ppm. 76.2.3 HUMAN EXPERIENCE No evidence was observed for the induction of photoirritation nor of allergic or photoallergic contact dermatitis when Roundup herbicide (41 % IPA salt of glyphosate, water, and a surfactant) was evaluated in 346 volunteers (Maibach, 1986). Roundup was less irritating than a standard dishwashing detergent and a general all-purpose cleaner and was no different than baby shampoo. Acquavella et al. (1999) evaluated effects from 1513 human ocular exposures to various Roundup formulations reported to an American Association of Poison Control Centers (AAPCC) certified regional poison control center during the years 1993 through 1997. The majority of the reported exposures were judged by the poison center specialists to result in either no injury (21 %) or transient minor symptoms (70%). In no case did exposure result in permanent change to the structure or function of the eye. The exposure potential of the general population and applicators to glyphosate have been reviewed by Williams et al. (2000). Exposure of the general population to glyphosate is very low and occurs primarily from the diet. Glyphosate has been registered for use in food crops for over 20 years, and glyphosate is now used in a wide range of crops. The initial uses for glyphosate were for preplanting or preemergence applications and resulted in negligible residues in the crops. Later uses have included applications when the crops are present, either using directed spray techniques, applications close to harvest, or herbicide-tolerant crops. These uses can result in residues in edible commodities, although they are still at very low levels. The reference dose (RID) for glyphosate based on the developmental toxicity study with rabbits (NOEL of 175 mg per kilogram of body weight per day) and using a lOO-fold safety factor is calculated to be 2.0 mg(kg body weight)day (U.S. EPA, 1999). The RID represents the level at or below which daily aggregate dietary exposure over a lifetime will not pose appreciable risks to human health. The U.S. EPA generally has no concern for exposures below 100% of the RID. The theoretical maximum residue contributions (TMRC) and percentage of RIDs for the U.S. popUlation was estimated to be 0.029960 or 1.5% of the RID and 0.064388 or 3.2% or the RID for children (1-6 years old) (U.S. EPA, 1999). Because the qualitative nature of glyphosate residues is well understood and the aggregate exposure is not expected to exceed 100% of the RID, the U.S. EPA concludes that there is reasonable certainty that no harm will result from aggregate exposure to glyphosate residues. Dermal contact is the most likely route of exposure for applicators; activities such as mixing and loading of glyphosate
and extended applications using hand sprayers have the highest potential for exposure. Inhalation is considered to be a minimal route of exposure under most circumstances because of glyphosate's extremely low vapor pressure. Biological measurements estimating the amount of pesticide that has penetrated into the body, the internal dose, provide the most relevant information for safety assessments. Lavy et al. (1992) found that, of 355 daily urine samples analyzed from silvicultural workers, none contained quantifiable levels of glyphosate, with a limit of quantification of 10 ppb. Cowell and Steinmetz (1990) found that, of 96 urine samples analyzed from silvicultural workers, only 5 contained quantifiable levels of glyphosate. The highest measurement was 14 ppb. In a recent pilot study with three farmers and their families, there were no quantifiable residues of glyphosate in the study except one farmer with a urinary glyphosate measurement of 12 ppb on the day of a 5hour, hand-wand sprayer application to weeds along a fence line (Alexander et aI., 1999). This application method is similar to those in the previous silvicultural studies. In a worst case analysis, Williams et al. (2000) estimated that an adult worker's peak acute exposure to glyphosate during application was 56.2 J.!g(kg body weight)day and, for a 5-day working week, the chronic applicator exposure was 8.5 J.!g(kg body weight)day. Comparison of these values to lowest relevant NOEL of 175 mg/kg/day in a the rabbit developmental toxicity study produced margins of exposure (MOEs) of 3114 and 20,588 in acute and chronic exposure, respectively. Actual exposures are anticipated to be significantly less. lauhiainen et al. (1991), in addition to biological and inhalation monitoring of glyphosate of forestry workers during application, had each worker receive a medical examination on the first and last days that Roundup® herbicide was applied and a follow-up examination 3 weeks after the last application day. No changes were noted in hematology, clinical chemistry, electrocardiogram, pulmonary function, blood pressure, or heart rate. Accidental exposures to small volumes of glyphosate have not produced serious effects. In spite of this experience, it has been stated that glyphosate is a leading cause of pesticide poisoning in California. California's Department of Pesticide Regulation (CD PR) pesticide incident program accepts telephone inquiries from physicians, who are required to report pesticide incidents, as well as from the general public. Although many calls are purely informational or report effects limited to topical irritation, all telephone calls are recorded as "poisonings," incorrectly suggesting some degree of systemic illness. In 1994, CDPR reported only l3 "definite" or "probable" calls related to glyphosate exposure alone among the 1995 total calls received (California Environmental Protection Agency, 1996). Eleven of these 13 cases reported only minor and reversible eye irritation likely due to accidental overexposure. Of the remaining two cases, one involved a worker who reported a headache in addition to eye irritation. The other case involved symptoms related to ingestion and/or aspiration of hydrocarbon solvent. The latter case cannot be related to glyphosate itself or to a marketed formulation, because commercial preparations of glyphosate are not formulated using hydrocarbon solvents. The
References
CDPR noted in its 1994 Pesticide Illness Surveillance Report that greater than 80% of the people affected by glyphosate experienced only irritant effects and that, of the 515 pesticide-related hospitalizations recorded over the 13 years on file, none was attributed to glyphosate. Statements to the effect that glyphosate is a major cause of clinical poisoning in California are clearly not substantiated by the available data. It has become customary to generically refer to any organic compound containing phosphorous as an "organophosphate." However, there are actually different classes of "organic phosphate" compounds that are determined by the atoms attached to the phosphorus. The phosphorus atom of a true organophosphate is attached only to oxygen atoms (O'Brien, 1967). The structure of glyphosate is different in two important respects. First, the phosphorus atom is attached to the remainder of the molecule by a carbon atom, not an oxygen. This classifies the glyphosate molecule as an organophosphonate (O'Brien, 1967). Second, there are no other side chains attached to the phosphorus atom. Glyphosate consists of a glycine moiety and a phosphonomethyl moiety. These distinctions are important for the following reason. The nature of the groups attached to the phosphorous determine how strongly the molecule will interact with the enzyme cholinesterase (O'Brien, 1967). The groups attached to the phosphorus of a true organophosphate allow it to be readily hydrolyzed by cholinesterase. In contrast, the carbon-phosphorus bond of an organophosphonate is not easily broken (Roberts and Caserio, 1965). This interaction is responsible for the subsequent inhibition of the enzyme and disruption of normal nerve function. The phosphorus of an organophosphonate such as glyphosate, on the other hand, does not react with or inhibit cholinesterase. Thus, this type of molecule does not interfere with normal nerve function (Williams et aI., 2000). Large amounts of glyphosate-based herbicides are occasionally deliberately ingested to attempt suicide and may result in serious gastrointestinal, cardiovascular, pulmonary, and renal effects and possibly death. Aggressive supportive care is recommended (Tominack et aI., 1989). Glyphosate is sometimes mistakenly referred to as an organophosphate, thus contributing to the incorrect perception that glyphosate is a cholinesterase inhibitor similar to the organophosphate insecticides. As a result, some patients who have ingested glyphosate-based herbicides have received inappropriate medical treatment which may have worsened their condition. Atropine or 2-PAM (Pralidoxime) are not indicated in the treatment of glyphosate exposure.
REFERENCES Acquavella, J. E, Weber, J. A., Cullen, M. R., Cruz, O. A., Martens, M. A., Holden, L. R., Riordan, S., Thompson, M., and Farmer, D. R. (1999). Human ocular effects from self-reported exposure to Roundup herbicides. Hum. Exp!. Toxico!. 18,479-486. Alexander, B. H., Mandel, J. S., Baker, B. A., and Honeycutt, R. (1999). "The Farm Family Exposure Pilot Study." Unpublished draft final report. Bodden, R. M. (1988). "Metabolism Study of Synthetic I3C/14 C-Labeled Glyphosate and Aminomethylphosphonic Acid in Laying Hens." Unpublished report, Hazleton Laboratories America, Inc., Madison, W1.
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Branch, D. K. (1981). "Primary Eye Irritation of Isopropylamine Salt of Glyphosate to Rabbits Eyes." Unpublished report, Monsanto Environmental Health Laboratory, St. Louis, MO. Busch, B. (1987). "Primary Eye Irritation Study of Monosodium Salt of Glyphosate in New Zealand White Rabbits." Unpublished report, Food and Drug Research Laboratories, Inc., Waverly, NY. California Environmental Protection Agency (1996). "California Pesticide Illness Surveillance Program Summary Report 1994." Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento. Colvin, L. B., and Miller, J. A. (1973). "Residue and Metabolism-The Gross Distribution of N-Phosphonylmethylglycine-1 4 C in Rabbits." Unpublished report, Monsanto Company, St. Louis, MO. Cowell, J. E., and Steinmetz, J. R. (1990). "Assessment of Forest Worker Exposures to Glyphosate During Backpack Foliar Applications of Roundup® Herbicide." Unpublished report, Monsanto Company, St. Louis, MO. Dudek, B. R. (1987). "Acute Toxicity of Rodeo® Herbicide Administered by Inhalation to Male and Female Sprague-Dawley Rats." Unpublished report, Monsanto Environmental Health Laboratory, St. Louis, MO. Franz, J. E., Mao, M. K., and Sikorski, J. A. (1997). "Glyphosate: A Unique Global Herbicide," ACS Monograph No. 189. Am. Chem. Soc., Washington,D.C. Jauhiainen, A., Rasanen, K., Sarantila, R., Nuntineg, J., and Kangas, J. (1991). Am. Ind. Hyg. Assoc. J. 52,61-64. Lavy, T. L., Cowell, J. E., Steinmetz, J. R., and Massey, J. H. (1992). Conifer seedling nursery worker exposure to glyphosate. Arch. Environ. Contam. Toxico!. 22,6-13. Maibach, H. I. (1986). Irritation, sensitization, photoirritation, and photosensitization assays with a glyphosate herbicide. Contact Dermatitis 15,152-156. National Toxicology Program (1992). "Technical Report on Toxicity Studies of Glyphosate (CAS No. 1071-83-6) Administered in Dosed Feed to F344fN Rats and B6C3F, Mice." Toxicity Report Series Number 16, NIH Publication 92-3135, July 1992, National Toxicology (NTP), U.S. Department of Health and Human Services, Research Triangle Park, Ne. O'Brien, R. D. (1967). Organophosphates: Chemistry and inhibitory activity. In "Insecticides: Action and Metabolism," pp. 32-54. Academic Press, New York. Ridley, W. P., and Mirly, K. (1988). "The Metabolism of Glyphosate in Sprague-Dawley Rats. I. Excretion and Tissue Distribution of Glyphosate and Its Metabolites Following Intravenous and Oral Administration." Unpublished report, Monsanto Environmental Health Laboratory, St. Louis, MO. Roberts, J. D., and Caserio, M. S. (1965). Organophosphorus compounds. In "Basic Principles of Organic Chemistry," pp. 1194-1215. Benjamin, New York. Tominack, R. L., Conner, P., and Yamashita, M. (1989). Clinical management of Roundup herbicide exposure. Jpn. J. Toxieol. 2, 187-192. U.S. Environmental Protection Agency (U.S. EPA) (1992). Pesticide tolerance proposed rule. Fed. Reg. 57, 8739-8740. U.S. Environmental Protection Agency (U.S. EPA) (1993). "Re-registration Eligibility Decision (RED): Glyphosate." Office of Prevention, Pesticides and Toxic Substances, U.S. Environmental Protection Agency, Washington, D.C. U.S. Environmental Protection Agency (U.S. EPA) (1999). Glyphosate: Pesticide tolerance, Final Rule-40 CFR, Part 180 [Opp-300835; FRL-6073-51. Fed. Reg. 64(71), 18360-18367. Wester, R. C., Melendres, J., Sarason, R., McMaster, J., and Maibach, H. 1. (1991). Glyphosate skin binding, absorption, residual tissue distribution, and skin decontamination. Fundam. App!. Toxieol. 16,725-732. World Health Organization (WHO) (1994). "Glyphosate," Environmental Health Criteria No. 159. International Programme of Chemical Safety (JPCS), World Health Organization, Geneva. Williams, G. M., Kroes, R., and Munro, 1. e. (2000). Safety evaluation and risk assessment of the herbicide roundup and its active ingredient, glyphosate, for humans. Regul. Toxico!. Pharmacol. 31, 117-165.
CHAPTER
77 Inhibitors of DNA Biosynthesis-Mitosis: Benzimidazoles-The Benzimidazole Fungicides Benomyl and Carbendazim Ronald L. Mull RLM Strategies, Inc.
Leon W. Hershberger DuPont Agricultural Products
77.1 AGRICULTURAL USE AND SCIENTIFIC INTEREST The benzimidazole fungicides benomyl and carbendazim are extensively used worldwide on a variety of crops (vegetables, fruits, nuts, cereals, cotton, omamentals, mushrooms, and others) for numerous fungal pests. They have been in use for approximately 30 years with an excellent safety record. They are toxicologically very similar because benomyl is rapidly converted to carbendazim. Therefore, it is advisable to consider information available from both products to have a complete understanding of the toxicity of benomyl. Studies with both benomyl and carbendazim are considered in this chapter. A complete contemporary toxicology database exists to support the widespread registration of these products. Relatively little of this data has been published in the open literature by the pesticide manufacturers, but most of the available data have been reviewed and commented upon by panels of experts and are available to the pUblic. The World Health Organization (WHO) Environmental Health Criteria (EHC) 148 (benomyl) and 149 (carbendazim) were published in 1993 (WHO, 1993a, b); the Food and Agriculture Organization (FAO) and WHO published the Joint FAOIWHO Meeting on Pesticide Residues (JMPR) 1995 evaluation of be no my I and carbendazim toxicology in 1996 (FAOIWHO, 1996a, b) and 1998 evaluation of residues in 1999 (FAOIWHO, 1999a, b). These documents will serve as excellent resources for the reader wishing to probe further than is done here. References in this chapter to unHandbook of Pesticide Toxicology Volume 2. Agents
published work may be found in the WHO and FAOIWHO documents, along with summaries of individual studies. Benomyl and carbendazim are of low acute toxicity and present little likelihood of acute systemic illness from their use in agriculture; in fact, none have been reported from either product after years of widespread use. However, both have been extensively studied because of reproductive and developmental effects seen in laboratory animals given high oral doses and because they cause aneuploidy (numerical chromosome aberrations) in mammalian cells both in vitro and in vivo. Studies employing new techniques have demonstrated thresholds for the aneugenic effects. Their fungicidal activity is due to the ability to bind to tubulin and disrupt microtubule assembly during cell replication. Studies have shown that mammalian cells have a much lower capacity for binding in comparison to fungal tubulin. Benomyl, but not carbendazim, is presently registered in the United States; however, both products are registered and sold in many countries throughout the rest of the world. Benomyl is sold only as a wettable powder formulation whereas carbendazim is sold in several formulation types. Outside the United States, both are sold alone and as mixtures with other fungicides. These mixtures aid in preventing the resistance that such site-specific pesticides are prone to cause. In the United States, benomyl is commonly mixed in the spray tank with other fungicides. Exposure of the general population is likely limited to residues in the diet whereas the skin is the primary exposure
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Copyright © 200 1 by Academic Press. All rights of reproduction in any form reserved.
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CHAPTER 77 Benzimidazoles
route for persons using these fungicides in agriculture, with limited exposure from inhalation. Extensive available data show that a wide margin of safety exists between potential environmental and occupational exposures and those that cause mammalian toxicity. There are no reports of human ingestion of large amounts of benomyl or carbendazim in cases of accident or suicide.
77.2 POTENTIAL FOR HUMAN HEALTH EFFECTS Benomyl and carbendazim present low potential risk for causing acute illness. They are of relatively low toxicity in all mammalian species tested. Both are well absorbed after oral exposure, but poorly absorbed by dermal exposure. Illness following acute exposure has not been reported except for skin irritation. Benomyl, but not carbendazim, has been implicated in skin sensitization in humans. Given the expected low exposures and the poor absorption through the skin, it is not likely that the exposures of either the general population or agricultural users will result in systemic illness. Exposure of the general population is most likely to occur through residues on food. Estimates of dietary exposure based on regional or national food consumption patterns have been made in Europe, the United States, and elsewhere. These estimates indicate exposures well below the acceptable daily intake (AD!) established by the World Health Organization in 1993 (WHO, 1993a, b). Regulatory bodies and the WHO have concluded that the exposures likely to result from use of these fungicides in agriculture pose little risk of human health effects (FAOIWHO, 1996a, b, 1999a, b; WHO, 1993a, b). Benomyl and carbendazim cause reproductive and developmental effects when given to certain laboratory animals in high oral doses. In laboratory studies, developmental toxicity was demonstrated when benomyl or carbendazim was given by gavage administration. Similar developmental toxicity was not observed in dietary exposure studies. Bolus dose exposures, similar to the gavage laboratory animal dosing, are not expected to occur in humans, and the dietary experiments represent the more likely approximation of any human effects from exposure. These compounds are not directly mutagenic, but have been widely studied because they induce numerical chromosome aberrations (aneuploidy). Disruption of microtubule assembly in the actively dividing cell results in aneuploidy, an effect which has a demonstrable no-ob served-effect level (NOEL) and threshold doses below which effects are not seen (Bentley et aI., 2000). Their fungicidal activity is thought to be due to the ability to bind to tubulin during cell replication, causing nondisjunction and cell death. Studies have shown that mammalian cells have less affinity for this binding than do fungal cells, and human cells have a low capacity for binding. Benomyl and carbendazim were not carcinogenic after lifetime dietary exposures in rats and up to 2 years of feeding in dogs. However, benign liver tumors have been shown following lifetime feeding in mouse strains having high background incidences of these tumors. Tumors were not seen in a mouse strain
with a low background incidence of liver tumors, suggesting that the relevance of the mouse liver tumors to other species is questionable. Efforts to better understand the propensity of the mouse to develop liver tumors continue (Fox and Gonzalez, 1996). Both have been thoroughly evaluated in acute, subchronic, and chronic studies in rats, mice, rabbits, and dogs. In extended feeding studies, the dog was the most sensitive species tested. The 2-year NOEL for benomyl was 13 mg/kg/day (females) and 14 mg/kg/day (males). In carbendazim studies, the NOELs were similar to those of the benomyl dog studies, if adjusted for the molecular weight difference. Based on the 2-year benomyl dog study and applying a lOO-fold safety factor, the AD! for benomyl would be 0.1 mg/kg. Based on the 2-year carbendazim dog study and applying a lOO-fold safety factor, the AD! for carbendazim would be 0.03 mg/kg/day (FAOIWHO, 1996a, b). The potential dietary exposure of humans to benomyl residues was estimated to be 0.000218 mg/kg/day in the United States (Eickhoff et aI., 1989). This is considerably lower than the Environmental Protection Agency (EPA) reference dose (0.05 mg/kg/day). Because postharvest usages of be no my1were removed from the Benlate® fungicide label in 1989, this value is likely to represent an overestimate of current dietary intake. The potential benomyl and carbendazim dietary exposure was evaluated for five different geographic regions by the 1998 JMPR meeting. The international estimated daily intake (EID!) ranged between 1 and 14% ofthe WHO AD! (0.03 mg/kg/day) (FAOIWHO,1999b). Agricultural usage results in primarily dermal exposure to the sprayers or other agricultural workers. Studies of dermal absorption of benomyl and carbendazim, both in vivo and in vitro, show them to be poorly absorbed by this route. Topical effects such as skin irritation and sensitization are the primary complaints from agricultural use. Mild eye irritation is also possible. Inhalation exposure is limited due to the low vapor pressure and to the size of the airborne droplets from spray application, which is generally well above the respirable range. No systemic toxicities in humans have been shown to result from either benomyl or carbendazim exposure. The exposure levels from studies on occupational exposure to benomyl vary. Procedures such as mixing operations that involve the direct handling of fungicide formulations have the highest potential to cause exposure. Average values of potential dermal deposition during mixing, during field reentry and harvesting, and during simulated use on home gardens were 26 mg per mixing cycle, 12 mg on reentry, 5.9 mg per hour of harvesting, and < 1 mg per application cycle, respectively (Everhart and Holt, 1982). Similar results for harvesting were reported by Zweig et al. (1983), who found that benomyl exposures averaged 5.39 mg/hour for strawberry pickers. All values were calculated for "worst case" situations in which no protective clothing would be worn. These estimated exposures must be reduced by the limited degree of skin absorption, to judge potential systemic exposures. To determine potential dermal exposure from contact with foliage during reentry, Liesivuori et al. (1988) studied dissi-
77.3 Physical and Chemical Properties
pation rates for benomyl following applications in three rose greenhouses. The mean half-lives for benomy I and carbendazim were 44 and 53 hours, respectively. When benomyl was sprayed on apples, foliage residues dissipated quickly during the first 3-7 days, but at least 15% of the original deposit remained after 12 days. Exposure by inhalation is potentially greatest in manufacturing operations where handling takes place in an enclosed area on a daily basis. Industrial hygiene surveys in manufacturing facilities over a 4.5-year period (1974-1979) reported Benlate dust concentrations averaged 0.708 mg/m 3 [6-8 hour timeweighted average (TWA); range is 0-2.01 mg/m 3 ] (DuPont, 1979). In an agricultural setting, potential human exposure to benomy I via inhalation was calculated to be 0.08 mg per work cycle for the mixing and loading of formulations for aerial spraying; 0.003 mg for field reentry; 0.002 mg/hour for hand-harvesting; and 0.003 mg per application for simulated home use (Everhart and Holt, 1982). No inhalation exposure to benomyl could be detected among lawn-care workers (limit of detection equivalent to 0.004 mg/m3 ) (Leonard and Yeary, 1990). More recent investigations (Hoekstra et aI., 1996; Lavy et aI., 1993; NIOSH, 1994) have utilized measurements of dermal deposition, inhalation, and excretion of the urinary metabolites 4-hydroxy-2-benzimidazole carbamate (4-HBC) and methyl-5hydroxy-2-benzimidazolecarbamate (5-HBC) to obtain a more refined estimate of systemic exposure. Lavy et al. conducted a year-long monitoring study of tree nursery workers engaged in varying tasks and have estimated the exposures to benomyl based on urinary excretion of 4-HBC and 5-HBC. They estimated the margins of safety to be from 14,000 to 69,400 for a 10-day exposure period and even greater margins of safety for a 30-day period. Hoekstra et al. also found low levels of exposure for greenhouse nursery workers involved in typical operations in Florida. The systemic exposure was estimated from urinary excretion of 5-HBC, and they found the highest average concentration in the urine to be 23.8 Il-mol 5-HBC per mole of creatinine. Additionally, inhalation exposure to both benomyl and BIC (bulylisocyanate) was measured and found to be quite low, the highest average for benomyl being 32.8 Il-g/m 3 and for BIe, 6.6 ppb. The American Council of Governmental and Industrial Hygienists (ACGIH) has recommended an 8-hour TWA threshold limit value (TLV) of 10 mg/m 3 for benomyl and carbendazim. The Occupational Safety and Health Administration (OS HA) permissible exposure limit (PEL) is 10 mg/m3 (total dust) and 5 mg/m3 (respirable dust). 77.2.1 MEDICAL SURVEILLANCE AND TESTING The primary adverse health consequence associated with exposure to benomyl or carbendazim is contact dermatitis. Patch testing with benomyl confirmed positive dermal reactions in a number of patients (Fregert, 1973; Kuehne et aI., 1985; Savitt,
1675
1972; Schuman and Dobson, 1985; van Joost et aI., 1983; van Ketel, 1976). One report also suggests the possibility of false negative readings for an individual tested with benomy I (Larsen et aI., 1990). Despite the widespread use of benomyl and carbendazim, adverse skin effects are relatively rare (Cronin, 1980; Lisi et aI., 1986; van Joost et aI., 1983). Some researchers have suggested that allergic effects attributed to benomyl may be cross-reactions to other pesticides (Larsen et aI., 1990; Matsushita and Aoyama, 1981). Fertility in male workers was studied due to early animal study results that implicated the testes as a target of toxicity. A group of 286 male workers with potential exposure to benomyl and carbendazim were studied for fertility patterns by assessing the birth rates among their spouses. Workers were divided into three exposure categories: average exposure, below-average exposure, and varied exposures. Birth rates for each group were at or above expected levels when compared to control populations, and it was concluded that benomyl exposure had no adverse effect on fertility rates among exposed workers (Gooch, 1978). Media allegations of a relationship between exposure to benomyl and the occurrence of birth defects prompted investigators in several countries to undertake epidemiology studies. A study was conducted in Italy to look at cases of anophthalmia and their relationship to benomyl exposure. Parental exposure to benomyl did not appear to be a factor, and the authors concluded that an association between benomyl use and congenital microphthalmia and anophthalmia appears unlikely (Spagnolo et aI., 1994). In another study of farm workers in Norway investigators found no association between exposure to benomyl and cases of children born with anophthalmia (Kristensen and Irgens, 1994). A third study in England found no clustering of cases of anophthalmia or microphthalmia in any region in England, refuting media allegations of clusters in farming areas where benomyl may have been used (Dolk et aI., 1988).
77.2.2 EXPOSURE-RELATED ILLNESS The products may cause a temporary contact dermatitis. However, no specific human symptoms of benomyl or carbendazim toxicity are known. In a practical sense, exposures are likely to be from benomyl or carbendazim plus a mixture partner so it is important for the medical provider to try to determine the probable constituents of the mixture. Treatment for overexposure should be symptomatic and supportive.
77.3 PHYSICAL AND CHEMICAL PROPERTIES The structures of benomyl and carbendazim are shown in Fig. 77.1. The physical properties are given in table from Fig. 77.1.
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CHAPTER 77
Benzimidazoles
o 11 CNtt- C41ig
©c Name
I
©:
j-NHCOOCH'
H
N)- NHCOOCH, N
carbendazim
benomyl
Common
Benomyl
Carbendazim
/uPAC
methyl I -(butylcarbamoyl)-benzimidazol-2ylcarbamate
methyl benzimidazol-2-ylcarbamate
CAS
Methyl [I-[(butylarnino)carbonyl]-IHbenzimidazol-2-yl]carbamate
methyllH-benzimidazol-2-ylcarbamate
17804-35-2
CAS number Empirical Formula Molecular Weight
C14H l8 N403
10605-21-7 C9 H9N 302
290.62
191.19
Physical State
Colorless crystal
Colorless crystal
Benomyl
Carbendazim
Melting Point (0C)
Decomposes shortly after starting to melt at about 100°C.
302-307°, with decomposition
Aqueous Solubility (at 25°C)
at pH 5: 3.6 mg/L at pH 7: 2.9 mglL at pH 9: 1.9 mg/L
at pH 4: 28 mglL at pH 7: 8 mglL at pH 8: 7 mg/L
Vapor Pressure
<5.0 x 10-6 Pa @ 25° C
6.5 x 10-8 Pa @ 20° C
Henry's Law Constanta
at pH 5: <4.0 x 10-4 Pa-m 3/mol at pH 7: <5.0 x 10-4 Pa-m 3/mol
at pH 4: 4.42 x 10-7 Pa-m 3/mol at pH 7: 1.55 x 10-6 Pa-m 3/mol
at pH 9: <7.7 x 10-4 Pa-m3/mol
at pH 8: 1.77 x 10-6 Pa-m 3/mol
23.4 (from unbuffered water)
at pH 5: 24 at pH 7: 32 at pH 9: 31
Physical Properties
OctanollWater Partition Coefficient
Dissociation Constant (pKa)
4.48
a Henry's law constant was calculated from the vapor pressure and aqueous solubility data. Thus for carbendazim it was calculated using a vapor pressure determined at 20°C and a solubility determined at 25°C. Figure 77.1
Structures of benomyl and carbendazim.
77.4 ABSORPTION, DISTRIBUTION, METABOLISM, AND EXCRETION The metabolism and toxicokinetics of benomyl and carbendazim have been reviewed by WHO (1993a, b) and JMPR (FAOIWHO, 1996a, b); more recently, the metabolism has been reviewed by Anderson (1999). In general, both benomyl and carbendazim are well absorbed following oral administration (80-85%) but poorly absorbed following dermal administration
(1-2%). After oral and intravenous treatment of rats with benomy I low levels of residues were found in blood and some tissues with minimal bioaccumulation potential. Benomyl is primarily metabolized in animals to carbendazim through the loss of the n-butylcarbamoy1side chain prior to further metabolism (Gardiner et aI., 1974). Carbendazim then undergoes aryl hydroxylation-oxidation of the benzimidazole ring at the 5 and 6 positions followed by sulfate or glucuronide conjugation as the primary metabolic pathway in
77.4 Absorption, Distribution, Metabolism, and Excretion
dogs and rats (Anderson, 1999). The hydrolysis of carbendazim to 2-aminobenzimidazole is another significant pathway in rats (Krechniak and Klosowska, 1986). Krupka (1974) has shown that benomyl does not inhibit either acetyl or butyryl cholinesterase in vitro, as did Belasco (1979a), who examined in vitro inhibition of acetylcholinesterase using bovine red blood cells. Absorption Benomyl and carbendazim were shown to be well absorbed following oral administration to rats, mice, dogs, and hamsters (Belasco, 1969; Culik, 1981a, b; Douch, 1973; Monson, 1990). Benomyl, or its metabolites, appeared in rat blood within one hour following a single oral gavage dose of 14C-benomyl in peanut oil or when benomyl was administered mixed with ground chow at a dose level of 1000 mg per kilogram of body weight (Belasco et aI., 1969). Krechniak and Klosowska (1986) measured urinary excretion of radiolabeled carbendazim and its metabolites 5-HBC and 2-AB (z-amine-lH-benzimidazole) following oral dosing of male rats with radiolabled carbendazim in diethyl glycolethanol. Total absorption was found to be about 85% based on the degree of urinary excretion of the metabolites. In another study, carbendazim was rapidly absorbed (80-85%) in both male and female rats following oral dosing. With absorption of carbendazim being so rapid, the urinary excretion half-life for both males and females was approximately 12 hours, with over 40% of the dose eliminated within the first 12-hour collection interval (Monson, 1990). A dermal absorption study was done with a radiolabeled benomyl formulation (50% wettable powder) in male rats, with four rats killed at each of 0.5-, 1-, 2-, 4-, and lO-hour intervals. A dose of 0.1, 1, 10, or 100 mg of benomyl was applied to a shaved area approximating 16% of the animals' skin. Blood levels of radioactivity peaked between 2 and 4 hours postapplication, reaching 0.05 mg/l in the low dose at 2 hours and 0.1 mg/l at the high dose in 4 hours (Belasco, 1979b). An in vitro study of the penetration through human skin of a similar benomyl formulation applied at a spray-tank dilution showed poor penetration. Even less penetration occurred when the benomyl was applied to the skin dry (Ward and Scott, 1992). Percutaneous absorption of carbendazim was also shown to be very low in rats, with only 0.03% of a 60-mg/rat applied dose excreted after 24 hours (Dom and Keller, 1980). Distribution In rats exposed orally or intravenously, benomy1 and its metabolites were cleared rapidly from blood and exhibited minimal potential for bioaccumulation (Han, 1979; Monson, 1990). Rats examined 24 hours after a single oral dose of radiolabeled 14C-benomyl (1000 mg per kilogram of body weight) had radiolabellevels of 3-13 ppm in blood and 2-4 ppm in the testes (Belasco, 1969). However, no residues of benomyl, carbendazim, 5-HBC, or 4-HBC were detected 24 hours after the last of 10 oral benomyl doses (200 mg/kg/day). This was further confirmed in a study where blood from rats fed a diet containing 2500 ppm benomyl for 1 year had no
1677
detectable amounts of benomyl, carbendazim, or 4-HBC. Although traces of 5-HBC were present (0.2 ppm), no 5-HBC was found in the testes (Belasco, 1969). Studies were reported by Krechniak and Klosowska (1986) in which similar results were seen with carbendazim-treated rats. Metabolism The metabolic pathways for benomyl and carbendazim are shown in Fig. 77.2. Studies show that benomyl is extensively hydrolyzed to carbendazim, which is then further metabolized. A possible route under certain conditions is the conversion of be no my1to STB (3-butyl-l ,3,5triazine (1 ,29)benzimidozoI2,4(lH,3H)dione). Investigations using rats exposed to benomyl by intravenous (Han, 1979), dermal (Belasco, 1979b), and inhalation (WHO, 1993a) routes have shown that 5-HBC is the major urinary metabolite, with some carbendazim also found. The proposed metabolic pathway for metabolism in mammals is shown in Fig. 77 .2. There are two divergent pathways in the biotransformation of benomyl: the release and subsequent degradation of the n-butylcarbamoyl moiety at NI of benomyl, and the metabolism of the remaining carbendazim molecule. The products of benomyl and carbendazim metabolism containing a hydroxyl functional group typically underwent conjugation reactions. Conjugation products were often more polar and more readily excreted than their parent compounds. Production and conjugation of 5-HBC appeared to be the primary path of elimination from mammalian systems. Parent compound accounted for less than 5% of the urinary radiolabel. When the metabolism of carbendazim was compared in rats and mice, only quantitative differences were evident (Dom et aI., 1983). Almost all of the urinary metabolites were in the form of glucuronide and sulfate conjugates, cleavage of which liberated 5-HBC. Excretion Benomyl, carbendazim, and their metabolites were rapidly eliminated in the urine and feces in all mammalian test systems examined. Greater than 98% of the administered radiolabel appeared in the urine or feces of rats within 72 hours following treatment with 14C-carbendazim by using a variety of dosing regimens. Groups of Sprague-Dawley rats were gavaged with a single dose of 50 mg/kg, 14 days preconditioning with gavage dosing of unlabeled carbendazim followed by a single dose of 50 mg/kg of the radiolabeled material or a single gavage dose of 1000 mg/kg (Monson, 1990). Urinary excretion, which was dose dependent, accounted for 41-66% of the dose. The remainder of the radioactivity was virtually all found in the feces, with only traces of radiolabel found in the expired air. Results similar to those in rats were obtained with mice and hamsters (Han, 1978). In mice, 64% of the dose ( 14C-benomyl) was eliminated in the urine and 11. 7 % in the feces during the first 24 hours after dosing. In hamsters, 44.3% was eliminated in the urine and 14.8% in the feces over the same 24-hour period. By 72 hours postdose, 92.7 and 97% had been eliminated in mice and hamsters, respectively. Urinary elimination was by far the major pathway of clearance. However, in the dog this was
1678
CHAPTER 77 Benzimidazoles
o
HO)lN~ H n-BUlylcrbamic acid
~
/
r?r- H
0
V.~~O/ ~_
N
+
Carbendazim
Other metabolites: 4·HBC 5,6-DHBC 5,6-DDBC 5-Hydroxy-6-methoxycarbendazim 5-Hydroxy-2-aminobenzimidazole 5,6-D-hydrodiol-2-aminobenzimidazole
r\-H 0 HO~ ..~~O/ N
o
0
/~
HO~ Acetoacetic acid
Figure 77.2
~
5-HBC
sulfate conjugate
glucuronide conjugate
Metabolism of benomyl and carbendazim in mammals.
not the case. A beagle dog was fed a diet supplemented with unlabeled benomyl for 7 weeks (2500 ppm) and then administered an equivalent 14C-Iabeled dose via capsule. The majority of the dose (83.4%) was eliminated in the feces with only 16.2% of the dose eliminated in the urine over the following 72-hour collection intervaL The reason for such a high level of radio1abel in dog feces cannot be reliably commented on without further characterization of the radiolabeled material found in the feces. Urinary excretion of metabolites was measured following intravenous dosing of radiolabeled carbendazim in albino rats. Twelve hours postdosing the urinary metabolite composition was found to be 94% 5-HBC, 3% 2-AB, and 3% carbendazim (Krechniak and Klosowska, 1986).
77.5 MAMMALIAN TOXICITY 77.5.1 SUMMARY Benomyl and carbendazim have low tOXICIty in acute toxicity studies. The results were consistent in oral (LDso > 1000 mg/kg in the dog to > 10,000 mg/kg in the rat), dermal (LDso > 10,000 mg/kg), and inhalation (LCso > 5.9 mg/l) studies in several animal species. Benomyl, but not carbendazim, caused dermal sensitization in guinea pigs by the maximization test but not by the Buehler method. They are mild to moderate irritants to both the eye and skin. Subchronic toxicity studies showed the gastrointestinal tract, testes, liver, and bone marrow as target organs following high
oral doses of benomyl or carbendazim. The liver was a primary target of toxicity in feeding studies with these fungicides. Mechanistic studies have shown that xenobiotic metabolizing enzymes in the liver are induced. Hepatotoxicity is indicated by changes in serum enzymes, elevated organ weights, or histopathological effects. Following subchronic inhalation exposures to benomyl, rats had localized nasal irritation and degeneration of the olfactory epithelium at high doses although no systemic effects were observed. Later study showed these changes to be specific to the route of exposure (Hurtt et aI., 1993). The NOEL for rats exposed to benomyI by inhalation for 90 days was 10 mg per cubic meter of air. Genetic toxicity studies indicate that benomyl and carbendazim do not induce gene mutations or DNA damage and repair. In addition, neither compound causes structural chromosome aberrations in vivo in somatic or germ cells. However, consistent with their mode of action, these fungicides are positive in tests that assess numerical chromosome aberrations (aneuploidy) in mammalian cells in vitro and in somatic and germ cells in vivo. This activity is observed only when the compounds are administered at relatively high doses. Studies utilizing more sophisticated techniques have shown a threshold for induction of aneuploidy. In chronic feeding studies, rats fed diets that contained up to 2500 ppm of benomyl for 2 years had no signs of chronic toxicity. Rats fed up to 10,000 ppm of carbendazim had body weight, blood, and liver effects. There was no evidence of increased tumor incidence.
77.5 Mammalian Toxicity Mice fed benomyl or carbendazim in long-term studies had notable liver effects. In multiple studies, mice that ingested diets that contained 500 ppm of either fungicide had nonneoplastic lesions in the liver as well as an increased incidence of benign liver tumors. These long-term studies were conducted with CD-l mice, a strain that has a high incidence of spontaneous liver tumors. When carbendazim was fed to a strain of mice with a low incidence of spontaneous tumors, no compound-induced neoplasms were observed. These compounds are known to induce liver enzymes, and the tumor increases in mice are believed to result from induction of cytochrome P450 enzymes and modulation of the growth of spontaneous neoplasms. Liver injury (primarily increases in serum enzyme activities and cholesterol levels) was evident in long-term dog studies with benomyl or carbendazim. Decreased protein and albuminto-globulin ratio in the blood and increased pituitary and thyroid weights were also noted at high doses. The lowest NOEL for 2 years of chronic oral exposure in dogs was 500 ppm, based on clinical and microscopic evidence of liver injury at 2500ppm. Reproductive toxicity studies in rats showed that benomyl and carbendazim cause effects on the reproductive system. Decreased sperm counts, decreased testicular weights, and histopathological changes occurred at higher dose levels. Reproductive changes in rats occurred at approximately the same or at higher doses than those resulting in general toxicity. Accordingly, benomyl and carbendazim are not selective reproductive toxins. The potential for benomyl and carbendazim to cause developmental toxicity was assessed in studies in pregnant rats and rabbits. They are toxic if administered in a single large oral dose but not when given in the diet because of the resulting higher blood levels after bolus dosing. Developmental effects included reduced fetal weight and anomalies of the eyes, skull, and head. In feeding studies, toxic effects to the dams (reduced food consumption and body weight gain) occurred at lower doses than those which produced fetal effects (reduced fetal weights). Results from feeding studies are believed to be more appropriate than gavage studies in determining human risk from daily ingestion or dermal exposure. When benomyl is administered by gavage, some studies show that toxicity to the developing fetus occurs at a lower dose than that causing toxicity to the dam. However, in a more recent expanded rabbit gavage study of benomyl, no fetal effects occurred at levels below maternal toxicity. Both benomyl and carbendazim were found to be negative when evaluated for neurotoxic potential in traditional hen studies. Additionally, no evidence of mammalian neurotoxicity was seen when rats were dosed with benomyl, either in acute doses or in the feed for 90 days. NOELs were based on general toxicity. The NOEL in rats for neurotoxicity from acute exposure to benomyl was 2000 mg/kg, the highest dose tested. In other tests, both compounds were shown not to inhibit acetylcholinesterase.
1679
77.5.2 ACUTE TOXICITY Benomyl and carbendazim as well as formulations made from them were investigated in acute studies in several animal species over the years. They are generally of low oral toxicity, quite low dermal toxicity, and low toxicity by inhalation. They are not classified as irritating to either the skin or the eye. Benomy1 was found not to be a sensitizer by the Buehler method whereas it was found to be a sensitizer by the maximization method (Matsushita et aI., 1977). Ford (1981) studied technical carbendazim and a 75% wettable powder in male guinea pigs, finding no evidence of dermal sensitization. Martin et al. (1987) assessed a 50% carbendazim formulation in both male and female guinea pigs and found no evidence of sensitization. Carbendazim is not a dermal sensitizer. Numerous acute toxicity studies were summarized and reported in WHO (1993a, b) and FAOIWHO (1996a, b). Studies representing the range of findings are presented in Table 77 .2. Benomyl has an oral LDso(rat) > 10,000 mg/kg and an inhalation LDso(rat) of >4 mg/l. Oral LDsos were determined in rats dosed with the minor metabolites 2-AB (>3400 mg/kg), 5-HBC (>7500 mg/kg), BUB (2-(3-butylureido)benzimidazole) (>17,000 mg/kg), and STB (>17,000 mg/kg). Carbendazim (the major metabolite) has been shown to have a similar range of acute toxicity to that of benomyl. The dermal irritation potential of benomyl and carbendazim or their formulations has been assessed. Results varied from "mild irritation" to "not an irritant" (Vick and Brock, 1987a, b). The potential for eye irritation by benomyl and carbendazim and their formulations has been evaluated in rabbits. Edwards (1974a), Vick and Valentine (1987), and have assessed the eye irritation and found them to be mild to moderate irritants. Others studied both technical benomyl and 50% wettable powder formulation as well as a suspension in mineral oil, finding mild conjunctival irritation and minor, transitory corneal opacity (WHO, 1993a). 77.5.3 SUB CHRONIC TOXICITY 77.5.3.1 Oral Several studies of 10 days to 3 months of repeated dosing by both gavage and feeding have been reviewed (Table 77.3). Feeding at the higher doses caused liver changes in rats, dogs, and mice whereas high gavage doses have also resulted in changes in the testis, bone marrow, or gastrointestinal tract. In early studies, both benomyl and carbendazim were given to rats gavaged daily for 5 days/week for 2 weeks. The technical material was given at doses of 0, 200, and 3400, and with carbendazim, 5000 mg/kg/day (Sherman, 1965; Sherman and Krauss, 1966). With benomyl, mortality occurred at 3400 (four of six rats) and with carbendazim, two of six rats died at 3400. Slight testicular changes were noted at the 200 mg/kg dose. Testicular effects, including degeneration of germinal epithelium
1680
CHAPTER 77
Benzimidazoles
Table 77.2 Summary of Acute Toxicity Studies Route and
Sex
Material
species
(No./dose)
tested
Vehicle
Results
Reference
Oral, rat
Male (10)
Benomyl
Peanut oil
LDSO> 10,000 mg/kg
Sherman (l969a)
Carbendazim
Corn oil
LDSO > 5000 mg/kg
Davidse (1975)
Female (10) Oral,
Male (10)
guinea pig
suspension
Oral, mouse
Male (10)
Carbendazim
Propylene glycol
LDSO > 15,000 mg/kg
Oral, rat
Male (10)
Benlate OD
Corn oil
LDso > 12,000 mglkg
Hostetler (1977)
(50% benomyl) Oral, rat
Male/female (10)
Carbendazim
Sesame oil
LDso > 15,000 mg/kg
Kramer and Weigand (1971)
Dermal, rabbit
Male (10)
Carbendazim
Aqueous paste
LDso> 10,000 mg/kg
Edwards (1974a)
Dermal, rabbit
Male/female (5)
Benlate C
Aqueous paste
LDso > 2000 mg/kg
Vick and Brock (1987c)
Fungicide 1991
50% wettable
LDso > 10,000 mg/kg
Busey (1968a)
(benomyl)
powder LCso > 2000 mg/kg
Gargus and Zoetis (I 983c )
LCso > 4.01 mg/L
Busey (1968b)
LCso > 1.65 mg/L
Littlefield and Busey (1969)
Dust
LCso > 5 mg/L
Nash and Ferenz (1982)
Particulate in air
LCSO (1 hr) > 5.9 mg/I
Sarver (1975)
(50%WP, carbendazim) Dermal, rabbit Dermal, rabbit Inhalation, rat Inhalation, dog Inhalation, rat
Male/female (4) Male/female (10) Male (6) Male (10) Male/female (10)
Benlate PNW
50% wettable
(benomyl)
powder
Fungicide 1991
50% wettable
(benomyl)
powder
Fungicide 1991
50% wettable
(benomyl)
powder
Carbendazim, 75%WP
Inhalation, rat
Male (6)
Carbendazim
and reduction or lack of sperm, erosion and thickening of the gastric mucosa as well as liver changes were seen at the higher doses. lanardhan et at. (1987) conducted a 90-day gavage study of carbendazim in rats dosed daily at 0, 16, 32, or 64 mg/kg. They reported numerous hematological and clinical chemistry effects in all dose groups. However, reported alterations were minimal, were out of the expected ranges in the control animals, and did not show a dose response. Absence of raw data and the variability of the results confound interpretation of the results of the study. Ninety-day feeding studies have been conducted in rats with both benomyl and carbendazim. Except for the increased liver weights seen in rats fed benomyl, there were no remarkable differences in the effects noted. In male and female rats fed benomyl diets for 90 days at levels of 0, 100, 500, and 2500 ppm, there was no clinical evidence of toxicity and there were no effects on blood or urine parameters. Female rats fed 2500-ppm diets had slightly elevated liver weights, but no microscopic abnormalities. At the end of the 90-103-day feeding period, 10 of 16 males and females from each dose group were killed and evaluated. The remaining 6 males and females from each group were used in a one-generation reproduction study. No compound-related effects were noted in reproduc-
tive parameters. Gross and microscopic evaluation of tissues and organs showed no significant effects at dietary levels of up to 2500 ppm, equivalent to 198 mg/kg/day for males and 215 mg/kg/day for females (Sherman, 1967). In a similarly designed 90-day feeding study in rats with carbendazim also fed at dietary levels of 0, 100, 500, and 2500 ppm), no compound-related effects were seen during clinical observations, blood or urine analyses, or gross and microscopic examinations. No reproductive effects were found in the subset of animals used for a one-generation reproduction study. The liver-to-body-weight ratio was slightly increased over controls in females in the top dose group. No significant compound-related effects were seen in this study at the highest dose tested, 2500 ppm, equivalent to 197 mg/kg/day for male and 210 mg/kg/day for female rats (Sherman, 1968). Ninety-day feeding studies of benomyl and carbendazim were also conducted in dogs. The pattern of toxicity was similar and indications of liver effects were seen with both compounds. Male and female beagle dogs were fed diets containing 0, 100, 500, or 2500 ppm of benomyl. Clinical chemistry results indicative of compound-related liver effects, which included a slight increase in the activity of alkaline phosphatase and glutamic pyruvic transaminase and a decrease in the albuminto-globulin ratio, were seen at the top dose. There were no
77,5 Mammalian Toxicity
1681
Table 77.3 Summary of Subchronic Toxicity Studies NOEL
Material
mg/kg/day
ppm
NOAEL
Reference
2S00ppm (l98-2IS mg/kglday)
Sherman (1967)
2S00ppm
Sherman (1968)
SOO ppm
Sherman (1968)
2,7 (M&F)
SOO ppm [14.4 mglkglday (M) and 11.3 mg/kg/day (F)]
Sherman (l970a)
7.5 (M&F)
300 ppm
Tit et al. (1972)
None
None
Busey (l968d)
NA
2S0
2S0 mg/kg/day
Hood (1969)
NA
10 mg/m 3
lOmg/m3
Warheit et al. (1989)
0, 200, 3400 mg/kglday
NA
200
200 mg/kg/day
Sherman and Krauss (1966)
Carbendazim technical
0, 200, 3400, or SOOO mg/kg/day
NA
None
None
Sherman (l96S)
10-Dose dermal study in rabbits
Carbendazim technical
0, 2000 mg/kg
NA
None (local skin effects at treatment sites)
None
Dashiell (1974)
Enzyme induction in rats and mice (28-day feeding study)
Carbendazim and benomyl
0,10,30,100,300,1000, 3000ppm
Epoxide hydrolase induced at 1000 ppm (NOEL 300 ppm)
Guengerich (1981)
Study
tested
Dose levels
90-Day feeding study in rats
Benomyl formulation
0, 100, SOO, 2S00 ppm
SOO
41 (M) 44 (F)
90-Day feeding study in rats
Carbendazim formulation
0, 100, SOO, 2S00 ppm
2S00
197 (M) 210 (F)
90-Day feeding study in dogs
Benomyl formulation
0, 100, SOO, 2S00 ppm
SOO
18 (M) 19 (F)
90-Day feeding study in dogs
Carbendazim formulation
0,100, SOO, IS00-2S00 ppm
100
90-Day feeding study in dogs
Carbendazim formulation
0, 100,300, or 1000 ppm (increased to 2000 after 6 weeks)
300
IS-Dose dermal study in rabbits
Benomyl formulation
0, 1000 mg/kg
NAa
IS-Dose dermal study in rabbits
Benomyl formulation
90-Day inhalation study in rats
Benomyl technical
0, 50, 2S0,SOO, 1000, SOOO mg/kg/day 0, 10, SO, 200 mg/m3
14-Day gavage study in rats
Benomyl technical
14-Day gavage study in rats
Glutathione-S-transferase induced at 3000 ppm (NOEL 1000 ppm)
aNA, not applicable,
histopathological abnormalities in any dogs fed test material and no other evidence of toxicity was observed. The NOEL was 500 ppm benomyl, approximately 18 mg/kg/day for male dogs and 19 mg/kg/day for females (Sherman, 1968). Dogs were also fed carbendazim in other 90-day feeding studies. In a study by Sherman (1970a) original dietary carbendazim concentrations of 0, 100, 500, and 2500 ppm were changed to a top dose of 1500 ppm because of reduced food intake at 2500 ppm. There was histological evidence of liver injury in dogs in the 1500-ppm group. The study NOEL was considered to be 500 ppm, a daily dosage equivalent to 14.4 mg/kg/day for males and 11.3 mg/kg/day for females. Til et al. (1972) fed dogs carbendazim at dietary levels of 0, 100, 300, and 1000 ppm for 13 weeks. The top dose was raised to 2000 ppm after 6 weeks of feeding. There were no clinical signs of toxicity or changes in body weight, hematological parameters, and kidney or liver function. Slight increases were seen in relative liver and thyroid weights with a decrease in
relative heart weights at the top dose, but histopathologic evaluation showed no compound-related changes. The NOEL was judged to be 300 ppm, equivalent to 7.5 mg/kg/day of carbendazim. Twenty-eight-day feeding studies were done to investigate hepatic toxicity and enzyme induction with benomyl or carbendazim in mice and rats of both sexes fed dietary concentrations of 0, 10, 30, 100, 300, 1000, and 3000 ppm. Rats were found to have increased liver weights. These effects were noted in animals fed carbendazim diets at 1000 and 3000 ppm and benomyl diets at 3000 ppm. Both benomyl and carbendazim induced hepatic epoxide hydrolase (EPH) in a dose-dependent fashion in mice and rats, with females of both species exhibiting a greater sensitivity to induction. Glutathione-s-transferase (GST) induction was qualitatively similar to EPH induction; however, the extent of induction was not as great. Significant induction occurred in rats and mice that received a dietary concentration of 3000 ppm (Guengerich, 1981).
1682
CHAPTER 77 Benzimidazoles
77.5.3.2 Dermal
77.5.3.3 Inhalation
The repeated-dose dermal toxicity of be no my I and carbendazim was studied in rabbits. Applications of a 50% benomyl formulation to either intact or abraded abdominal skin sites were made 5 days/week for 3 weeks at a dose rate of 1000 mg per kilogram of body weight. Slight edema, erythema, and atonia were noted at all skin sites as well as some desquamation. Compound-related changes in body weight or organ weights were not seen. Microscopic evaluation of the testes showed degenerative changes in the spermatogenic elements of the seminiferous tubules (Busey, 1968d). A more extensive study of the benomyl formulation was done by Hood (1969), applying doses of 0, 50, 250, 500, 1000, and 5000 mglkg to non occluded abraded dorsal skin sites for 6 hours per day, 5 days/week for 3 weeks. Decreased body weight gains and diarrhea, oliguria, and hematuria were seen in both males and females at the top two doses. Mild to moderate skin irritation was most notable at the top dose. Average testicular weights and testes-to-body-weight ratio changes were only noted at the 1000-mglkg dose. Histopathologic changes were not reported. Technical carbendazim in a 50% aqueous paste was applied to shaved, intact dorsal skin sites of rabbits daily, 6 hours/day, for 10 days at doses of 0 or 1000 mglkg. No adverse effects were reported on body weight, clinical signs, organ weights, gross pathology, or histopathology of selected organs. Focal epidermal necrosis and polymorphonuclear cell infiltration was noted in five ofthe six rabbits (Dashiell, 1975).
Effects observed in rats exposed to benomyl by inhalation were localized to the respiratory tract. A 90-day study was conducted in rats with a dust of benomyl technical. Groups of 20 male and 20 female rats were exposed, nose-only, 6 hours/day, 5 days/week. Particle sizes were within range of respirability for the rat. Male rats exposed to 200 mg/m 3 had low mean body weights and low food consumption. Male and female rats exposed to 200 mg/m 3 and male rats exposed to 50 mg/m 3 had degeneration of the olfactory epithelium. No other pathological abnormalities were observed. The study NOEL was 10 mg/m 3 , based on the respiratory tract toxicity (Warheit et aI., 1989). Hurtt et al. (1993) have utilized both respiratory and oral exposures of rats to demonstrate that the effects on the nasal epithelium are only seen when exposure is by inhalation.
77.5.4 CHRONIC TOXICITY Feeding studies were conducted in rats and mice to evaluate the chronic toxicity and carcinogenicity of benomyl and carbendazim. Dogs were also fed the respective chemicals to evaluate chronic toxicity in a nonrodent species (Table 77.4). The liver was the primary target for chronic toxicity, and a subchronic study conducted to address hepatic enzyme induction was previously described.
Table 77.4 Summary of Chronic Toxicity Studies Material
Dose
Study
tested
(ppm)
ppm
mg/kg/day
(ppm)
Reference
Two-year feeding study in rats
Benomyl (formulation)
0,0,100,500,2500
2500
109 (M) 128 (F)
2500
Sherman (1969c)
Two-year feeding study in rats
Carbendazim (formulation)
Two-year feeding study in mice
Benomyl (technical)
0,0,100,500,250010,000, 5000 0, 500, 1500, 50007500
Two-year feeding study in mice
Carbendazim (technical)
0, 500, 1500, 37507500
Two-year feeding study in NMRKfmice
Carbendazim
0, 50, 150, 300, or 5000
300
34.4 (M) 41.9 (F)
One-year feeding study in dogs
Carbendazim (technical)
0,100,200,500
200
6.4 (M) 7.2 (F)
Two-year feeding study in dogs
Carbendazim
0, 150,300, or 2000
300
Two-year feeding study in dogs
Carbendazim
0, 100,500, 1500
100
aNOEL set by peer review pathologists.
NOEL
500 None
500; nonea
21.6 (M) 24.5 (F) NA
81 (M) 125 (F)
2.6
NOAEL
500
Sherman (1972)
None
Weichman (1982) Peer review: Hardisty (1990); Frame and Van Pelt (1990)
500;
Wood (1982) Peer review: Hardisty (1990); Frame and Van Pelt (1990)
nonea
Same as NOEL
500 [16.5 mg/kg/day (M) and 17.1 mg/kg/day (F)]
Stadler (1986)
300
Reuzel et al. (1976)
100
Sherman (1972)
77.5 Mammalian Toxicity
77.5.4.1 Rats In long-term studies with male and female rats, no compoundrelated toxicity was observed with benomyl. With carbendazim, effects on body weight, blood, and liver were noted. There was no evidence of carcinogenicity with either compound in rats. Groups of 36 male and 36 female weanling rats were fed diets for 2 years that contained benomyl at dietary concentrations of 0, 100, SOO, and 2S00 ppm. During the study, there were no compound-related effects on mortality, body weight, food consumption, or clinical signs. There were no changes in blood or urine parameters, no organ weight effects at the 1- or 2-year sacrifice, and no histopathological lesions of note. The study NOEL was the highest dose tested, 2S00 ppm of benomyl, 109 mg per kilogram body weight per day for male rats and 128 mg per kilogram body weight per day for females (Lee, 1977; Sherman, 1969c). Carbendazim was fed to groups of 36 male and 36 female Sprague-Dawley rats in a study of similar design. Initial doses contained 0, 100, SOO, 2S00, or SOOO ppm of carbendazim. The dietary concentration for the 2S00-ppm group was raised to 7S00 ppm at 18 weeks, then raised again after two weeks to 10,000 ppm for the remainder of the study. In the study, male and female rats fed 2S00-1O,000-ppmand female rats fed SOOOppm diets had a low body weight gain. There were no effects on food consumption, food efficiency, clinical signs, or mortality. During the last 6 months of the study, female rats fed 10,000ppm diets were anemic. There were also increases in alkaline phosphatase and glutamic pyruvic transaminase (GPT) activities in high-dose rats. None of the histopathological changes were related to carbendazim intake. A review of tissues confirmed that there was no evidence of histopathological effects on the testes. The NOEL was SOO ppm, an average dose of 22 mg per kilogram body weight per day for male rats and 2S mg/kg/day for female rats (Sherman, 1972). The results in these studies were similar to those in another chronic study in which groups of 60 male and 60 female Wistar-derived rats were fed carbendazim diets at concentrations of 0, ISO, 300, and 2000 ppm for 2 years. The 2000-ppm concentration was increased to SOOO ppm after 1 week and to 10,000 ppm after 2 weeks for the rest of the study. Decreased body weights, and blood and liver effects occurred in rats fed the high-dose diets, but there were no carcinogenic or other effects from the dietary intake of carbendazim. Except for an increased incidence of diffuse proliferation of parafollicular cells of the thyroid in the high-dose females (Til et aI., 1976a). The NOEL was 300 mg/kg, a dosage equivalent to approximately IS mg/kg/day. Because the molar ratio (on a weight basis) of benomyl to carbendazim is approximately 1.S:1, it is not surprising that studies with carbendazim can result in a somewhat lower NOEL than studies with benomyl.
77.5.4.2 Mice In contrast to studies in rats, there was an increased incidence of benign liver tumors in long-term studies in mice with either
1683
benomyl or carbendazim. The data and slides from the original evaluation of liver tumors were subsequently peer-reviewed by a panel of pathologists, who concluded that there was in increase in benign but not malignant liver tumors. The mechanism of tumor induction appears to be an indirect effect that promotes the growth of spontaneous tumors in CD-1 mice rather than a direct genotoxic effect of the test compound. Groups of 80 male and 80 female CD-1 mice were fed diets that contained benomyl at concentrations of 0, SOO, lS00, and SOOO ppm for two years (the highest concentration was reduced from 7S00 to SOOO mg/kg after 37 weeks because the mice experienced marked reductions in body weight). The body weights of mice in the low- and intermediate-dosage groups were not affected. There were no effects in clinical signs, mortality, or hematological parameters; however, there were a number of significant findings from histopathological evaluation of animal tissues. Increased liver weights in the lS00- and SOOO-7500 mg/kg groups and decreased testicular weights in the 5000-7S00 mg/kg males were noted. Significant increases in the incidence of hepatocellular carcinomas and combined neoplasms were also reported initially at the SOO and lS00 but not 7500 ppm. It was noted, however, that there were no differences between control and treated groups with respect to time-to-tumor or histomorphological appearance of the tumors. There was not a no-effect level in the study (Weichman, 1982). In a similar 2-year study, groups of 80 male and 80 female CD-1 mice were fed diets containing carbendazim at concentrations of 0, 500, lS00, and 7S00 ppm for 2 years. Because of early mortality, the concentration of high-dose diets given to male mice was reduced from 7500 to 3750 ppm. Mortalities continued and the group was terminated at 73 weeks. Increased mortality also occurred among males fed lS00-ppm diets. Female mice continued to receive 7S00-ppm diets. There were no compound-related effects on clinical signs or hematological parameters. At necropsy, there were increased liver weights in females in the 7S00-ppm group. Notable histopathological changes were initially reported to include an increased incidence of nonneoplastic liver lesions as well as hepatocellular carcinomas, hepatocellular adenomas, and combined hepatocellular neoplasms. Tumor incidences were elevated in mice from all carbendazim dietary groups (Wood, 1982). A subsequent independent peer review of the slides and data for both the benomyl and the carbendazim mouse studies by three pathologists concluded that both compounds produced benign, but not malignant, hepatocellular neoplasms. There were also an increased incidence of multiple hepatocellular adenomas and a slight increase in the incidence of focus or foci of cellular alteration. A NOEL could not be established for histopathological lesions in either study (Frame and Van Pelt, 1990; Hardisty, 1990). There are two additional reports on long-term effects of carbendazim in the mouse. An 80-week study was conducted with groups of 100 male and 100 female Swiss (SBF) mice fed diets containing carbendazim at concentrations of 0, 150, 300, and SOOO ppm. Increased liver weights, an increased incidence of liver adenomas and carcinomas, and an increased incidence
1684
CHAPTER 77
Benzimidazoles
of liver nodules were present (Beems et aI., 1976; Mohr, 1977). Another study to confirm the findings of the first was conducted with HOE NMRKf (SPF-71) mice, a strain with a low incidence of spontaneous tumors. The dietary concentrations were 50, 150,300, and 5000 ppm. Although there histopathological effects were observed in livers of mice fed the 5000-mg/kg diets, there was no compound-related effect on the incidence or time of onset of tumors, and the total number of benign and malignant tumors was comparable among the groups of mice. There was no evidence of carcinogenicity at any dose tested (Donaubauer et aI., 1982). It appears that benomyl and carbendazim are not directly carcinogenic, but act to promote a common tumor type in CD-1 mice. It is likely that adaptive metabolic responses rather than a genotoxic insult cause the tumors. Several observations support this conclusion: (1) There is no evidence that benomyl or metabolites interact directly with DNA. (2) The absence of a decreased latency period indicates that benomyl is not acting as a classical tumor promoter. (3) There have been no significant differences identified between rats and mice in the metabolism of benomyl that would explain the different sensitivities to tumor formation. (4) The different results between the rat and mouse lifetime studies, and between the NMRKf mice and other strains, indicate that mice having high rates of spontaneous liver tumor formation are uniquely affected by benomyl or carbendazim. Viewed together, study results indicate that increased incidence of hepatic neoplasia in mice fed benomyl or carbendazim does not indicate risk of carcinogenicity in other mammalian species.
concentrations of serum cholesterol were slightly elevated in dogs in the 500-ppm group. One dog had a rare thyroid follicular cell adenoma, but the tumor was considered a chance finding. No histopathological lesions resulted from intake of carbendazim. The NOEL was 500 ppm, a mean daily intake of approximately 16.5 mg/kg/day for male dogs and approximately 17.1 mg/kg/day for females (Stadler, 1986). In an earlier study, dogs were fed 0, 100, 500, or 15002500 ppm of carbendazim in the diets for 2 years. Albumin concentration was decreased in dogs fed the high-concentration carbendazim diets. Evidence of hepatic cirrhosis and/or mild chronic toxic hepatitis was seen in dogs from the 500-ppm and 1500-2500-ppm carbendazim groups. These effects were attributed to compound intake. Testicular changes in two dogs that were fed 100 ppm were not attributed to carbendazim, because similar effects were not seen at 500 ppm. The NOEL was 100 mg/kg of carbendazim (Sherman, 1972). In another 2-year study with carbendazim, dogs received diets containing 0, 150, 300, or 2000-5000 ppm. Biochemical evidence of liver effects and increases in weights of livers, pituitary, and thyroid were noted in the high-dose animals. The NOEL was 300 ppm (Reuzel et aI., 1976). Although sporadic findings were noted in dogs fed benomy1 or carbendazim, the common target organ in these studies was the liver. The no-effect levels ranged from 100 to 500 ppm based primarily on biochemical markers of liver effects rather than overt toxicity due to test material. 77.5.5 NEUROTOXICITY
77.5.4.3 Dogs Dogs were fed benomy1or carbendazim to evaluate chronic toxicity in a nonrodent species. In all studies, evidence of liver injury was noted after extended intake of test material. Dogs (five per sex per group) were fed diets with carbendazim concentrations of 0, 100, 200, and 500 ppm in a I-year study. There were no compound-related effects on body weight, food consumption, clinical observations, or organ weights. The
Neurotoxicity studies have been conducted in hens and mammals (Table 77.5). Two rat neurotoxicity studies have been summarized (FAOIWHO, 1996a, b) and an acute study was reported earlier (De si, 1983). A number of earlier studies conducted in hens with benomyl or carbendazim were also reported (Goldenthal, 1978; Jessup, 1979; Jessup and Dean, 1979). In vitro studies by Belasco (1979a) and Krupka (1974) demon-
Table 77.5 Summary of Neurotoxicity Studies Material
Dose
Study
tested
levels
NOEL or NOAEL
References
Acute oral in rats (gavage)
Benomyl
0, 500, 1000, or 2000 mg/kg
Neurotoxicity >2000 mg/kg; other, none
Foss (1993)
90-Day oral in rats (feeding)
Benomyl
0, 100,2500, or 7500 ppm
Neurotoxicity> 7500 ppm; other 2500 ppm
Foss (1994)
Neurotoxicity in hens
Benomyl
500, 2500, or 5000 mg/kg
Neurotoxicity >5000 mglkg
Goldenthal (1978)
Acute delayed neurotoxicity in chickens
Benomyl
500, 2500, or 5000 mg/kg
Neurotoxicity >5000 mg/kg
Jessup (1979)
Neurotoxicity in hens
Carbendazim
500 or 2500 mg/kg
Neurotoxicity >2500 mg/kg
Jessup and Dean (1979)
77.5 Mammalian Toxicity
strated that benomyl did not inhibit either acetyl cholinesterase or butyryl cholinesterase. In an acute neurotoxicity study, rats were given single oral doses of benomyl at doses up to 2000 mg/kg. Doses of 500 mg/kg and above affected body weight and feed consumption. Motor activity was reduced in the female group dosed with 2000 mg/kg on the day of dosing; however, the reduced activity was associated with general toxicity and not with specific neurotoxicity. Neither the FOB parameters nor the neurohistological evaluation showed evidence of any neurotoxicity due to benomyl (Foss, 1993). In a 90-day study, rats exhibited reduced body weight and feed consumption and increased motor activity in the groups fed 7500 ppm in the diet. There were no specific neurotoxic end points identified for rats fed benomyl, because all FOB (functional observational battery) measurements and histopathological evaluations were similar to control parameters (Foss, 1994). The study NOEL was set at 2500 ppm, a dose equivalent to 126-266 mg/kg/day, based on end points of general toxicity. The NOEL for neurotoxicity resulting from administration of benomyl in the feed is 7500 ppm. The mammalian neurotoxicity studies indicate that benomyl is not a specific neruotoxicant. The significant effects on motor activity parameters (reduced in the acute study and increased in the subchronic study) appear to be related to general toxicity in animals that received high doses. In another rat study, no effects on electroencephalogram (EEG), behavior, or cholinesterase activity in the cerebellum, brain stem, white matter, or cerebral cortex were detected after daily oral doses of 250 or 500 mg/kg for 90 days (Desi, 1983). Similar conclusions were reached in early neruotoxicity studies conducted with benomyl and carbendazim in hens (Goldenthal, 1978; Jessup, 1979; Jessup and Dean, 1979). Transient ataxias in hens treated with benomyl were related to acute toxicity and not to any specific neurotoxic action. One benomyl study noted demyelination and axonal degeneration in all treated hens as well as in positive and negative control groups. These lesions were found to be due to Marek's disease, which was present in birds throughout the study. Similar ataxias were seen in hens treated with carbendazim. 77.5.6 GENOTOXICITY Benomyl and carbendazim have long been recognized for their ability to cause numerical chromosome aberrations (aneuploidy) both in vitro and in vivo because of their ability to bind tubulin and to disrupt microtubule assembly during cell division. They inhibit fungal growth by binding to tubulin (Davidse, 1975; Davidse and Flach, 1976b) and can also bind to mammalian tubulin (Albertini et aI., 1988; De Brabander et aI., 1976b) but with much less affinity (Albertini et aI., 1993; Ireland et aI., 1979; Russell et aI., 1992). Studies selected to show the range of testing done are presented in Tables 77 .6-77 .8. These substances do not interact directly with DNA and have given negative results for gene mutations, structural chromosome aberrations (clastogenicity), and DNA damage and repair
1685
in the majority of studies conducted (Bentley et aI., 2000). Many studies using various test systems have been conducted to define the mutagenic potential of these chemicals both in vitro and in vivo, including studies in nonmammalian cells and in mammalian somatic and germ cells. This has resulted in the predictable situation of having apparently conflicting results in some cases. WHO expert panels reviewed the available studies, both public and proprietary, for both compounds in 1993 (WHO, 1993a, b). They provided a service for those interested in this data by tabulating those studies that presented "sufficient detail to evaluate the reasons for the conflicting data". A similar presentation is available in the FAOIWHO (1996a, b) JMPR publications. One of the reasons for some of the conflicting data was the presence of phenazine contaminants [2-amino-3-hydroxyphenazine (AHP) and 2,3-diaminophenazine (DAP)] in some of the carbendazim test samples produced prior to the mid-1980s. These contaminants were shown to be quite effective in inducing gene mutations in the Salmonella typhimurium Ames test whereas pure carbendazim or benomyl was negative. These contaminants are mutagenic at very low concentrations in the Ames test. Concentrations of > 5 ppm DAP and 10 ppm AHP in carbendazim test samples resulted in positive test results (Sarrif et aI., 1994a). Since that time, most of the carbendazim manufacturers have changed the process to reduce the level of these impurities to below 0.5 ppm for DAP and 3.5 ppm for AHP, and technical carbendazim has been demonstrated to be nonmutagenic in subsequent Ames tests. These contaminants are not produced when benomyl is metabolized to carbendazim. Cytotoxicity and, with benomyl, toxicity resulting from the release of the n-butylcarbamoyl moiety also contributed to the conflicting results (Sarrif et aI., 1994a). Toxicity relating to the n-butylcarbamoyl moiety is likely to be responsible for the reported positives in the mouse lymphoma assay with benomyl because no such activity is reported in similar assays using purified carbendazim. In two in vitro studies evaluating numerical chromosome aberrations, structural chromosome aberrations were observed at benomyl concentrations greater than those causing aneuploidy or polypoloidy. This observation supports the view that the clastogenic effects produced by benomyl resulted from cytotoxicity due to the n-butylcarbamoyl moiety rather than direct interaction with DNA. However, the positive clastogenic finding in these studies with benomyl appears to have little significance in vivo because structural chromosome aberrations have not been observed in animal studies (Sarrif et al.,1994b). A number of additional studies have been reported since the WHO reviews in 1993 and the FAOIWHO reviews in 1995. Those using newer techniques of antikinetichore antibodies and fluorescence in situ hybridization (FISH) have provided conclusive evidence of the lack of direct interaction with DNA and the existence of thresholds for the aneugenic effects of benomyl and carbendazim (Bentley et aI., 2000; Bjorge et aI., 1996; Elhajouji et aI., 1995; Elhajouji et aI., 1997; Jeffay et aI., 1996; Mailhes and Aardema, 1992; Sarrif et aI., 1994b; WHO, 1993a, b).
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CHAPTER 77 Benzimidazoles
Table 77.6 Summary of In Vitro Genotoxicity Studies Test
Material
Concentration
Test
organism
tested
range
Result
Reference
Salmonella-Ames assay
Salmonella typhimurium
Benomyl (technical)
0-J0,000 iJ-g/plate
Negative
Russell (1978a, b); Rickard (1986);
2-Amino-3-hydroxyphenazine
0-0.1 iJ-g/plate (with S9)
Positive :0::0.02 iJ-g/plate
Arce (1984a)
2,3-Phenazinediamine
0-0.1 iJ-g/plate (with S9)
Positive :0::0.01 iJ-g/plate
Arce (1984b)
CHOfHPRT assay
Chinese hamster ovary (CHO) cells
Benomyl (technical)
0-172 iJ-M (with S9) 0-120 iJ-M (without S9)
Negative
Fitzpatrick (1980)
Mouse lymphoma assay
L5178Y cells
Benomyl (technical)
0-25 iJ-M (with S9) 0-15 iJ-M (without S9)
Positive at 10 and 12.5 iJ-M (without S9)
McCooey et al. (l983a)
Carbendazim (purified technical)
0-200 iJ-M (with & without S9)
Negative
McCooey et al. (1983b)
Rat hepatocytes
Carbendazim (technical)
0-125 iJ-glml (654iJ-M)
Negative
Tong (1981b)
Mouse hepatocytes
Benomyl (technical)
0-500 iJ-glml (1722 j.lM)
Negative
Tong (l98Ia)
Mouse hepatocytes
Carbendazim (technical)
0-125 iJ-g/ml (654 iJ-M)
Negative
Tong (198Jc)
Unscheduled DNA synthesis
Note: Concentrations given in parentheses are the iJ-M equivalent of the concentrations given in the original report.
As noted previously, inhibition of microtubular function by benomyl and carbendazim results from the interaction with a non-DNA target (tubulin). Therefore, it is reasonable to expect that a "threshold" can be demonstrated for the aneugenic effect (Dellarco et aI., 1985; Parry et aI., 1994) where a critical number of target sites must be occupied before aneuploidy is expressed. Studies designed to demonstrate the presence of thresholds utilizing new and more sensitive techniques for the detection of aneuploidy at low concentrations have been reported. Elhajouji et al. (1995), Marshall et al. (1996), and
Bentley et al. (2000) have employed fluorescent DNA probes and fluorescence in situ hybridization techniques to demonstrate clear thresholds for benomyl and carbendazim in cultured human lymphocytes. Bentley et al. (2000) evaluated the threshold for both benomyl- and carbendazim-induced aneuploidy using cultured binucleate human lymphocytes to track numerical chromosome abnormalities for six chromosomes. The results showed that there were similar dose responses for each chromosome and equimolar threshold concentrations for benomyl and carbendazim.
Table 77.7 Summary of In Vitro Chromosome Aberration Studies Test
Material
Concentration
organism
tested
range
Result
NOEL
Reference
R3-5 human-mouse hybrid cell line
Benomyl (technical)
0, 1.5,3.0,7.5, and 15 iJ-glml
Aneuploidy Polyploidy Clastogenicity
Effect at all concs. 1.5 iJ-glml 7.5 iJ-g/ml
Athwal and Sandhu (1985); Sandhu et al. (1988)
Chinese hamster ovary (CHO) cells
Benomyl (technical)
0, 1.4,2.8,5.7,11.3, and 22.7 iJ-g/ml (without S9)
Polyploidy Clastogenicity
1.4 iJ-g/ml 11.3 iJ-g/ml
Sasaki (1988)
0,3.1,6.2,12.5,22.7, 49.9, and 90.6 iJ-g/ml (with S9)
Polyploidy Clastogenicity
Not induced 22.7 iJ-glml
77.5 Mammalian Toxicity
1687
Table 77.8 Summary of In Vivo Genotoxicity Studies Test
Material
species
tested
Dose levels
Result
Reference
Bone marrow chromosome aberrations
Mouse
Benomyl (technical)
0,625, 1250, 2500, 5000 mglkg by gavage
Negative
Stabl (1990)
Bone marrow micronucleus assay
Mouse
Benomyl (technical)
0, 1250, 2500, 5000 mg/kg by gavage
Positive :::2500 mg/kg
Sasaki (1990)
Mouse
Benomyl (technical)
0, 100, 2500, 5000 mg/kg by gavage
Positive :::2500 mglkg
Bentley (1992a)
Mouse
Carbendazim (technical)
0, 66, 1646, 3293 mg/kg by gavage
Positive :::1646 mg/kg
Bentley (I 992b )
Rat
Benomyl (formulation)
0, 500, 2500, 5000 ppm in feed 7 days
Negative
Culik (1974)
Test
Dominant lethal assay
Some investigators have reported possible increases in chromosomal aberrations (Carbonell et aI., 1990) and sister chromatid exchanges (Lander and Ronne, 1995) in workers exposed to multiple pesticides. Others have found either lack of genotoxicity (Dolara et aI., 1993) or possible effects (Dolara et aI., 1994) from such exposures. Benomyl and carbendazim are often included in these types of studies because of their widespread use. However, there is no conclusive evidence of cytogenetic effects in humans from exposure to these two compounds. The weight of extensive available evidence is that benomyl and carbendazim do not directly interact with cellular DNA and do not induce either point mutations or chromosome aberrations. This has been demonstrated in both mammalian and nonmammalian systems in vitro and in vivo, and in somatic cells as well as in germ cells (Bentley et aI., 2000; FAOIWHO, 1996a, b; Sarrif et aI., 1994b; WHO, 1993a, b).
77.5.7 REPRODUCTIVE AND DEVELOPMENTAL TOXICITY Because of their mechanism of action on the dividing cell, these compounds have been thoroughly studied in various studies of reproductive and developmental toxicity (Tables 77.9-77.12). In early testing of benomyl, it was learned that high doses by gavage could cause testicular effects. Biotransformation studies showed the rapid conversion of benomyl to carbendazim so many studies were conducted with carbendazim rather than benomyl. It is clear that, despite extensive study, benomyl and carbendazim continue to hold interest for those investigating reproductive and developmental effects of chemicals. It was also recognized early in testing of benomyl that teratogenic effects seen with bolus (gavage) dosing in rats were not seen in feeding studies with either rats or rabbits. This is important in assessment of risk of health effects in humans, be-
Table 77.9 Studies of Reproductive Effects via Dietary Exposure Dosing Species
Sex
duration
Doses given
Results and comments
Reference
Chemical
Rat, SpragueDawley
MIF
2 generations
0, 100, 500, 3000, or 10,000 ppm of diet
Decreased body weights in parents and fewer, smaller pups at 10,000; spenn count decreased with testicular changes at 3000 and 10,000 ppm; no effect on mating or fertility indexes
Mebus (1990)
Benomyl
Rat, SpragueDawley
MIF
3 generations
0, 100, 500, 5000, or 10,000 ppm of diet
Lower avg. litter weights at 5000 and 10,000; no effects on fertility, gestation, lactation, or pup viability
Shennan (1972)
Carbendazim
Rat, Wistar
MIF
3 generations
0, 150, 300, or 2000 ppm of diet
No reproductive or teratogenic effects at up to 2000 ppm of diet
Koeter et al. (1976a, b)
Carbendazim
Rat, Wistar
M
70 days on and 70 days off feed
0, 1,6.3 or 203 mg/kg
Decreased spenn count, testicle weight; reversal of all effects at end of recovery phase
Bames et al. (1983)
Benomyl
1688
CHAPTER 77
Benzimidazoles
Table 77.10 Studies of Developmental Toxicity Effects via Dietary Exposure Dosing Species
durationa
Doses given
Results and comments
Rat, SpragueDawley
g.d.6-15
0,9,44,210, or 373 mg/kg body weight
No significant maternal or fetal effects
Rat, Wistar
g.d.7-16
0, 169, 298, or 505 mg/kg body weight
Decreased fetal weight at top dose but no major malformations or anomalies
Rat, SpragueDawley
g.d.6-15
0,9,46,218,432,or 626 mg/kg body weight
No adverse maternal or fetal effects
Rat, Wistar
g.d.6-15
0, 600, 2000, or 6000 ppm of diet
Maternal body weight decrease at top dose; no evidence of teratogenicity
Koeter (l975a)
Carbemdazim
Rabbit, New Zealand
g.d.8-16
0, 100, or 500 ppm of diet
No developmental toxicity but inadequate pups or litters for full evaluation
Busey (I 968c )
Benomyl
Rabbit, New Zealand
g.d.6-18
0, 600, 2000, or 6000 ppm of diet
Delayed ossification and increased supernumerar ribs and skull bones at top dose; inadequate data for full evaluation
Koeter (l975b)
Carbendazim
ag.d.
Reference
Chemical Benomyl
Kavlock et al. (1982)
Benomyl
Carbendazim
= gestation day(s).
cause it is unlikely that anyone will unintentionally receive a large oral dose of these fungicides. 77.5.7.1 Reproductive Toxicity Studies
Reproduction studies of varying designs have been conducted in rats, mice, and hamsters. Comprehensive multigeneration feeding studies have been reported with both benomyl and carbendazim. Gavage dosing and, in one case, intraperitoneal and intratesticular injection have also been used (Lim and Miller, 1997). Feeding Studies of Reproductive Toxicity Reproduction feeding studies are summarized in Table 77.9. A two-generation reproduction study in Sprague-Dawley rats was done to recent international guidelines with benomyl at dietary concentrations of 0, 100,500,3000, and 10,000 ppm (Mebus, 1990). Parental rats were fed the treated diets for 71 days before breeding to produce the FI offspring. Selected FI rats were fed diets with the same concentrations of benomyl for at least 105 days after weaning and were mated to produce the F2a generation. The FI females were mated again, to non sibling males, to produce the F2b litters. Indices of reproductive function assessed for the FO and FI adults included mating, fertility, gestation, viability, lactation, percentage of pups born alive, and percentage litter survival. Clinical observations were recorded and body weight parameters, food consumption, and food efficiency were measured. Parental rats were killed after litter production and subjected to gross and microscopic pathology evaluation. Selected weanlings were also examined grossly. Effects were observed in both parents and offspring fed the two highest dietary concentrations. No increase in parental mortality was seen at any dose; however, body weight pa-
rameters and overall food consumption by FO and FI rats were decreased at 10,000 ppm. The number of F2a and F2b pups alive before culling on day 4 was decreased, and the offspring of this dose group had low body weights at birth. Offspring of rats treated with 3000 ppm of benomyl also had decreased body weights at some evaluations during lactation. Decreased testicular weight, histopathologic changes, and depressed sperm counts were seen in FO and FI male rats at 3000 and 10,000 ppm. Histopathologic changes included atrophy and degeneration of the seminiferous tubules in the testes of rats at 3000 and 10,000 ppm and oligospermia in the epididymides of FO rats at 10,000 ppm and of FI rats at 3000 and 10,000 ppm. There were no compound-related differences in mating indices, fertility indices, or gestation length. The NOEL for adults and offspring was 500 ppm, equivalent to about 37 mg/kg/day. A three-generation feeding study was conducted in SpragueDawley rats with carbendazim at doses of 0, 100, 500, 5000, and 10,000 ppm of diet (Sherman, 1972). Parental rats were fed the diets for approximately 80 days and mated to produce the FI generation. The number of matings, of pregnancies, and of pups per litter at birth were recorded. Litter body weights and number alive were recorded during lactation. The parents were again mated to produce the FIb litters. Subsequent matings produced F2a, F2b, F3a, and F3b litters. Selected rats from the control, and the 5000- and 1O,000-ppm groups were examined for gross and histologic lesions. No carbendazim-related effect was apparent on fertility, gestation, viability, or lactation. Average litter weights at weaning were decreased in all generations fed 5000 and 10,000 ppm. The NOEL was 500 ppm, the same as found in the two-generation study with benomyl. Another three-generation feeding study was conducted with Wistar rats, at carbendazim dietary levels of 0, 150, 300, and
77,5 Mammalian Toxicity
1689
Table 77,11 Studies of Reproductive Effects via Gavage Exposure Dosing Results and comments
Reference
Chemical
Killed at various days after treatment ended; no treatment-related effects
Carter (1982)
Benomyl
0, 200, or 400 mg/kg
Done to evaluate sperm effects; found effects on all stages at 400mg/kg
Carter and Laskey (1982)
Benomyl
62 days
0, 1,5, 15, or 45 mg/kg
No hormonal or reproductive effects; testicular effects at 45
Linder et al. (1988)
Benomyl
M
Single dose
0,25,50,100,200,400, or 800mg/kg
Germ cell sloughing at 100 and higher; no long-term effects at 25 and 50
Hess et al. (1991)
Benomyl
Rat, SpragueDawley
MIF
10 days
400 mg/kg
Serial breeding protocol, 75% infertile at 5th week; 50% still infertile at 32 weeks
Carter et al. (1987)
Carbendazim
Rat
MIF
I generation
0, 50, 100, 200, or 400 mg/kg
Reproductive effects at 200 and 400; sperm effects at 50 and above
Gray et al. (1988, 1990)
Carbendazim
Hamster
MlfF
1 generation
0, 50, 100, 200, or 400 mg/kg
No reproductive effects; testicular and sperm effects at 400
Gray et al. (1988, 1990)
Carbendazim
Rat, Holtzman
F
g,d, 1-8a
0,25,50,100,200,400, or 1000 mg/kg
Toxicity at 1000 with decreased serum LH (luteinizing hormone) and increased serum estradiol
Carbendazim
Rat
F
8 days
Induced pseudopregnancy; reported reduced uterine competency
Carbendazim
Mouse
M
5 days
0, 250, 500, or 1000 mglkg
Hamster
F
Single dose
0, 250, 500, 750, or 1000 mglkg; then or 1000 in follow-up study
Rat, SpragueDawley
M
Single dose
Rat, SpragueDawley
M
Single dose
Dog
M
Rat, SpragueDawley
M
Species
Sex
duration
Doses given
Rat, SpragueDawley
M
10 days
°
Rat, SpragueDawley
M
10 days
Rat, Wistar
M
Rat, SpragueDawley
Hamster
or 200 mglkg
°or 400 mg/kg
Sperm quality study; found decreased testicular weight and sperm effects at 1000
Evanson et al. (1987)
Carbendazim
Reduced pregnancy rate and number of live pups at 750 and 1000; early pregnancy loss at 1000 given during microtubule sensitive meiotic stage
Perrault et al. (1992)
Carbendazim
Time and testicular effects; saw various adverse effects on sperm and testicles
Nakai et al. (1992)
Carbendazim
0,50,100,200,400, or 800mg/kg
Follow-up time and testicular effects; found dose-dependent adverse effects on spermatids and seminiferous tubules at doses of 100 and above
Nakai et al. (1992)
Carbendazim
4-hr inhalation
0,065 and 1.65 mg/l
Reduced spermatogonic activity at 14 days but not 28 days postexposure at 1.65 mgll
Littlefield and Busey (1969)
Benomyl
Single dose
100 mg/kg
Electromagnetic and light microscopy for study of morphologic changes in spermatids; concluded that abnormalities seen are in common with those reported in testes with several chemicals and in mutant animals
Nakai et al. (1997)
Carbendazim
Single dose
1000 mglkg
Direct oocyte effect, not endocrine effect; oocyte aneuploidy results in early pregnancy loss
Jeffay et al. (1996)
Carbendazim
°
°or 400 mg/kg
(continues)
1690
CHAPTER 77
Benzimidazoles
Table 77.11 (continued) Dosing Species
Sex
duration
Doses given
Results and comments
Reference
Chemical
Rat, SpragueDawley
M
Single dose
100mglkg
Found rapid direct effects on spermatids aside from effects caused by tubular blockage from sloughing
Naki and Hess (1997)
Carbendazim
Rat, SpragueDawley
M
ip and testicular injectiona
ip, 250 (ben) and 164 (mbc) mglkga
Compared benomyl and mbc effects; concluded that all of benomyl-related testicular damage and microtubuleassembly disruption could be explained by the mbc contribution
Lim and Miller (1997)
Benomyl and carbendazim
No malformations, only reproductive effects
Piersma et al. (1995)
Benomyl
testicular injection, 400 (ben) and 262 (mbc) flg/kg Rat, Wistar
ag.d.
F
OECD screen protocol
0, 10, 30, 90 mg/kg
= gestation day(s); ip = intraperitoneal; ben = benomyl; mbc = carbendazim.
2000 ppm (Koeter et aI., 1976a, b). In a slightly different study design, the Fla and F2a litters were discarded at weaning and the FIb and F2b litters used to produce succeeding generations. F3a pups were used for a teratology study and the F3b pups were used in a 4-week toxicology study. Body weight gain was unaffected, but all generations in the treated groups weighed more than the controls. There was no effect on fertility, survival, litter size, or lactation. Necropsy of rats from the 4-week toxicity study showed increased relative liver weights and decreased relative spleen weights in females fed 2000 ppm as well as decreased relative ovarian weights in all treated groups. Histopathologic evaluation of the livers failed to show any compound-related changes. There were no apparent carbendazim-related effects on reproduction and no teratogenic effects at doses of carbendazim of up to 2000 ppm of the diet. A feeding study was done in male Wistar rats to assess effects on testicular function and recovery after cessation of exposure to benomyl (Barues et aI., 1983). The treated diets provided doses of approximately 0, 1, 6, and 203 ppm. The test groups were split into two groups each to be killed after 70 days of feeding or after 70 days of recovery. Decreased relative testicular weights and slightly lower fertility index were found in all treated groups, with ejaculated sperm counts down in the top dose group. Copulatory behavior was unaffected and dominant lethal mutation was not induced by treatment with benomyl. Plasma testosterone and gonadotropin levels were unchanged. Studies of the recovery group showed that all changes had reversed.
Gavage Studies of Reproductive Toxicity Numerous studies ranging from single doses to daily dosing for one generation have been done with rats, mice, and hamsters to investigate various aspects of reproductive function (Table 77.11). A single dose of 100 mg/kg or more is genera~ly required for any effect to be seen with either benomyl or carbendazim. Multiple doses similarly are generally without effect at doses lower
than 100 mg/kg/day, although Gray et al. (1988) reported some effects on sperm at 50 mg/kg and above of carbendazim in a one-generation reproduction study. A similar study done in hamsters by the same group found no effects at up to 400 mg/kg. Male reproductive effects of benomyl and carbendazim in rats have been extensively studied. Naki and Hess (1997) found that carbendazim has a direct effect on spermatids, in addition to the effects caused by tubular blockage from sloughing of spermatids. Lim and Miller (1997) used intraperitoneal and direct testicular injection of benomyl and carbendazim to compare the degree of effects at measured testicular concentrations of both. They concluded that all of the testicular damage and microtubule disruption seen after treatment with benomyl can be explained by the carbendazim released from the benomyl rather than benomyl itself. Effects on female reproductive organs have also been studied. Although numerous authors have listed benomyl with other pesticides in suggesting that it may have endocrine activity, none of the studies conducted to date have shown direct or indirect evidence of hormonal activity. Spencer et al. (1996) showed that benomyl and carbendazim affected uterine weight without producing effects on hormone activity. Jeffay et al. (1996) concluded that the effects seen on oocytes is direct and not mediated by endocrine changes. The aneugenic effect on the oocyte results in early pregnancy loss.
77.5.7.2 Developmental Toxicity Studies The developmental toxicity of benomyl and carbendazim has been shown to be similar in the rather extensive testing which has been reported in rabbits, rats, mice, and hamsters. There appears to be no contribution from the n-buty1carbamoyl moiety to the effects noted.
Feeding Studies of Developmental Toxicity (See Table 77.10.) Sherman et al. (1970a) fed carbendazim to rats during
77.5 Mammalian Toxicity
1691
Table 77.12 Studies of Developmental Toxicity Effects via Gavage Exposure
Species
Dosing duration
Doses given
Results and comments
Reference
Chemical
Mouse, CD-l
g.d.7-17a
0, 50, 100, or 200 mg/kg
No maternal toxicity; increased fetal abnormalities at 100 and 200mg/kg
Kavlock et at. (1982)
Benomyl
Rat, SpragueDawley
g.d.7-16
0, 3, 10, 30, 62.5, or 125 mg/kg
No maternal toxicity; decreased fetal weight and increased mortality at 125 mg/kg, increased skeletal variations at 62.5; microphthpalmia seen in few fetuses
Staples (1980)
Benomyl
Rat, SpragueDawley
g.d.7-16
0, 3, 6.25, 10, 20, 30, or 62.5 mg/kg
Follow-up to better define microphthalmia; only 2 fetuses with malformations, both in top dose group, I with microphthalmia; no teratogenicity seen at lower doses
Staples (1982)
Benomyl
Rat, Wistar
g.d.7-16
0, 15.6, 31.2, 62.5, or 125 mg/kg
Lower maternal body weight at top dose, some anomalies at top 2 doses, including microphthalmia
Kavlock et at. (1982)
Benomyl
Rat, Wistar
g.d. 7 to lactation day 15
0, 15.6, or 31,2 mg/kg
No effects at 100 days after birth except for reduction in testes and seminal vesicle weights at top dose
Kavlock et at. (1982)
Benomyl
Rat
g.d.7-21
0, 31.2, or 62.4 mg/kg
Ocular and craniocerebral malformations at both doses with protein-deficient diet
Zeman et at. (1986)
Benomyl
Rat
g.d.7-21
0, 31.2, or 62.4 mg/kg
Ocular and craniocerebral malformations at both doses with protein-deficient diet
Hoogenboom et at. (1991)
Benomyl
Rat, Holtzman
g.d.I-8
0, 100, 200, 400, or 600mg/kg
Killed at g.d. 11 or 20; no maternal toxicity; saw developmental delays and embryo or fetal toxicity at 200 and above, decreased fetal weight, litter size, and delayed ossification at all doses
Cummings et at. (1992)
Carbendazim
Rat, SpragueDawley
g.d.7-16 (guideline study)
0,5, 10, 20, or 90 mg/kg
Maternal toxicity at top dose; malformations at top dose; fetal toxicity at 20 and 90; NOEL 20 (maternal) and 10 (fetal)
Alvarez (1987)
Carbendazim
Rat, SpragueDawley
g.d.8-15
19.1 mg/kg
11 rats, one dose; reported fetal toxicity and malformations
Delatour and Besse (1990)
Carbendazim
Rat, Wistar
g.d.6-15
0, 20, 40, or 80 mg/kg
No malformations; embryo or fetal toxicity at two top doses
lanardhan et at. (1984)
Carbendazim
Rabbit
g.d.8-16
0, 40, 80, or 160 mg/kg
No malformations; dose-related increase in fetal toxicity
lanardhan et at. (1984)
Carbendazim
Rabbit, New Zealand
g.d.7-19 (guideline study)
0,10,20, or 125 mg/kg
Maternal toxicity at top dose; malformations at top dose; fetal toxicity at 20 and 125; NOEL 20 (maternal) and 10 (fetal)
Christian et al. (1985)
Carbendazim
Rabbit, New Zealand
g.d.7-28 (guideline study)
0, 15, 30,90, or 180mg/kg
Maternal toxicity at top dose; fetal toxicity and variations at top dose; NOEL 90 mg/kg for both maternal and fetal toxicity
Munley (1995)
Benomyl
Hamster
Single dose, g.d. 10
0, 15, 30,75, or 150 mg/kg
NOEL 30 mg/kg; malformations, embryotoxicity, resorptions at higher doses
Minata and Biernacki (1982)
Carbendazim
(continues)
1692
CHAPTER 77
Benzimidazoles
Table 77.12 (continued)
Species
Dosing duration
Doses given
Results and comments
Reference
Chemical
Rat, SpragueDawley
g.d.6-15
0, 10,30,60, 100,300, 1000, or 3000 mglkg
Malformations at 60 and 100, all resorbed (early) at 300, 1000, and 3000; NOEL 10 mg/kg
Hofmann and Peh (1987a, b)
Carbendazim
Rat, Wistar
14 days prior to mating to postnatal day 6
10, 30, or 90 mg/kg
No malformations, reproductive effects only
Piersma et al. (1995)
Benomyl
Rat, Wistar
g.d.6-15
90 and 270 mg/kg
Malformations and reproductive effects
Piersma et al. (1995)
Benomyl
Rat, Holtzman
g.d.I-8
0':100 mg/kg
Embryotoxicity, decreased growth, malformations
Cummings et al. (1992)
Carbendazim
Rat, Wistar
g.d.6-15
0, 40, 60, 80, 100 mg/kg
NOEL 20 mg/kg; all embryos dead at 100; malformations at 40, 60, and 80 mg/kg
Lu et al. (1995)
Carbendazim
ag.d.
= gestation day(s).
gestation days 6-15 at dietary concentrations providing doses of approximately 0,9,46,218,432,626, and 747 mg/kg/day. No mortality, effects on body weight, or clinical signs of toxicity were seen. Food intake was reduced at the top dose during the period of dosing but returned to control afterward. There was no reduction in implantation sites, increase in resorptions, or differences in live to dead fetuses. A repeat of this study design using benomyl at dietary concentrations providing doses of approximately 0,9,44,210, and 373 mg/kg/day gave results identical to those of the earlier carbendazim study, with the exception that three litters in the top dose group had some fetuses with hydronephrosis and retarded ossification. A later benomyl study of similar design fed rats dietary concentrations providing daily doses of 0, 169, 298, or 505 mg/kg/day on gestation days 7-16. A decrease in fetal body weight was seen at the top dose, but no major malformations or abnormalities were noted (Kavlock et aI., 1982). An early feeding study in rabbits fed benomyl at dietary concentrations of 0, 100, and 500 ppm during gestation days 8-16 showed no evidence of developmental toxicity or malformations. However, there were inadequate numbers of litters or pups available for examination (Busey, 1968c) to fully evaluate this study. Carbendazim was fed to rabbits in a later study at dietary concentrations of 0, 600, 2000, and 6000 ppm during gestation days 6-18 (Koeter, 1975b). A significant increase was found in the number of supernumerary ribs and skull bones at the top dose and ossification was delayed or absent in these fetuses. Some study design deficiencies and lack of adequate data presented prevent the clear understanding of the developmental toxicity potential of carbendazim (WHO, 1993b).
Gavage Studies of Developmental Toxicity (See Table 77 .12.) Studies utilizing gavage administration were done in rabbits, mice, rats, and hamsters with benomyl or carbendazim. It appears that, when such bolus exposure of the dam is utilized, the fetus is sometimes somewhat more sensitive than is the dam. However, the comprehensive studies that have been done allow
one to identify a NOEL for the fetus which may be used in risk assessment. Additionally, in the most recent study of benomyl in rabbits done to the current expanded testing protocol no fetal effects were shown at doses below those causing maternal effects. In a standard Segment 11 study, benomyl was given to Sprague-Dawley rats by gavage at doses of 0, 3, 10, 30, 62.5, and 125 mg/kg/day on gestation days 7-10. Test animals were observed daily for clinical signs of toxicity or changes in behavior. No clinical signs of toxicity were seen in any dose group. Body weight gain and incidences of pregnancy, corpora lutea, implantation sites, and sex ratios were unchanged compared to controls. Fetal body weight was decreased at 62.5 and125 mg/kg and there was an increase in fetal mortality at 125 mg/kg. Malformations noted included microphthalmia, anophthalmia, and hydrocephaly. These appeared possibly treatment related at the higher dose levels. Histopathologic evaluation of the fetal eyes showed irregularities in fetuses from the two top doses. Major skeletal malformations were seen at the top dose and other skeletal variations were increased at both doses (Staples, 1980). A follow-up study was done to better determine a NOEL for the microphthalmia and hydrocephaly (Staples, 1982). Groups of rats from the same strain and the same supplier were given doses of 0, 3, 6.25, 10, 20, 30, and 62.5 mg/kg from gestation days 7-16. Following a gross pathologic evaluation, reproductive status was determined on a litter basis. The number of implantation sites, resorptions and dead, live, and stunted fetuses as well as the mean weight of live fetuses per litter were determined. Microscopic examination was done of fetal heads. At the top dose, mean fetal body weight was decreased. Two fetuses had malformations; both were from the 62.5-mg/kg dose group and from different litters. One had unilateral microphthalmia and the other showed internal hydrocephaly. No malformations were seen at the lower doses. Kavlock et al. (1982) evaluated the teratologic potential of Wistar rats given daily doses of 0, 15.6, 31.2, 62.5, or
77.5 Mammalian Toxicity 125 mg/kg during gestation days 7-16. Maternal toxicity was seen in decreased body weight at 125 mg/kg. Developmental toxicity was evident as increased incidence of fetal abnormalities at the top two doses. These included microphthalmia, hydrocephaly, encephaloceles, fused vertebrae, and fused ribs. The lower two doses were without adverse effect on the fetus. Kavlock et at. (1982) also evaluated the postnatal effects of benomyl on pups. Dams were given benomyl at daily doses of 0,15.6, or 31.2 mg/kg from gestation day 7 to lactation day 15 and pups were evaluated periodically for 100 days after birth. At 100 days of age, organs weighed included adrenals, liver, kidney, ovaries, testes, and ventral prostate plus seminal vesicles. Organ weights were not affected by benomyl exposure except that weight of the testes and the ventral prostate and seminal vesicles was reduced at the top dose. There were no compound-related effects on litter size at birth or weaning, fetal body weight, growth, survival, or locomotor activity. Additionally, Kavlock et at. (1982) studied the effects of benomyl given to CD-l mice at doses of 0, 50, 100, and 200 mg/kg on gestation days 7-17. Pups were delivered (by Cesarean section) on day 18 and the number of live, dead, and resorbed fetuses determined. Fetuses were examined grossly for abnormalities. Maternal toxicity was not noted at any dose. Fetal developmental toxicity appeared to be shown at all doses by a significant dose-response trend for increased incidence of subnormal vertebrae and supernumerary ribs as well as a significant increase in enlarged renal pelvises in pups from the top-dose dams. The number of abnormal litters and fetuses was also significantly increased at the top two doses and fetal weights were decreased. Zeman et at. (1986) and Hoogenboom et at. (1991) studied the effects of benomyl on Sprague-Dawley rats given doses of 31.2 and 62.4 ml/kg on gestation days 7-21. Ocular abnormalities were seen in 43% of the fetuses from the top dose group. These abnormalities increased in incidence to 63% when the dams were fed a protein-deficient diet and 62.4 mg/kg of benomyl. Craniocerebral abnormalities (primarily hydrocephaly) were found in fetuses from dams given protein-deficient diets plus 31.2 mg/kg of benomyl. In a standard Segment 11 study, Sprague-Dawley rats were given daily doses of 0, 5, 10, 20, or 90 mg/kg of carbendazim on gestation days 7-16. Maternal toxicity was seen at the top dose as decreased body weight, increased mean liver weight, and decreased pregnancy rate. Fetal body weight was reduced at the top two doses and malformations were increased in incidence at the top dose. The malformations included hydrocephaly, microphthalmia, anophthalmia, malformed scapulae, and axial skeletal malformations (vertebral, rib, and sternal fusions, exencephaly, hemivertebrae, and rib hyperplasia). The NOEL was found to be 20 mg/kg for the dams and 10 mg/kg for the fetus (Alvarez, 1987). Cummings et at. (1992) dosed Holtzman rats daily at rates of 0, 100, 200, 400, and 600 mg/kg of carbendazim on gestation days 1-8. Dams were killed on day 11 or 20. Maternal toxicity was not noted. Crown-rump length, head length, number of somites, and number of embryos per dam were reduced at
1693
doses of 200 mg/kg and greater at day 11. Increased resorptions, decreased live litter size and decrease fetal body weight, and delayed ossification were seen at day 20 in all treated groups. Delatour and Besse (1990) dosed 11 Sprague-Dawley rats at 19.1 mg/kg on gestation days 1-15. They reported an increase in dead fetuses and fetal skeletal and external malformations with a decrease in fetal body weight in the treated group. Janardhan et at. (1984) studied rats and rabbits to determine effects of carbendazim on fetal survival and development. Wistar rats were given daily doses of 0, 20, 40, or 80 mg/kg on gestation days 6-15. Half of the animals were killed on day 21 and half were allowed to have normal deliveries. Rabbits were dosed at 0, 40, 80, or 160 mg/kg on gestation days 6-18 and killed on day 31. There was a dose-related increase in dead and resorbed fetuses in rats killed at day 21 and in rabbits. There was no effect on mean fetal weight and no malformations were seen. In rats allowed to deliver, there appeared to be a doserelated reduction in pups per litter and mean fetal weight was increased in the top two dose groups. Mortality at 21 days postpartum was 3-3.5 times greater at these doses also as compared with controls. Christian et at. (1985) studied the developmental toxicity of carbendazim in a typical Segment 11 study. New Zealand rabbits were given daily doses of 0, 10, 20, or 125 mg/kg on gestation days 7-19. Maternal toxicity and fetal malformations were noted at the top dose. Fetal toxicity and variations were noted at 20 and 125 mg/kg. The NOEL was 20 for the dam and 10 for the fetus. Munley (1995) did a more recent, contemporary, expanded Segment 11 study of New Zealand rabbits dosed with benomyl at doses of 0, 15, 30, and 180 mg/kg on gestation days 7-28. Maternal toxicity and fetal toxicity and variations were seen at the top dose. The NOEL for both maternal and fetal toxicity was determined to be 90 mg/kg. Hamsters were given a single dose of carbendazim at 0, 15, 30, 75, or 150 mg/kg on gestation day 10. An increase in malformations and embryotoxicity was seen at 75 and 150 mg/kg (Hofmann and Peh, 1987a, b). Piersma et al. (1995) used benomyl in a proposed Organization for Economic Cooperation and Development (OECD) screening protocol for reproductive toxicology. Male and female Wistar rats were gavaged daily at doses of 0, 10, 30, and 90 mg/kg from 14 days prior to mating onward. Males were killed after 28 days of exposure and females were killed after 4-6 days of lactation, approximately 54 days of exposure. Adults and pups were necropsied and evaluated. Male rats showed the expected testicular degeneration after 28 days at the top dose. Dams in the high-dose group had decreased body weights, and pup weights were decreased as well. However, there was no increase in pup mortality and, unexpectedly, no malformations were found. The investigators then proceeded to conduct a typical developmental toxicity study, giving dams daily doses of 0, 90, or 270 mg/kg of be no my1 during gestation days 6-15 for comparison with the earlier results. Body weight gain was reduced at both 90 and 270 mg/kg. Postimplantation loss and total litter
1694
CHAPTER 77
Benzimidazoles
resorptions were increased in both groups as well. Ophthalmic abnormalities were increased in a dose-related fashion. The authors note that the developmental toxicity study has 10 exposure days whereas the OECD protocol exposes the females for 54 days. They speculate that changes in maternal metabolism during the premating phase exposure may somehow protect against fetal malformations. Lu et al. (1995) have reported exposure of female rats to carbendazim at daily doses of 0, 20, 40, 60, 80, or 100 mg/kg during gestation days 6-15. Complete embryo mortality was seen at the top dose, with increased malformations at 40, 60, and 80 mg/kg. The dams in all treated groups showed reduction in body weight gains and liver weights were increased in all but the 20-mg/kg group. The NOEL would appear to be 20 mg/kg for both dam and fetus.
REFERENCES Unpublished references that are dealt with in WHO (1993a) are marked with an asterisk (*). Those dealt with in WHO (I 993b) are marked with a dagger (t). Unpublished references that are dealt with in FAOIWHO (l996a) are Those dealt with in FAOIWHO (1996b) marked with a double dagger are marked with a section symbol (§). Albertini, S., Brunner, M., and Wurgler, F. E. (1993). Analysis of the six additional chemicals for in vitro assays of the European Economic Communities' EEC Aneuploidy Programme using Saccharomyces cerevisiae D61. M and the in vitro porcine brain tubulin assembly assay. Environ. Mol. Mutagen. 21, 180-192. Albertini, S., Friedreich, U., Holderegger, c., and Wurgler, F. E. (1988). The in vitro porcine brain tubulin and assembly assay: Effects of a genotoxic carcinogen (aflatoxin Bl), eight tumor promotors and nine miscellaneous substances. Mutat. Res. 201, 283-292. t Alvarez, L. (1987). "Teratogenicity Study ofINE-965 (carbendazim) in Rats." Unpublished report HLR 281-87, E. 1. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. t Anderson (1999). t Arce, G. T. (1984a). "Mutagenicity Evaluation in Salmonella typhimurium." Unpublished report HLR 72-84, E. 1. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. t Arce, G. T. (1984b). "Mutagenicity Evaluation in Salmonella typhimurium." Unpublished report HLR 71-84, E. 1. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. Athwal, R S., and Sandhu S. S. (1985). Use of human x mouse hybrid cell line to detect aneuploidy induced by environmental chemicals. Mutat. Res. 149, 73-81. Barnes, T. B., Veriangieri, A l., and Wilson, M. C. (1983). Reproductive toxicity of methyl-l-(butylcarbamoyl)-2-benzimidazole carbamate (benomyl) in male Wistar rats. Toxicology 28, 103-115. tBeems, R B., Til, H. P., and van der Heijden, C. A (1976). "Carcinogenicity Study with Carbendazim (99% MBC) in Mice." Central Institute for Nutrition and Food Research (TNO), The Hague. (Unpublished summary report A08129, Hoechst AG, Frankfurt, and BASF AG, Ludwigshafen, Germany.) Belasco, 1. J. (1969). Belasco, 1. l. (1970). *Belasco, 1. J. (I 979a). "Study Showing the Absence of Acetylcholinesterase Inhibition with a Wettable Powder Formulation (50% Benomyl)." Unpublished report BrrOX 4, E. I. du Pont de Nemours and Co., Inc., Wilmington, DE. *Belasco, I. J. (l979b). "2- 14 C-Benomyl (50 WP) Adsorption through Rat Skin. Part n. Effect of Time and Dose Applied, with Supplement." Unpublished report BIME 47 and supplement HLR 117-79, E. I. du Pont de Nemours and Co., Inc., Wilmington, DE.
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(TNO), The Hague. (Unpublished report A04120, Hoechst AG, Frankfurt, and BASF AG, Ludwigshafen, Gennany.) tKoeter, H. B. W. M. (1975b). "Effect of HOE 17411F (=BAS 3460F) on Pregnancy of the New Zealand White Rabbit." Central Institute for Nutrition and Food Research (TNO), The Hague. (Unpublished report A09804, Hoechst AG, Frankfurt, and BASF AG, Ludwigshafen, Gennany.) tKoeter, H. B. W. M., et al. (1976a). tKoeter, H. B. W. M., et al. (1976b). tKramer, M., and Weigand, W. (1971). "(HOE 17411 O.E) Toxicological Examination." Unpublished report A00936, Hoechst AG, Pharmaceuticals Research, Toxicology Section, Frankfurt, Gennany. Krechniak, J., and Klosowska, B. (1986). The fate of 14-C-carbendazim in rat. Xenobiotica 16(9), 809-815. Kristensen, P., and Irgens, L. (1994). Clusters of anophthalmia. Br. Med. J. 206, 308-309. Krupka, R M. (1974). On the anti-cholinesterase activity of benomyl. Pestic. Sci. 5, 2[[-216. Kuehne, G., Heise, H., Plottke, B., and Puskeiler, T. (1985). Dennatitis after Benlate contact. Z. Gesamte. Hyg., Grenzgeb. 31,710-711. Lander, E, and Ronne, M. (1995). Frequency of sister chromatid exchange and hematological effects in pesticide-exposed greenhouse sprayers. Scand. J. Work Environ. Health 21, 283-288. Larsen, A. I., Larsen, A., Jepsen, J. R, and Jorgensen, R (1990). Contact allergy to the fungicide benomyl? Contact Dermatitis 22,278-281. Lavy, T. L., Mattice, J. D., Massey, J. H., and Sku1man, B. W. (1993). Measurements of year-long exposure to tree nursery workers using multiple pesticides. Arch. Environ. Contam. Toxicol. 24, 123-144. *Lee, K P. (1977). "The Two-Year Feeding Study in Rats with Benomy1 with Supplemental Pathology Report." Unpublished report HLR 66-77, E.I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. Leonard, J. A., and Yeary, R A. (1990). Exposure of workers using hand-held equipment during urban application of pesticides to trees and ornamental shrubs. Am. Ind. Hyg. Assoc. J. 50, 605-609. Liesivuori, J., Liukkonen, S., and Pirhonen, P. (1988). Reentry intervals after pesticide application in greenhouses. Scand. J. Work Environ. Health 14(Suppl. 1),35-36. Lim, J., and Miller, M. (1997). The role of the benomyl metabolite carbendazim in benomyl-induced testicular toxicity. Toxicol. Appl. Pharmacol.142, 401410. Linder, RE., Rehnberg, G. L., Strader, L. E, and Diggs, J. P. (1988). Evaluation of reproductive parameters in adult male Wistar rats after subchronic exposure. J. Toxicol. Environ. Health 25, 285-298. Lisi, P., Caraffini, S., and Assalve, D. (1986). A test series for pesticide dermatitis. Contact Dermatitis 15, 266-269. *Littlefield, N. A., and Busey, W. M. (1969). "Four-Hour Acute Inhalation Exposure Test in Dogs Using a Wettable Powder Fonnulation (50% Benomyl)." Hazelton Laboratories, Inc., Falls Church, VA. (Unpublished report HLR 192-69, E. I. du Pont de Nemours and Co., Inc., Wilmington, DE.) Lu, S. Y., Hong-Wei, L., and Shun-Cheng, W. (1995). Taiwan Agricultural Chemicals and Toxic Substance Research Institute. Plant Protection Bulletin (Taichung) 37(3), 331-338. Mailhes, J. B., and Aardema, M. J. (1992). Benomyl-induced aneuploidy in mouse oocytes. Mutagenesis 7, 303-309. Marshall, R R, Murphy, M., Kirkland, D. J., and Bentley, K S. (1996). Fluorescence in situ hybridisation with chromosome-specific centromeric probes: A sensitive method to detect aneuploidy. Mutat. Res. 372, 233-245. tMartin, D. A., Henry, J. E., and Brock, W. J. (1987). "Closed-Patch Repeated Insult Dennal Sensitization Study (Buehler Method) with Benlate C Fungicide in Guinea-Pigs." Unpublished report HLR 510-87, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. Matsushita, T., and Aoyama, K (1981). Cross reactions between some pesticides and the fungicide benomyl in contact allergy. Ind. Health. 19, 77-83. Matsushita, T., Yoshioka, M., Aoyama, K, Aritmatsu, Y., and Nomura, S. (1977). Experimental study on contact dennatitis caused by fungicides benomyl and thiophanate-methyl. Ind. Health 15, 141. *McCooey, K. T., Arce, G. T., Sarrif, A. M., and Krahn, D. E (1983a). "L5178Y Mouse Lymphoma Cell Assay for Mutagenicity." Unpublished
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report HLR 86-83, E. I. du Pont de Nemours and Company, Haskell Laboratory, Newark, DE. *McCooey, K T., Arce, G. T., Sarrif, A. M., and Krahn, D. E (1983b). "L5178Y Mouse Lymphoma Cell Assay for Mutagenicity." Unpublished report HLR 253-83, E. I. du Pont de Nemours and Company, Haskell Laboratory, Newark, DE. *Mebus, C. A. (1990). "Reproductive and Fertility Effects with DPX-1991529 (Benomyl). Multigeneration Reproduction Study in Rats." Unpublished report HLR 765-90, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. Minata, M., and Biernacki, B. (1982). Embryotoxicity of carbendazim in rats, rabbits and hamsters. Bull. Vet. Inst. Pulaway 2, 42-52. tMohr, U. (1977). "Review of Liver Sections from Mice and Rats Fed with Carbendazim." Department of Experimental Pathology, Medical School, Hannover, Gennany. (Unpublished report MBCrrOX 6 from Hoechst AG, Frankfurt, and BASF AG, Ludwigshafen, Gennany.) *Monson, K D. (1990). "Metabolism of [phenyl(U)_I4CjCarbendazim in Rats." Unpublished report AMR [[41-88, E. I. du Pont de Nemours and Co., Inc., Wilmington, DE. +Munley, S. M. (1995). "Developmental Toxicity Study of DPX-Tl991-529 (Benomyl)." Unpublished report HLR 164-95, E. I. du Pont de Nemours and Co., Haskell Laboratory, Newark, DE. Nakai, M., Hess, R. A., Moore, B. J., Guttroff, R E, Strader, L. E, and Linder, R E. (1992). Acute and long-tenn effects of the fungicide carbendazim (methyI2-benzimidazole carbamate; mbc) on the male reproductive system in the rat. J. Androl. 13, 507-518. Nakai, M., et al. (1997). Naki, M., and Hess, R A. (1997). Effects of carbendazim (methyl 2-benzimidazole carbamate; mbc) on meiotic spennatocytes and subsequent spermatogenesis in the rat testis. Anat. Rec. 247, 379-387. Naki, M., Hess, R. A., Matsuo, E, Gotoh, Y., and Nasu, T. (1996). Further observations on carbendazim-induced abnonnalities of spennatid morphology in rats. Tissue Cell 477-485. tNash, S. D., and Ferenz, R (1982). "Inhalation Median Lethal Concentration (LC50) in Rats-EPA Protocol (Carbendazim 75% Wettable Powder)." Unpublished report HLR 365-82, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. NIOSH (1994). "Ornamental Plant Nurseries, Florida." NIOSH Hazard Evaluation and Technical Assistance Report No. 92-0381-2445. Parry, J. M., Fielder, R. J., and McDonald, A. (1994). Thresholds for aneuploidy-inducing chemicals. Mutagenesis 9, 503-504. Perrault, S. D., Jeffay, S., Poss, P., and Laskey, J. W. (1992). Use of the fungicide carbendazim as a model compound to detennine the impact of acute chemical exposure during oocyte maturation and fertilization on pregnancy outcome in the hamster. Toxicol. Appl. Pharmacol. 114,225-231. Piersma, A. H., Verhoef, A., and Dortant, P. M. (1995). Evaluation of the OECD 421 Reproductive Toxicity Screening Test Protocol using 1-(butylcarbamoyl)-2-benzimidazole carbamate (benomyl). Teratog. Carcinog., Mutagen. 15,93-100. tReuzel, P. G. J., Hendriksen, C. E M., and Til, H. P. (1976). "Long-Tenn (TwoYear) Toxicity Study with Carbendazim in Beagle Dogs." Central Institute for Nutrition and Food Research (TNO), The Hague. (Unpublished report A06583, BASF AG, Ludwigshafen, and Hoechst AG, Frankfurt, Gennany.) *Rickard, L. B. (1983a). "Mutagenicity Evaluation in Salmonella typhimurium." Unpublished report HLR 97-83, E.I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. *Rickard, L. B. (1983b). "Mutagenicity Evaluation in Salmonella typhimurium." Unpublished report HLR 98-83, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. *Rickard, L. B. (1986). *Russell, J. E (1978a). "Mutagenic Activity of2-Benzimidazolecarbamic Acid, I(Butylcarbamoyl)-methyl Ester in the SalmonellalMicrosome Assay." Unpublished report HLR 18-78, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. *Russell, J. E (1978b). "Mutagenic Activity of 2-Benzimidazolecarbamic Acid, I(Butylcarbamoyl)-methyl Ester in the SalmonellaIMicrosome Assay." Un-
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Benzimidazoles
published report HLR 31-78, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. *Russell, J. F., et a!. (1992). §Russell, L. D. (1992). "Review of Prof. Hilscher's Report 'Effects of Carbendazim on Spermatogenesis'." Unpublished report, University of Illinois. (Unpublished report A49229, Hoechst AG, Frankfurt.) Sandhu, S. S., Gudi, R. D., and Athwal, R. S. (1988). A monochromosomal hybrid cell assay for evaluating the genotoxicity of environmental chemicals. Cell. BioI. Toxicol. 4,495-506. Sarrif, A. M., Arce, G. T., Krahn, D. F., O'Neil, R. M., and Reynolds, V. L. (l994a). Evaluation of carbendazim for gene mutations in the Salmonella/Ames plate-incorporation assay: The role of aminophenazine impurities. Mutat. Res. 321, 43-56. Sarrif, A. M., Bentley, K. S., Fu, L. l., O'Neil, R. M., Reynolds, V. L., and Stahl, R. G. (1994b). Evaluation of be nomy I and carbendazim in the in vivo aneuploidy/micronucleus assay in BDFI mouse bone marrow. Mutat. Res. 310, 143-149. tSarver, J. W. (1975). "Acute Inhalation Toxicity-One Hour Head Only (Carbendazim)." Unpublished report HLR 58-75, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. Sasaki, Y. F. X. (1988). Benomyl in-vitro cytogenetics test. IET-88-0043. The Institute of Environmental Toxicology, Kodiara Laboratories, Suzuki-cho 2-772, Kodaira, Tokyo 187, Japan. * Sasaki, Y. F. X. (1990). "Benomyl: Micronucleus Test in Mice." Kodaira Laboratories, Institute of Environmental Toxicology, Tokyo (Unpublished report IET 89-0046, E. I. du Pont de Nemours and Co., Inc., Wilmington, DE.) Savitt, L. E. (1972). Contact dermatitis due to benomyl (letter). Arch. Dermatol. 105,926-927. Schuman, S. H., and Dobson, R. L. (1985). An outbreak of contact dermatitis in farm workers. J. Am. Acad. Dermatol. 13,220-223. tSherman, H. (1965). "Acute Oral Test." Unpublished report HLR 125-65, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. Sherman, H. (1967). *Sherman, H. (1968). "Three-Month Feeding Study in Dogs Using a Wettable Powder Formulation (50% Benomyl)." Unpublished report HLR 269-68, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. *Sherman, H. (l969a). "Acute Oral LD50, Test in Rats Using Technical Benomyl (>95% Benomyl) and a Wettable Powder Formulation (50% Benomyl)." Unpublished report HLR 17-69, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. *Sherman, H. (l969b). "Acute Oral ALD Test in a Dog Using Technical Benomy I (>95% Benomyl)." Unpublished report HLR 168-69, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. *Sherman, H. (1969c). "Long Term Feeding Study in Rats with I-Butylcarbamoyl-2-benzimidazolecarbamic Acid, Methyl Ester (INT-1991)." Unpublished report HLR 232-69, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. tSherman, H. (l970a). "Three-Month Feeding Study in Dogs Using a Wettable Powder Formulation (50% MBC)." Unpublished report HLR 283-70, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. *Sherman, H. (I 970b). "Long-Term Feeding Study in Dogs with I-Butycarbamoyl-2-benzimidazolecarbamic Acid, Methyl Ester (INT-199 I)." Unpublished report HLR 48-70, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. *. tSherman, H. (1972). "Long-Term Feeding Studies in Rats and Dogs with 2-Benzimadazolecarbamic Acid, Methyl Ester (INE-965) (50% and 70% MBC Wettable Powder Formulations): Parts I and n." Unpublished report HLR 195-72, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. *Sherman, H., and Krauss, W. C. (1966). "Acute Oral Test [Benomyll." Unpublished report HLR 100-66, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. *Sherman, H., Barnes, J. R., and Krauss, W. C. (1967). "Ninety-Day Feeding Study with I-Butylcarbamoyl-2-benzimidazolecarbamic Acid, Methyl Ester (INT-1991)." Unpublished report HLR 11-67, E. I. du Pont de Nemours and Co., Inc., Haskell, Laboratory, Newark, DE.
Sherman, H., Culik, R., and Jackson, R. A. (1975). Reproduction, teratogenic and mutagenic studies with benomyl. Toxicol. Appl. Pharmacol. 32, 305315. Sherman, H. et a!. (1970a). Spagnolo, A., Bianchi, F., Calabro, A., Calzolari, E., Clementi, M., Mastroiacovo, P., Meli, P., Petrelli, G., and Tenconi, R. (1994). Anophthalmia and benomyl in Italy: A multicenter study based on 940,615 newboms. Reprod. Toxicol. 8, 397-403. Spencer, F., Chi, L., and Zhu, M. (1996). Effect of benomyl and carbendazim on steroid and molecular mechanisms in uterine decidual growth in rats. J. Appl. Toxicol. 16,211-214. §Stadler, J. C. (1986). "One-Year Feeding Study in Dogs with Carbendazim." Unpublished report HLR 291-86, E. I. du Pont de Nemours and Co .. Inc., Haskell Laboratory, Newark, DE. *Stahl, R. G., Jr. (1990). "In Vivo" Evaluation of INT-1991-259 for Chromosome Aberrations in Mouse Bone Marrow." Unpublished report HLR 401-90, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. *Staples, R. E. (1980). "Teratogenicity Study in the Rat after Administration by Gavage of Technical Benomyl (>95% Benomyl): Parts I, Il and Ill." Unpublished report HLR 649-80, E. I. du Pont de Nemours and Co .. Inc., Haskell Laboratory, Newark, DE. * Staples, R. E. (1982). "Teratogenicity Study in the Rat Using Technical Benomyl (>95% Benomyl) Administered by Gavage and Supplement with Individual Animal Data." Unpublished report HLR 587-82, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. Til, H. P., Beems, R. B., and Grout, A. P. (1981). Determination of the acute oral toxicity of carbendazim in mice. The Hauge, Central Institute for Nutrition and Food Research (TNO), Unpublished report No. A-21604, prepared for BASFAG, Ludwigshafen, Germany. tTil, H. P., Koellen, c., and van der Heijden, C. A. (l976a). "Combined Chronic Toxicity and Carcinogenicity Study with Carbendazim in Rats." Central Institute for Nutrition and Food Research (TNO), The Hague. (Unpublished report A08128, BASF AG, Ludwigshafen, and Hoechst AG, Frankfurt, Germany.) tTil, H. P., Koeter, H. B. W. M., and van der Heijden, C. A. (I 976b ). "Multigeneration Study with Carbendazim in Rats." Central Institute for Nutrition and Food Research (TNO), The Hague. (Unpublished report AI0295, Hoechst AG, Frankfurt.) tTil, H. P., van den Muelen, H. C., Feron, V. J., Seinen, W., and de Groot, A. P. (1972). "Sub-chronic (90 Day) Toxicity Study with W 17411 in Beagle Dogs." Central Institute for Nutrition and Food Research (TNO), The Hague. (Unpublished report A00292, Hoechst AG, Frankfurt.) *Tong, C. (l981a). "Hepatocyte Primary CulturelDNA Repair Assay on Compound 10,962-02 (Benomyl) Using Mouse Hepatocytes in Culture." Naylor Dana Institute, Valhalla, NY. (Unpublished report HLO 741-81 from E. I. du Pout de Nemours and Co., Inc., Wilmington, DE.) tTong, c. (l98Ib). "The Hepatocyte Primary CulturelDNA Repair Assay on Compound 11,201-01 Using Rat Hepatocytes in Culture." Naylor Dana Institute, Valhalla, NY. (Unpublished report HLO 743-81, E. I. du Pont de Nemours, Inc., Wilmington, DE.) tTong, c. (l98Ic). "The Hepatocyte Primary CulturelDNA Repair Assay on Compound 11,201-01 Using Mouse Hepatocytes in Culture." Naylor Dana Institute, Valhalla, NY. (Unpublished report HLO 744-81, E. I. du Pont de Nemours, Inc., Wilmington, DE.) *Turney, R. T. (1979). "Rat Inhalation Study-Benlate." Unpublished report HLR 116-79, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. Van Joost, T. H. et al.(1983). "Benomyl." Van Joost, T. H., Naafs, B., and van Ketel, W. G. (1983). Sensitization to benomy I and related pesticides. Contact Dermatitis 9, 153-154. Van Ketel, W. G. (1976). Sensitivity to the pesticide benomyl. Contact Dermatitis 2, 290-291. *Vick, D. A., and Brock, W. J. (l987a). "Primary Dermal Irritation Study with Benlate 50 DF Fungicide in Rabbits." Unpublished report HLR 300-87, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE.
References
tYick, D. A., and Brock, W. J. (l987b). "Primary Dermal Irritation Study with Benlate C Fungicide in Rabbits." Unpublished report HLR 299-87, E.I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. tYick, D. A., and Brock, W. J. (l987c). "Acute Dermal Toxicity Study of Benlate C Fungicide in Rabbits." Unpublished report HLR 303-87, E.I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. Yick, D. A., and Valentine, R. (1987). "Primary Eye Irritation Study with Benlate C Fungicide in Rabbits." Unpublished report HLR 287 -87, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. *Ward, R. S., and Scott, R. C. (1992). "Benomyl: In Vitro Absorption of a 500 g kg-I WP Formulation through Human Epidermis." Imperial Chemical Industries (ICI), Fernhurst, Naslemer, Surrey, UK. (Unpublished report CTLIP/3659, E. I. du Pont de Nemours and Co., Inc., Wilmington, DE.) Warheit, D. B., Kelly, D. P., Carakostas, M. C., and Singer, A. W. (1989). A 90day inhalation toxicity study with benomyl in rats. Fundam. Appl. Toxieol. 12, 333-345. *, +Weichman, B. E. (1982). "Long Term Feeding Study with Methyl 1-(Butylcarbamoyl)-2-benzimidazolecarbamate in Mice (INT-1991; >95% Benomyl): Parts I, Hand rn." Unpublished report HLR 20-82, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE.
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tWood, C. K. (1982). "Long-Term Feeding Study with 2-Benzimidazolecarbamate, Methyl Ester «99% MBC, INE-965) in Mice. Parts I and H." Unpublished report HLR 70-82, E. I. du Pont de Nemours and Co., Inc., Haskell Laboratory, Newark, DE. World Health Organization (WHO) (l993a). "Environmental Health Criteria. 148. Benomyl." World Health Organization, Geneva. World Health Organization (WHO) (l993b). "Environmental Health Criteria. 149. Carbendazim." World Health Organization, Geneva. Zelesco, P. A., Barbieri, I., and Graves, J. A. M. (1990). Use of a cell hybrid test system to demonstrate that benomyl induces aneuploidy and polyploidy. Mutat. Res. 242, 329-335. Zeman, F. J., Hoogenboom, E. R., Kavlock, R. J., and Semple, J. L. (1986). Effects on the fetus of maternal benomyl exposure in the protein-derived rat. J. Toxieol. Environ. Health 17,405-417. Zweig, G., Gao, R., and Popendorf, W. (1983). Simultaneous dermal exposure to captan and benomyl by strawberry harvesters. J. Agrie. Food Chem. 31, 1109-1113.
CHAPTER
78 Cyprodinil: A Fungicide of the Anilinopyrimidine Class Felix Waechter, Edgar Weber, and Thomas Hertner Syngenta Crop Protection AG
78.1 INTRODUCTION
78.2.2 SYNONYMS
Anilinopyrimidines are a new chemical class of fungicides that are highly active against a broad range of fungi. Three compounds are currently on the market: Cyprodinil (Syngenta Crop Protection AG), Mepanipyrim (Kumiai Chemical Industry Co., Ltd.), and Pyrimethanil (AgrEvo GmbH). They all have similar structures and differ only with respect to their substituent at position 4 of the pyrimidine ring, which is a cyclopropyl group for Cyprodinil, a 1-propynyl group for Mepanipyrim, and a methyl group for Pyrimethanil (Fig. 78.1). The biological mode of action includes inhibition of methionine biosynthesis and secretion of hydrolytic enzymes. Anilinopyrimidines show no cross-resistance with other fungicide groups. Extensive field monitoring has shown reliable performance of anilinopyrirnidine products over many years and no cases of practical resistance have been detected under commercial conditions in all key target pathogens including Botrytis and Venturia. However, under conditions of forced selection, a potential resistance risk of these pathogens could be demonstrated in field and laboratory trials. This chapter addresses the toxicological profile and human safety aspects of Cyprodinil as a representative example of anilinopyrimidines. Where published data are available, reference is made to the toxicological profile of the other two marketed anilinopyrimidines, Mepanipyrim and Pyrimethanil.
78.2 IDENTITY, PROPERTIES, AND USES 78.2.1 CHEMICAL NAME Cyprodinil (Fig. 78.1) is (4-cyclopropyl-6-methyl-pyrirnidin2-yl)-phenyl-amine (according to IUPAC) or 4-cyclopropyl-6methyl-N -phenyl-2-pyrirnidinamine (according to CA). Handbook of Pesticide Toxicology Volume 2. Agents
The common name Cyprodinil (ISO draft) is in general use. Trade names include Unix® (Syngenta), Chorus® (Syngenta), Stereo® (Syngenta), Switch® (Syngenta), and Vangard™ (Syngenta USA). The CAS registry number is 121552-61-2 and the developmental code is CGA 219417. 78.2.3 PHYSICAL AND CHEMICAL PROPERTIES Cyprodinil has the molecular formula C14H15N3 and a molecular weight of 225.3. It is a white crystalline solid with a melting point of 75.9°C. The vapor pressure of Cyprodinil at 25°C is 5.1 x 10-4 Pa (crystal modification A) or4.7 x 10- 4 Pa (crystal modification B). Its solubility in water at 25°C is 20 mg/liter (pH 5.0), 13 mg/liter (pH 7.0), or 15 mg/liter (pH 9.0). It is quite soluble in most organic solvents. The partition coefficient log Pow (n-octanollwater) is 3.9 at pH 5.0 or 4.0 at pH 7.0 and pH 9.0 (Heye et aI., 1994). 78.2.4 HISTORY, FORMULATIONS, AND USES Ciba introduced Cyprodinil in France in 1993 for application on cereal grains. It is currently used as a foliar fungicide in cereal grains, grapes, pome fruit, stone fruit, strawberries, vegetables, field crops, and omamentals, and as a seed dressing on barley. Among others, products that contain Cyprodinil are marketed in most European countries, in North America, and in Japan. Cyprodinil controls a wide range of pathogens such as Pseudocercosporella herpotrichoides, Erysiphe spp., Pyrenophora teres, Rhynchosporium secalis, Septoria nodorum, Botrytis spp., Alternaria spp., Venturia spp., and Monilinia spp. Formulations for foliar applications are of the WG (water dispersible molecules) or EC (emulsifiable concentrates) type; those for seed dressing are of the FS (flowable concentrate for
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Copyright © 200 I by Academic Press. All rights of reproduction in any fonn reserved.
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CHAPTER 78
Cyprodinil
Cyprodinii
Mepanipyrim
Pyrimethanii
Figure 78.1 Pyrimethanil.
Chemical
structures of Cyprodinil,
Mepanipyrim,
and
seed treatment) type, Important mixing partners are triazoles, fludioxonil, fenpropidin, and acibenzolar-S-methyl.
(only phenyl labelled compound) and 100 mg/kg. Additional animals of both sexes received 14 consecutive doses of the nonlabelled test article followed by one single treatment with the phenyl labelled compound at 0.5 mg/kg. The biliary excretion was investigated in bile-fistulated males treated with 100 mg/kg of the phenyl-labelled test article. Independent of dose level and sex, Cyprodinil was rapidly absorbed from the gastrointestinal tract into systemic circulation. The excretion pattern was essentially independent of the sex, dose level, pretreatment, or label of the test substance administered. Approximately 48-68% of the administered dose was excreted in the urine, whereas 29-47% was found in the feces. Total excretion reached 92-97% of the administered dose within 48 h. In bile-fistulated rats, 39, 35, and 14% of the radioactivity were excreted via bile, urine, and feces, respectively. A comparison of these values with those obtained with the nonfistulated animals suggests that a significant amount eliminated via the feces was absorbed and reentered the intestinal tract by biliary excretion. A smaller amount seemed to be reabsorbed from the intestine and eliminated via the kidneys. Residues in tissues were generally low and rapidly depleted with no evidence for an accumulation or retention of radioactivity. 78.4.1.2 Dermal Administration
78.3 BIOLOGICAL MODE OF ACTION The anilinopyrimidine fungicides Cyprodinil, Pyrimethanil, and Mepanipyrim were identified as inhibitors of methionine biosynthesis in liquid cultures of Botrytis cinerea, In filamentous fungi, such as Neurospora grassa and Aspergillus nidulans, the cystathionine pathway has been established as the major route of homocysteine and methionine biosynthesis (Yamagata, 1989). Within this pathway, cystathionine ,B-Iyase, which catalyzes the synthesis of homocysteine from cystathionine, was identified as a target for these anilinopyrimidines (Fritz et aI., 1997; Masner et aI., 1994). The plant cell-wall degrading enzyme secretion system of pathogenic fungi was identified as a second target for anilinopyrimidine fungicides. Cell-wall digesting enzymes of pathogenic fungi are thought to play a critical role in the early step of plant infection, that is, in the absence of secretion, penetration of plant tissue by fungal mycelium is inhibited (Milling and Richardson, 1995; Miura et aI., 1994).
78.4 ABSORPTION, DISTRIBUTION, METABOLISM, AND EXCRETION 78.4.1 TOXICOKINETICS IN RATS 78.4.1.1 Oral Administration
The distribution and depletion kinetics in male and female rats were investigated with [U- 14 C]phenyl-cyprodinil and [U- 14 C]pyrimidyl-cyprodinil. The labelled test substance was administered to male and female rats at single oral doses of 0.5
The dermal absorption of Cyprodinil formulated as a WG-type product was determined in rats. The compound was applied for 6 h at two dose levels: the lower dose represented a typical spray dilution; the higher dose represented the highest applicable concentration. Residues in blood, skin, urine, and feces were determined during 48 h. Blood levels reached a maximum within 1-2 hours of exposure and decreased rapidly. The total amount absorbed within 48 h was roughly 20% of the administered low dose and about 3% of the high dose. 78.4.2 METABOLIC PATHWAYS IN RATS
The metabolic fate of Cyprodinil was investigated in urine, feces, and bile samples obtained from the toxicokinetic investigations. In urine and feces, there were no differences in the metabolite patterns of the phenyl and pyrimidyllabels, that is, there were no metabolites identified in these excreta that indicated a cleavage of the C-N -C bridge between the phenyl and pyrimidyl ring. Qualitatively, the metabolic pathways were not influenced by the dose level, by pretreatment with unlabelled Cyprodinil, or by the sex of the animals. Seven urinary, two biliary, and two fecal metabolites were identified, which in total accounted for 65-80% of the administered radioactivity. Cyprodinil was almost completely metabolized. No unchanged parent molecule could be found in urine, whereas minor amounts of unchanged Cyprodinil were found in feces. Most of the administered Cyprodinil was metabolized by sequential oxidation of the phenyl and pyrimidine ring (Fig. 78.2). The major phase 1 metabolite was identified as 4-cyclopropyl5-hydroxy-6-methyl- N -( 4-hydroxy )-phenyl-2-pyrimidinamine
78.4 Absorption, Distribution, Metabolism, and Excretion
Metabolite 2
Metabolite 5
1
1
(YNHUN~
HO~
N
HOYYNHut" HO~ No OH
LOH OH
Metabolite 3
Figure 78.2
1703
Metabolite 6
Phase I metabolism of Cyprodinil in the rat.
(metabolite 2). This metabolite was excreted in the urine as ,B-glucuronic acid conjugate as well as mono- and disulfuric acid conjugates. Although female rats formed the monosulfate almost exclusively, the males excreted equal amounts of the mono- and disulfate. Further oxidation of the methyl group led to the formation of 4-cyclopropyl-5-hydroxy-6hydroxymethy 1- N -(4-hydroxy )-pheny1-2-pyrimidinamine (metabolite 3), which was excreted in the urine in unconjugated form. Alternative pathways proceeded either by sequential oxidation of the phenyl ring to the 4-hydroxy and 3,4-dihydroxy derivatives (metabolites 4 and 5), followed by oxidation of the methyl group (metabolite 6), or started with hydroxylation of the methyl group as the first oxidation step (metabolite 9). Urinary and biliary metabolites were found to be conjugated with ,B-glucuronic acid and sulfuric acid. The major metabolites identified in feces were the 5-hydroxypyrimidine derivative of Cyprodinil (metabolite 1) and metabolite 4. Metabolites 1 and 4 also were present in conjugated form in urine and bile. In addition, liver and kidney tissue residues were analyzed for metabolites 12 h after single oral administration of [2-14C]pyrimidyl-Cyprodinil at 100 mg/kg to male rats. During this period, 17.0% of the total applied dose was eliminated via urine. The identified liver and kidney metabolites essentially confirmed the metabolic pathways proposed upon analysis of metabolites isolated from excreta. However, two additional
metabolites were found in liver and/or kidney tissue, but not in excreta. Metabolite 7 was identified as ring-hydroxylated Nphenyl-guanidine, a breakdown product of the pyrimidine ring moiety. Metabolite 8 (i.e., 4-cyclopropyl-6-methyl-pyrimidine2-ylamine) demonstrated a cleavage of the parent molecule between the pyrimidine and the phenyl ring. Metabolite 7 was found exclusively in the liver, where it represented the major metabolite. Minor amounts of metabolite 8 were found in the liver and kidneys. 78.4.3 TOXICOKINETICS AND METABOLISM IN LACTATING GOATS
[U- 14 C]phenyl or [2-14C]pyrimidyl labelled Cyprodinil were orally administered in gelatin capsules once daily to lactating goats for four consecutive days at dose levels of about 0.2 and 10 mg/kg. The absorption and excretion of radioactivity were measured over 78 h. Six hours after the last dose, the animals were sacrificed and the tissue residues were determined. Independent of the label and the dose administered, Cyprodinil was absorbed in goats to a lesser extent and more slowly than in rats. The major route of excretion was in urine and feces, whereas excretion via the milk was minimal. Residues of radioactivity in edible tissues were generally low. The metabolic pathways of Cyprodinil in lactating goats were similar to those observed in the rat.
1704
CHAPTER 78 Cyprodinil
78.4.4 TOXICOKINETICS AND METABOLISM IN LAYING HENS [U- 14 C]phenyl or [2-14C]pyrimidyl labelled Cyprodinil were orally administered in gelatin capsules to laying hens once daily for 4 consecutive days at dose levels of about 0.4 and 19 mg/kg. The excretion of radioactivity was measured over 78 h. Six hours after the last dose, the animals were sacrificed and the tissue residues were determined. In laying hens, Cyprodinil was rapidly and completely eliminated. Residues of radioactivity in eggs and edible tissues were very low. The distribution and excretion pattern was independent of the dose and the site of labelling. The metabolic pathways of Cyprodinil in laying hens were similar to those observed in the rat.
78.5 TOXICITY TO LABORATORY ANIMALS 78.5.1 ACUTE TOXICITY The acute toxicity profile of Cyprodinil is summarized in Table 78.1. According to these data, Cyprodinil is unlikely to present an acute hazard in normal use (Class Ill, according to the WHO hazard classification scheme). Cyprodinil is nonirritant to skin and eye when classified according to the EC Directive 83/467, but it may cause sensitization by skin contact.
• • • • • •
Rat, 28-day gavage Rat, 28-day dermal Rat, 90-day feeding Mouse, 90-day feeding Dog, 90-day feeding Dog, 12-month feeding
Rat, 28-Day Gavage Cyprodinil was administered by gastric intubation at 0, 10, 100, and 1000 mg/kg to groups of 10 male and female Tif:RAIf (SPF) rats for 5 days per week for 4 weeks. The main findings in both sexes were a hepatomegaly, as expressed by increased liver weight at 100 and 1000 mg/kg with a corresponding hepatocellular hypertrophy, and increased thyroid weights associated with follicular cell hypertrophy at 1000 mg/kg. The findings in the thyroid were considered to be secondary to liver stimulation. Further changes at 1000 mg/kg included decreased body weight development and food consumption, hypochromasia, and increased serum concentrations of albumin, globulin, bilirubin, phospholipids, and cholesterol. Rat, 28-Day Dermal Groups of five male and female Tif: RAIf (SPF) rats were dermally treated with Cyprodinil at doses of 0, 5, 25, 125, and 1000 mg/kg for 6 h per day, 5 days per week for 4 weeks (OECD Guideline 410). Except for some clinical observations of doubtful relationship to treatment (piloerection, dyspnea, hunched posture) and pathological changes at the application site, which were attributed to the application procedure, no effects of treatment were observed. In particular, Cyprodinil did not induce local irritations or effects on clinical chemistry parameters.
78.5.2 SUBCHRONIC TOXICITY The following subchronic toxicity studies were performed: Table 78.1 Acute Toxicity of CyprodiniI Species, strain, number, Parameter
and sex
Result
Acute oral LDsO
Rat, Tif RAIF (SPF),
Greater than 2000 mglkg
5m+5f Mouse, TifMAG (SPF),
Greater than 5000 mglkg
5m+5f Acute dermal LDso
Rat, Tif RAIf (SPF),
Greater than 2000 mglkg
5m+5f Acute inhalation LCso
Rat, Tif RAlf (SPF),
Skin irritation
Rabbit, New Zealand
Eye irritation
Rabbit, New Zealand
Sensitization
Guinea pig, Pirbright
Greater than 1200 mg/m 3a
5m+5f Nonirritantb
White,3f Nonirritantb
White,3m (maximization test)
White,lOm+lOf
aMaximum attainable concentration.
bClassified according to EC Directive 83/467.
Sensitizingb
Rat, 90· Day Feeding Groups of 10 male and female Tif:RAIf (SPF) rats received Cyprodinil at dietary concentrations of 0, 50,300,2000 and 12,000 ppm for 90 days. An additional control and high dose group was kept for a 4-week recovery period. Treatment resulted in partially reversible reduction of body weight development and food consumption at 12,000 ppm. Elevated serum concentrations of cholesterol and phospholipids occurred in both sexes treated at 2000 and 12,000 ppm, and increased serum activities of hepatic enzymes (alkaline phosphatase, y-glutamyltranspeptidase) were noted at 12,000 ppm. These changes were completely reversible after 4 weeks of recovery. Slightly elevated weights were found for liver, kidneys, and thyroid gland in both sexes treated at 2000 and 12,000 ppm. In the recovery group, only the thyroid weights remained slightly elevated. Histopathology revealed treatmentrelated effects in the kidneys (chronic tubular lesion) as well as in the liver (hepatocellular hypertrophy, foci of necrotic hepatocytes, hepatocellular inclusion bodies), thyroid (hypertrophy of thyroid follicles), and pituitary gland (hypertrophy of pituitary cells). The changes in the thyroid and pituitary gland might be related to liver stimulation. Except for the renal lesions, all changes were at least partially reversible within the 4-week recovery period. The elevated serum concentrations of cholesterol and phospholipids, seen after treatment with Cyprodinil for 90 days in
78.5 Toxicity to Laboratory Animals
rats, are different from those reported by Terada et al. (1998a, b) for the structurally closely related anilinopyrimidine Mepanipyrim. Treatment of male rats at 4000 ppm Mepanipyrim for 3 weeks decreased the serum cholesterol, triglyceride, and phospholipid levels (Terada et aI., 1998b). In addition, these authors reported hepatocellular fatty vacuolation after treatment for 13 weeks at 200 ppm and above (Terada et aI., 1998a), an effect that did not occur in Cyprodinil-treated rats at doses up to 12,000 ppm.
Mouse, 90-Day Feeding Cyprodinil was administered to groups of 10 male and female Tif:MAGf (SPF) mice over 90 days at dietary concentrations of 0, 500, 2000, and 6000 ppm. The animals tolerated the subchronic dietary administration without mortality or overt clinical signs. Body weight gain, food consumption, and hematological profile remained unaffected by the treatment. The liver was the main target organ of toxicity. Treatment of males at 6000 ppm caused increased absolute and relative liver weight. Histopathological changes at 2000 ppm and above comprised single cell necroses in males and depletion of glycogen in females. Clinical chemistry parameters were not investigated. Dog, 90-Day Feeding Groups of four male and females beagle dogs received Cyprodinil at dietary concentrations of 0, 200, 1500, 7000, and 20,000 ppm for 90 days (OECD Guideline 409). Treatment with Cyprodinil was well tolerated even at very high dose levels. The animals reacted to the treatment mainly by reduced food consumption and reduced body weight gain. In addition, increased numbers of blood platelets were recorded in the males treated at the top dose level of 20,000 ppm. The laboratory examinations indicated no deviations of possible toxicological significance. In particular, the treatment did not affect the plasma concentrations of cholesterol. This observation is in line with the lack of changes in clinical chemistry parameters in Mepanipyrim-treated dogs (Terada et aI., 1998a). The lipofuscin deposition in Kupffer cells and hepatocytes, as observed with Mepanipyrim (Terada et aI., 1998a), was not seen with Cyprodinil. Dog, 12-Month Feeding Groups of four male and female beagle dogs received Cyprodinil at dietary concentrations of 0, 25, 250, 2500, and 15,000 ppm (OECD Guideline 452). The animals tolerated the treatment without clinical signs or mortality and no effects on laboratory parameters were encountered. Reactions to the treatment mainly consisted of reduced body weight gain and food consumption. In addition, the livers of the male dogs contained minor amounts of lipofuscin-like pigments. Lipofuscinosis also has been reported upon treatment of dogs with Mepanipyrim (Terada et aI., 1998a). 78.5.3 CHRONIC TOXICITY AND ONCOGENICITY The following chronic toxicity and oncogenicity studies were performed:
1705
• Mouse, 18-month feeding • Rat, 24-month feeding
Mouse, 18-Month Feeding Cyprodinil was administered to groups of 60 male and female Tif:MAGf (SPF) mice over 18 months at dietary concentrations of 0, 10, 150, 2000, and 5000 ppm (OECD Guideline 451). The animals tolerated the chronic treatment without clinical signs. Most probably due to a reduced body weight development, survival was slightly increased in the top dose group. Slightly increased liver and kidney weights were noted in the top dose group animals. Histopathological examinations revealed increased incidences of slight to moderate hyperplasia of the exocrine pancreas in the top dose group males. The incidence and distribution of neoplastic changes was similar in treated and untreated control groups, and remained within the range of the historical controls. Rat, 24-Month Feeding Cyprodinil was administered to groups of 80 male and female Tif:RAIf (SPF) rats at dietary concentrations of 0, 5, 75,1000, and 2000 ppm over 24 months (OECD Guideline 453). The chronic dietary administration of Cyprodinil was tolerated without clinical signs or treatmentrelated mortality. Hematological examinations revealed a slight prolongation of the prothrombin time in the top dose group males. At the 12-month interim sacrifice, higher kidney to body weight ratios were observed for both sexes at 2000 ppm. At terminal sacrifice, males showed increased absolute and relative liver weights at the top dose level and hepatic sinusoidal cystic dilatation at 1000 ppm and above. The incidence and distribution of neoplastic lesions gave no indication of a carcinogenic effect. In particular, there was no induction of thyroid follicular cell tumors, which were found to be elicited in rats by treatment with the structurally closely related anilinopyrimidine Pyrimethanil (Hurley, 1998). Enhancement of hepatic thyroid hormone metabolism and excretion are considered to be the mode of action of thyroid tumorigenesis (Hurley, 1998). 78.5.4 EFFECTS ON LIVER XENOBIOTIC METABOLIZING ENZYMES IN THE RAT In subchronic oral toxicity studies in the rat, Cyprodinil caused increased liver weights and hepatocellular hypertrophy. Its possible effects on hepatic xenobiotic metabolizing enzymes were investigated in an explorative 28-day study where male Tif RAIf (SPF) rats were treated by oral intubation at dose levels of 0, 100, and 1000 mg/kg per day (Table 78.2). The treatment was without effect on body weight, but caused increased liver weights to 123 and 160% of control at 100 and 1000 mg/kg, respectively. Increased microsomal protein and cytochrome P450 contents at 1000 mg/kg were accompanied by an induction of the investigated cytochrome P450 dependent monooxygenase activities, that is, ethoxyresorufin O-deethylase and pentoxyresorufin O-depentylase as well as lauric acid 11- and 12hydroxylase. The most prominent effect was a nearly 30-fold
1706
CHAPTER 78
Cyprodinil
Table 78.2 Effects of Cyprodinil on Liver Xenobiotic Metabolizing Enzymes in the Male Rata Dose (mg/kg body weight/day) Parameter
0
Final body weight (g)
341
357
330
(10)
(11)
(26)
Absolute liver weight (g) Microsomal protein (mg/g liver) Cytosolic protein (mg/g liver) Cytochrome P450 (nmol/g liver) Ethoxyresorufin O-deethylase (nmol/min g-l liver) Pentoxyresorufin O-depentylase (nmol/min g-l liver) Lauric acid II-hydroxylase (nmol/ming-1liver)
100
1000
10.1
12.5**
16.2***
(0.5)
(0.9)
(1.7)
25.9
25.3
35.0*
(1.3)
(4.6)
(5.0)
127
123
114***
(5)
(5)
(3)
20.6
IS.1
46.6***
(2.7)
(3.3)
(11.3)
4.40
9.40**
43.5***
(1.25)
(3.S6)
(10.1)
US
3.56**
33.6***
(0.12)
(1.92)
(6.95)
22.5
22.S
46.0***
(I.S)
(5.4)
(S.S)
Lauric acid 12-hydroxylase (nmol/ming- 1 liver)
14.9
21.0*
45.7***
(2.6)
(5.2)
(7.1)
Glutathione S -transferase
S7.4
IS4**
305***
(32.3)
(49)
(36)
1034
[[29
1139
(204)
(l2S)
(160)
(iJ.mol/ming-l liver) Fatty acid /5-oxidation (nmol/min g-l liver)
QMean values; standard deviations are given in parentheses. Asterisks indicate statistical significant difference: * p < 0.05, ** p < 0.01, *** p < 0.001. Five animals were used in each dosage group.
induction of pentoxyresorufin 0 -depentylase, a marker enzyme activity for phenobarbitone-inducible cytochrome P450 isoenzyme CYP2B1 (Whitlock and Denison, 1994), Cytosolic glutathione S-transferase activity toward 2,4-dinitrochlorobenzene was induced 2- and 3.5-fold at the lower and higher dose level, respectively, whereas cyanide insensitive fatty acyl CoA f3 oxidation, a marker for hepatic peroxisome proliferation (Lake and Lewis, 1993), was not affected. According to these data, the hepatomegaly induced by Cyprodinil in the rat can be interpreted as an adaptation to a functional load. 78.5.5 EFFECTS ON LIVER AND PLASMA LIPIDS IN THE RAT
Upon subchronic feeding to rats, Cyprodinil caused increased plasma concentrations of cholesterol and phospholipids. In an exploratory study, male Tif:RAIf (SPF) rats were treated with Cyprodinil for 28 days by oral intubation at dose levels of 0, 100, and 1000 mg/kg per day. After a 20-h fasting period, the animals were sacrificed and analyzed for the content of liver free cholesterol and cholesterol esters and for the concentration of serum total cholesterol, as well as for the concentration
of cholesterol bound to high density lipoprotein (HDL), low density lipoprotein (LDL), and very low density lipoprotein (VLDL) (Table 78.3). In the liver, the concentration of free cholesterol level remained unaffected and the cholesterol ester concentration was, if at all, slightly reduced in both treatment groups. In the rat, cholesterol is mainly transported by HDL and to a lower extent also by LDL, whereas VLDL is of minor importance (Carroll and Feldman, 1989). Serum total cholesterol concentration was increased by a factor of about 2 at 1000 mg/kg. As expected, cholesterol concentrations were increased in all three serum lipoprotein fractions of animals treated with 1000 mg/kg, whereby the higher total serum cholesterol level could largely be attributed to increased HDL and LDL levels. These data confirm that, upon subchronic administration to rats, Cyprodinil interacts with lipid homeostasis. In special investigations, the structurally related anilinopyrimidine Mepanipyrim was also shown to interfere with lipid homeostasis in the rat (Terada et aI., 1998b). However, the effects elicited by the two anilinopyrimidines are quite different (Table 78.4). Mepanipyrim caused fatty liver that comprised increased liver cholesterol, phospholipid, and triglyceride con-
78.5 Toxicity to Laboratory Animals
1707
Table 78.3 Effects of Cyprodinil on Liver Free Cholesterol, Liver Cholesterol Esters, and Serum Cholesterol in the Male Rata Dose (mg/kg body weight/day) Parameter Liver free cholesterol (~g/mg
homogenate protein)
Liver cholesterol esters (~g/mg
homogenate protein)
Serum total cholesterol (~mol/ml
serum)
High density lipoprotein cholesterol (~mol/ml
serum)
Low density lipoprotein cholesterol (~mol/ml
serum)
Very low density lipoprotein cholesterol (~mol/ml
serum)
0
100
1000
8.3
7.7
8.2
(0.5)
(0.5)
(0.5)
1.8
1.4
1.3
(0.2)
(0.2)
(0.2)
1.78
1.90 (0.30)
3.89***
(0.22)
(0.74)
1.24
1.36
2.23***
(0.09)
(0.24)
(0.42)
0.64
0.67
1.86***
(0.16)
(0.15)
(0.34)
0.020
0.036
0.128
(0.021)
(0.014)
(0.110)
a Mean values; standard deviations are given in parentheses. Asterisks indicate statistical significant difference: < 0.05, ** p < 0.01, *** p < 0.001. Five animals were used in each dosage group.
*p
centrations. Cyprodinil was without an effect on liver cholesterol concentration (liver phospholipid and triglycerides were not measured) and did not cause fatty liver. In blood, Mepanipyrim decreased cholesterol and high-density lipoprotein cholesterol, phospholipid, and triglyceride concentrations, whereas Cyprodinil caused increased cholesterol concentration, cholesterol concentrations in high-, low-, and very low-density lipoprotein fractions, and phospholipid concentrations. It was suggested that the fatty liver induced by Mepanipyrim in the rat is the result of an inhibition of intracellular transport of very low-density lipoproteins from the Golgi apparatus to the cell surface (Terada et aI., 1999). This hypothesis is in accor-
dance with its presumed mode of action in pathogenic fungi, where the compound was shown to block intracellular trafficking and secretion of plant cell-wall digesting enzymes (Miura et aI., 1994). The mechanism by which Cyprodinil induces increased blood cholesterol and phospholipid concentrations in the rat is not known. However, in vitro investigations with Cyprodinil in primary cultured rat hepatocytes showed that the hypercholesterolemia observed with this compound was neither due to increased hepatic cholesterol synthesis nor a consequence of inhibition of bile acid synthesis or bile acid transport (data not shown).
Table 78.4 Cyprodinil and Mepanipyrim: Comparison of Effects on Selected Lipid Parameters in Liver and Blood Following Subchronic Treatment of Rats Parameter
Cyprodinila
Mepanipyrimb
Liver free cholesterol
No effect
No effect
Liver cholesterol esters
No effect
Liver total cholesterol
No effect
t t t t
Liver phospholipid
Not measured
Liver triglyceride
Not measured
Plasma or serum total cholesterol c
t t t
Plasma or serum phospholipidc Serum high-density
.j, .j, .j,
lipoprotein cholesterol aCyprodynil was administered to Tif:RAIf (SPF) male rats via gavage for 4 weeks at a rate of 1000 mg/kg per day. bMepanipyrim was administered to F344/DuCJ:j male rats via diet for 3 weeks at a rate of 4000 ppm. Data from Terada et al. (l998b). CCyprodinil: plasma; Mepanipyrim: serum.
1708
CHAPTER 78
Cyprodinil
78.5.6 EFFECTS ON REPRODUCTION AND DEVELOPMENT Possible effects of Cyprodinil on reproduction and development were investigated in the following studies: • Rat, teratogenicity • Rabbit, teratogenicity • Rat, two-generation reproduction Rat, Teratogenicity Cyprodinil was orally administered by gastric intubation to groups of 20--22 pregnant Tif:RAIf (SPF) rats from day 6 to day 15 of gestation at daily doses of 0, 20, 200, and 1000 mg/kg per day. The dams were sacrificed on day 21 of gestation and the fetuses were removed, weighed, sexed, examined for external malformations, and subjected to visceral and skeletal examination. In the dams, no treatment-related mortality occurred and no clinical signs of possible relevance were noted. Body weight development and food consumption were significantly reduced in the top dose animals. No treatment-related changes were noted upon necropsy. No treatment-related effects were observed on pregnancy rate, corpora lutea, implantations, early or late resorptions, sex or number of fetuses. Weight reduction and reduced ossification of digits and metacarpalia were observed in the top dose group. These findings were considered to be a consequence of the observed maternal toxicity at top dose treatment. In conclusion, the results of this study gave no indication of teratogenic effects. Rabbit, Teratogenicity Groups of 17-18 pregnant Russian rabbits were orally treated by gavage from day 7 to day 19 of gestation at daily doses of 0,5,30, 150, and 400 mg/kg per day. The does were sacrificed on day 29 of gestation and the fetuses were removed by hysterectomy, weighed, sexed, examined for external malformations, and subjected to visceral and skeletal examination. Neither mortality nor treatment-related clinical signs occurred in the does. The body weight development and the food consumption of the top dose animals was reversibly reduced during the treatment period. No treatment-related changes were noted upon necropsy. No treatment-related effects were observed on pregnancy rate, corpora lutea, implantations, early or late resorptions, sex, or number of fetuses. The fetal weights were similar in treated and untreated animals, and the external, visceral, and skeletal examinations of the fetuses revealed no treatment-related effects. In conclusion, the results of this study gave no indication of teratogenic effects. Rat, Two-Generation Reproduction Cyprodinil was administered to groups of 30 male and female Tif:RAlf (SPF) rats at dietary doses of 0, 10, 100, 1000 and 4000 ppm over two generations. In the FO generation, no treatment-related clinical signs or mortality were noted, but slight changes in body weight gain and food consumption were observed in the top dose group.
The parameters of fertility and reproduction showed no significant intergroup differences. At necropsy, slightly higher organ weights (liver, kidney, adrenal gland) and renal tubular basophilia were observed. The Fllitter sizes were not affected by treatment and the average pup weight at birth was similar in all groups. The body weight development of the top dose group pups was significantly reduced during the lactation period, but the pups reached all milestones of physiological development at the same time as the untreated animals. The parameters of fertility and reproduction were similar in treated and untreated groups. At necropsy, increased liver weights were noted, but the histopathological examination revealed no changes of toxicological significance. In the F2 litters of the top dose group, the mean pup weight at birth was slightly lower than the control value. All litters reached the milestones of physical development at similar time points. In conclusion, the results of this study gave no indication of effects on reproduction or fertility. 78.5.7 NEUROTOXIC EFFECTS The neurotoxic potential of Cyprodinil was investigated in the following studies: • Rat, single oral dose • Rat, 90-day feeding Rat, Single Oral Dose Groups of 10 male and female Tif:RAIf (SPF) rats received a single oral dose of 0, 200, 600, or 2000 mg/kg. The study included a functional observational battery that covered central nervous system (CNS) activity, CNS excitation, and sensorimotor, autonomic, and physiological functions. Neurological examinations covered sensorimotor functions, autonomic functions, and sensorimotor coordination. Motor activity was assessed and neuropathological examinations included different areas of the brain, the spinal cord, peripheral nerves, and muscles. The time of peak effect, determined in a range-finding test, was found to be approximately 2 h after treatment. There was no effect of treatment on mortality, body weight, or food consumption. Observations and functional tests showed relevant changes mainly at the time of peak effect. In females of the intermediate and high dose groups, reduced activity, hunched posture, piloerection, and increased responsiveness to sensory stimuli were observed; hunched posture was seen also in a few low dose females. In females of the two higher dose groups, signs lasted up to test day 4 and were considered to indicate toxicity. In addition, a dose-related decrease in body temperature was observed. Changes in motor activity parameters were limited to the time of peak effect. In high dose males and in females of the two higher dose groups, horizontal and vertical activity parameters were reduced.
78.6 Mutagenicity Macroscopic and microscopic examinations of the multiple areas of the central and peripheral nervous system, the eyes, optic nerves, and skeletal muscle of the male and female control and high dose animals did not reveal treatment-related neuropathologic changes. The results of this study gave no indication of neurotoxic effects. Rat, 90-Day Feeding Groups of 10 male and female Tif:RAIf (SPF) rats received Cyprodinil at dietary concentrations of 0, 80, 800, and 8000 ppm for 90 days. The study included a functional observational battery, covering CNS activity, CNS excitation, and sensorimotor, autonomic, and physiological functions. Neurological examinations covered sensorimotor functions, autonomic functions, and sensorimotor coordination. Motor activity was assessed and neuropathological examinations included different areas of the brain, the spinal cord, peripheral nerves, and muscles. There was no effect on mortality and clinical signs. In high dose animals, a moderately reduced body weight gain was seen throughout the treatment period. In high dose animals, absolute and relative liver weights were increased in both sexes and kidney weights were elevated in females. Microscopic examination revealed hepatocellular hypertrophy in the liver of high dose animals. In addition, chronic tubular lesions combined with tubular casts and single cystic changes in the kidneys, and hypertrophy of the follicular epithelial cells in the thyroid gland were seen in these animals. Observations and functional tests showed no effect of toxicological significance and no treatment-related effects on the different motor activity parameters. Neuropathologic examination of the eyes, optic nerves, and multiple areas of the central and peripheral nervous system of the male and female control and high dose animals revealed no treatment-related neuropathologic changes. The results of this study gave no indication of neurotoxic effects. 78.5.8 PHARMACOLOGICAL EFFECTS General pharmacological studies were performed with male ICR mice, male Wistar rats, and male Hartley guinea pigs. Single dose levels of 0, 150, 500, 1500, and 5000 mg/kg Cyprodinil were orally administered. In the in vitro experiments, the test material was dissolved in ethanol and applied at concentrations of 0, 0.1,1.0, and 10 J.lg/ml. Mouse, General Behavior and CNS Tests No animals died at any dose level. At 1500 mg/kg, slightly decreased spontaneous motor activity, slightly dilated pupil size, and slightly narrowed palpebral opening were observed within the first few hours. All signs disappeared by 6 h after treatment. At 5000 mg/kg, slight piloerection and abnormalities of body and limb position were seen in addition. All signs disappeared within 24 h in this dose group. Doses of 1500 mg/kg or higher prolonged hexobarbital-induced sleeping time. No effects on tonic extensor, clonic convulsions, or coma in electrically stimulated mice were noted at any dose level.
Rat, Body Temperature in any dosage group.
1709
Body temperature was not affected
Rat, Cardiovascular System Treatment with Cyprodinil had no effect on the systolic blood pressure. At 5000 mg/kg, the heart rate was significantly decreased at 1 h after dosing.
In Vitro, Autonomic Nervous System Cyprodinil inhibited the induced contractions of isolated guinea pig ileum at concentrations of 10 J.lg/ml. No effects were noted at lower concentrations. Mouse, Gastrointestinal System No effect was seen on intestinal transport at any dose level. Mouse, Skeletal Muscle System traction test at any dose level.
No effect was seen in the
Rat, Hematology No effects on prothrombin time or activated partial thromboplastin time were seen at any dose level. No hemolytic effects were noted.
78.6 MUTAGENICITY The mutagenic potential of Cyprodinil was investigated in five independent studies that covered different end points in eukaryotes and prokaryotes in vivo and in vitro. No induction of back mutations was noted in four strains of Salmonella typhimurium (TA 98, TA 100, TA 1535, and TA 1537) or in Escherichia coli WP2 uvrA. The compound was tested in the presence and in the absence of an extrinsic metabolic activation system (rat liver S9 fraction), covering a concentration range of 20-5000 J.lg/plate. The induction of gene mutations was further investigated at the hprt gene of Chinese hamster V79 cells. Concentrations up to the limit of toxicity of 96 and 30 J.lg/ml were tested in the presence and in the absence of rat liver S9 fraction, respectively. No increased incidence of gene mutations was detected in this in vitro system. Chinese hamster CHO cells were used for the detection of chromosome aberrations in vitro. Three experiments in the presence and in the absence of rat liver S9 fraction were performed by applying different exposure scenarios. Concentrations higher than 50 J.lg/ml could not be investigated due to cytotoxicity. In none of the experiments was an increased incidence of metaphases containing chromosome aberrations seen. Primary cultures of rat hepatocytes were treated with Cyprodinil at concentrations up to 80 J.lg/ml to investigate DNA damaging effects. Higher concentrations caused cytotoxicity. None of the tested concentrations caused enhanced unscheduled DNA synthesis, which would be indicative of the DNA damaging activity of the compound. The formation of micronuclei was investigated in bone marrow cells of Tif:MAGf (SPF) mice. Single oral doses of up to 5000 mg/kg body weight were administered 16, 24, and 48 h
1710
CHAPTER 78
Cyprodinil
before preparation of bone marrow. No cytotoxicity was observed at any dose level and the incidence of micronuc1eated polychromatic erythrocytes was not affected. Therefore, there is no evidence for a c1astogenic or aneugenic activity of Cyprodinil in somatic cells in vivo.
78.7 TOXICITY TO HUMANS 78.7.1 DIRECT OBSERVATIONS AND HEALTH RECORDS Three cases of moderate, reversible local irritation (erythema, swelling of eyelids) occurred among laboratory personnel during formulation development. No further complaints were noted.
78.7.2 DIAGNOSIS OF POISONING In animal studies, symptoms of acute intoxication were unspecific and transient only. The same can be expected for humans. However, no case of intoxication with Cyprodinil has yet been observed.
78.7.3 SENSITIZATION OBSERVATIONS No cases of skin sensitization were recorded in humans.
78.7.4 PROPOSED TREATMENT Whereas no specific antidote is known, symptomatic therapy is to be applied on persons who show symptoms after exposure to Cyprodinil.
REFERENCES Carroll, R. M., and Feldman, E. B. (1989). Lipids and lipoproteins. In "The Clinical Chemistry of Laboratory Animals" (w. F. Loeb and F. W. Quimby, eds.), pp. 95-116. Pergamon, New York. Fritz, R., Lanen, c., Colas, v., and Leroux, P. (1997). Inhibition of methionine biosynthesis in Botrytis cinerea by the anilinopyrimidine fungicide pyrimethanil. Pestic. Sci. 49, 40--46. Heye, U. l., Speich, J., Siegle, H., Steinemann, A., Forster, B., Knauf-Beiter, G., Herzog, J., and Hubele, A. (1994). CGA 219417: A novel broad-spectrum fungicide. Crop Protection 13,541-549. Hurley, P. M. (1998). Mode of carcinogenic action of pesticides inducing thyroid follicular cell tumors in rodents. Environ. Health Perspect. 106, 437445. Lake, B. G., and Lewis, D. F. V. (1993). Structure-activity relationships for chemically induced peroxisome proliferation in mammalian liver. In "Peroxisomes: Biology and Importance in Toxicology and Medicine" (G. Gibson and B. Lake, eds.). Taylor and Francis, London. Masner, P., Muster, P., and Schmid, J. (1994). Possible methionine biosynthesis inhibition by pyrimidinamine fungicides. Pestic. Sci. 42, 163-166. Milling, R. J., and Richardson, C. J. (1995). Mode of action of the anilinopyrimidine fungicide pyrimethanil. 2. Effects on enzyme secretion in Botrytis cinerea. Pestic. Sci. 45, 43-48. Miura, I., Kamakura, T., Maeno, S., Hayashi, S., and Yamaguchi, I. (1994). Inhibition of enzyme secretion in plant pathogens by mepanipyrim, a novel fungicide. Pestic. Biochem. Physiol. 48, 222-228. Terada, M., Mizuhashi, F., Tomita, T., Inoue, H., and Murata, K. (1998a). Mepanipyrim induced fatty liver in rats but not in mice and dogs. J. Toxicol. Sci. 23, 223-234. Terada, M., Mizuhashi, F., Tomita, T., and Murata, K. (1998b). Effects ofmepanipyrim on lipid metabolism in rats. J. Toxicol. Sci. 23,235-241. Terada, M., Mizuhashi, F., Murata, K., and Tomita, T. (1999). Mepanipyrim, a new fungicide, inhibits intracellular transport of very low density lipoprotein in rat hepatocytes. Toxicol. Appl. Pharmacal. 154, 1-11. Whitlock, J. P., and Denison, M. S. (1994). Induction of cytochrome P450 enzymes that metabolize xenobiotics. In "Cytochrome P450: Structure, Mechanism, and Biochemistry" (R. Ortiz de Montellano, ed.), 2nd ed., pp. 367-390. Plenum, New York Yamagata, S. (1989). Roles of O-acetyl-L-homoserine sulfhydrylase in microorganisms. Biochimie 71, 1125-1143.
CHAPTER
79 Captan and Folpet Elliot B. Gordon Makhteshim-Agan of North America, Inc.
79.1 INTRODUCTION Captan and foIpet are fungicides that have been in use for over 50 years. During this period there have been no reports of systemic toxicity and only a very low incidence of skin sensitization. This record of safe use is consistent with the toxicological data base developed for these compounds. Adverse findings in laboratory test systems consist of mutagenicity, carcinogenicity, and eye irritation. The first two outcomes resulted in the EPA classifications of captan and folpet as "probable human carcinogens"; the latter outcome resulted in reentry restrictions on farm workers. Analysis of the oncogenic findings has demonstrated that the etiology of this effect is compensatory proliferation of duodenal crypt cells following damage to duodenal villi caused by high dietary exposure. Analysis of mutagenicity delineates a clear distinction between positive in vitro and negative in vivo effects. This paradox is explained by rapid degradation in blood: captan degrades with a half-life of less than I s; folpet has a half-life of less than 5 s. Using a margin of exposure (MOE) analysis for human cancer, dietary exposure has a MOE of 1,000,000. The EPA's proposed 1996 cancer guidelines characterize this exposure level as "not likely at all to cause cancer in humans at low doses." In practical terms, this means that neither captan nor folpet is a human carcinogen. Draize rabbit studies show that these compounds are severe ocular irritants. Extensive experience, however, particularly with reentry operations, has shown that this laboratory phenomenon is not predictive of human experience. Captan and folpet remain highly valued, low risk fungicides.
79.1.1 OVERVIEW Captan and folpet are broad-spectrum protectant fungicides whose mode of action centers on their reaction with thiols. These compounds along with a third, captafol, are collectively called chloroalkylthio fungicides due to the presence of side chains that contain chlorine, carbon, and sulfur. Of the chloroalkylthio fungicides, captan and folpet predominate in Handbook of Pesticide Toxicology
Volume 2. Agents
agronomic practice today; captafol registrations in the United States were withdrawn in 1988. Related compounds associated with this fungicide class, but never registered in the United States, are dichlofluanid and tolylfluanid. These later two compounds have a fluorine atom substituted for one of the terminal chlorine atoms. Early investigations on captan and folmet focused on their mutagenicity. These assays, conducted in vitro, showed captan and folpet to be active. This led regulators to ascribe a genotoxic basis to the duodenal tumors that occur in mice, resulting in a linear low-dose extrapolation for cancer risk assessment. Teratogenicity studies of folpet were conducted following the perceived association of folpet's phthalimide moiety with the human teratogen S-thalidomide; these structures have since been shown to be toxicologically unrelated. Captan and folpet show developmental toxicity at maternally toxic doses; neither compound is teratogenic. A number of reviews have addressed the toxicology of the chloroalkylthio fungicides (Ecobichon, 1996; Edwards et a!., 1991; Elder, 1989; IARC, 1983; Saunders and Harper, 1994; Trochimowicz et a!., 2001; U.S. EPA, 1975). Data now show that a threshold-based nonmutagenic mode of action exists for the development of tumors in mice. Captan and folpet are not carcinogenic to farm workers because systemic exposure is absent; they will not cause tumors in consumers because the margin of exposure from dietary intake is approximately 1 million. The U.S. Environmental Protection Agency (U.S. EPA) has reviewed captan and folpet under some of the provisions of the Food Quality Protection Act (FQPA) of 1996 (U.S. Congress, 1996), but has notreevaluated them under the proposed new carcinogen risk assessment guidelines (US. EPA,1996). This chapter identifies the hazards associated with captan and folpet, and assesses risks for both cancer and noncancer endpoints. Controversy continues to surround their cancer risk characterization. The EPA classified these compounds as B2 carcinogens over 15 years ago (Engler, 1986; U.S. EPA, 1986a) and has yet to review the initial classification using the 1996 draft guidelines (U.S. EPA, 1999a, b).
1711
Copyright © 200 1 by Academic Press. All rights of reproduction in any form reserved.
1712
CHAPTER 79
Captan and Folpet
79.1.2 HISTORY AND USE
Captan was first registered in the Vnited States on March 8, 1949 as a fruit tree spray (NPIRS, 1999a) and its properties were described in 1953 (Kittleson, 1953). This compound proved extremely efficacious, spurring chemists to turn out a series of analogs in an attempt to capitalize on the fungicidal properties of the trichloromethylthio moiety (Horsfall and Rich, 1957; Kittleson, 1953; Lukens, 1966). Folpet was synthesized after captan; captafol was the last to be developed. As preventative fungicides, they are efficacious when applied prior to the establishment of pathogenic fungi. Captan and folpet are often used in integrated pest management (IPM) programs in conjunction with other fungicides. Registrations cover both agricultural and industrial uses (NPIRS, 1999b; V.S. EPA, 1985b). Captan is also efficacious as a bacteriostat in cosmetics (Elder, 1989). The V.S. EPA issued registration standards for captan (V.S. EPA, 1986b), folpet (U.S. EPA, 1987), and captafol (V.S. EPA, 1984a). A special review for captafol (V.S. EPA, 1985a) concluded in 1988 with the voluntary withdrawal of registrations. A special review for captan was completed in 1989 with the issuance of Position Document 4 (V.S. EPA, 1989). Reregistration Eligibility Decision documents (REDs) have now been promulgated for captan and folpet (V.S. EPA, 1999a, b). 79.1.3 TOXICOLOGICAL OVERVIEW
The principle governing the toxicity of captan and folpet centers on their rapid reaction with thiol groups (i.e., sulfhydryl, -SH groups). This reaction results in degradation of the parent compound. Thiophosgene is a key degradation product that also reacts with thiols as well as other functional groups. It is reactive and short-lived. The net result of these chemical interactions is that both captan and folpet elicit primary toxicological effects locally at the site of initial contact. In mice, dietary exposure results in local irritation of the gastrointestinal tract, predominantly in the duodenum. It is solely at this site that tumors eventually develop. Continued administration of high doses often leads to secondary effects such as decreased body weight or developmental retardation in fetuses and pups. Pursuant to the Food Quality Protection Act and subsequent guidance by the EPA, captan and folpet have been found to share a common mechanism of toxicity with regard to the development of duodenal tumors in mice (see discussion in Section 79.4). Although this finding requires that both compounds be considered for cumulative risk assessment, it affords the toxicologist the opportunity to combine mechanistic data for each into one unified data base. Captafol, in contrast, was found not to share a common mechanism of toxicity with captan or folpet. The paradigm that details the methods for conducting cumulative risk assessments are under development at EPA. The FQPA also requires consideration of the special sensitivities to infants and children, and the potential for endocrine disruption. There is no indication that captan or folpet shows disproportionate toxicity to infants or children. The mechanism by which
these compounds exert their toxicity would make such a distinction unexpected. The EPA has recognized this for captan but currently has assigned an additional threefold safety factor for folpet (see Section 79.3.4; V.S. EPA, 1999a, b). Captan or folpet shows no evidence of being endocrine disruptors. The EPA, however, is still developing its approach to identify endocrine disruptors. Captan and folpet induce mutagenic and clastogenic effects in a variety of in vitro assays. Mutagenic activity in vivo, however, does not occur. This paradox is explained by the extremely rapid degradation of these compounds in the intact animal. Whereas the delivered dose is negligible, captan and folpet are classic examples of the adage, "the (delivered) dose makes the poison." Despite the obvious potential for mutagenic events, the dose at sensitive targets in the intact animal, such as cellular DNA, is essentially zero. Although these fungicides have low acute toxicity, their interaction with biological tissues can cause irritation. Thus, a low percentage of exposed people exhibit ocular or respiratory irritation or dermal sensitization reactions. Persons handling these materials should also avoid inhalation due to their irritant properties. The compounds are not considered teratogens, reproductive toxins, or selective developmental toxins. The appearance of duodenal tumors in mice fed diets admixed with captan or folpet is a key toxicological finding that has heretofore been one of the central issues for the regulation of these compounds. Data show that a mode of action based on increased rates of cell proliferation, a threshold phenomenon, accounts for the incidence of tumors seen. This proliferative pressure is thought to promote nascent tumor cells that are normally resident within the duodenal crypt compartment. The EPA, however, continues to classify both compounds, under the 1986 Carcinogen Risk Assessment Guidelines, as probable human carcinogens (V.S. EPA, 1984b, 1986a). Human cancer risk assessments of captan and folpet show that farm workers are at no cancer risk because there is no systemic exposure. Persons exposed to residues in their food are not at risk because the margins of safety for both compounds are approximately 1 million (see Section 79.5).
79.2 PHYSICAL PROPERTIES AND CHEMICAL REACTIONS 79.2.1 OVERVIEW The toxicology of the chloroalkylthio fungicides is dependent on their physical properties and chemical reactions. The structures of captan and folpet along with typical ring degradates are shown in Figs. 79.1 and 79.2. The chemical identity and physical properties are noted in Table 79.1, and the rates of selected chemical reactions are shown in Table 79.2. The characteristic chemical moiety for captan and folpet is the trichloromethylthio side chain that is connected to an imide ring structure by way of a nitrogen-sulfur bond. Captan's ring is tetrahydrophthalimide (THPI) and folpet's is phthalimide. This ring imparts certain
79.2 Physical Properties and Chemical Reactions Chemical
Structure
Chemical
Structure
Folpet
Captan
4,5-cyclohexene-I,2-dicarboximide (THPl)
1713
~M'
Phthalimide(PI)
o
c»' o
4,5-epoxy-I,2-dicarboximide (THPI expoxide) Phthalamic Acid
~NH2 ~OH o
4,5-dihydroxy-I,2-dicarboximide (4,5-diOH THPI)
Phthalic Acid 7 -hydroxy-4,5-cydohexene-1 ,2-dicarboximide
~OH ~OH o
(ciltrans-3-0H THPI)
Figure 79.2 6-hydroxy-4,5-cydohexene-I,2-dicarboximide (cis/trans-5-0H THPI)
o l-amido-2-carboxy -4,5-cydohexene
CC I
(cis/trans-THPAM) 0
&
lbru2 (DOH
11
6-hydroxy-l-amido-2-carboxy -4,5-cydohexene (3-0H THP-arnic acid)
CNH,
the nitrogen-sulfur bond linking this moiety to the imide ring. Analogs with very stable bonds prove to be ineffective fungicides, whereas analogs with bonds that are overly labile degrade spontaneously (Horsfall and Rich, 1957; Lukens, 1966, 1967). The hydrolytic and thiol reactions serve to degrade the parent molecule and thus influence the toxicology outcome by effectively reducing or eliminating exposure.
I
(DOH
Figure 79.1
Folpet and its ring metabolites.
Captan and its ring metabolites.
physical properties to the molecule, but is subordinate to the side chain with regard to the fungicide's toxicological properties. The phthalimide ring is aromatic and, as such, is a resonance structure; THPI has one double bond between carbons 3 and 4. This hexene imide, unlike phthalimide, is nonplanar (Fickentscher et aI., 1977). 79.2.2 PHYSICAL PROPERTIES Captan and folpet have similar physical properties. They have low water solubility, low volatility, and melt at approximately the same temperature. Octanol-water coefficients are high for both, although folpet's Kow is somewhat higher than that of captan. 79.2.3 CHEMICAL REACTIONS Captan and folpet are unstable in aqueous solution, but the rate of hydrolysis is slow compared to their reaction with thiols. The key to their fungicidal efficacy is the balance between the reactivity of the trichloromethylthio moiety and the stability of
79.2.3.1 Hydrolysis The rates of aqueous hydrolysis increase in alkaline conditions and are more rapid for folpet than captan at comparable pH values (Table 79.2). At pH 5, for instance, captan is approximately eight times more stable than folpet; thus, in the acid conditions of the stomach, it would be expected that relatively more folpet degradation products would be present compared to captan. The higher hydrolytic rates for folpet are related to the higher standard free energy of the phthalimide ring structure compared to the THPI ring (Lukens, 1966). 79.2.3.2 Reaction with Thiols As previously noted, reactions with thiol groups are central to the fungicidal and toxicological properties of captan and folpet. This reaction has been studied with glutathione (GSH), proteins, and other thiol-containing compounds. In general, the thiol group is oxidized (e.g., GSH ~ GSSG and cysteine ~ cystine). Common to both captan and folpet is the generation of thiophosgene during degradation. This chemical entity appears to be a contributing toxicophore in that it rapidly reacts with a variety of functional groups in addition to thiols (Lukens, 1969; Lukens and Sisler, 1958b; Sharma, 1986). A general scheme of degradation for captan and folpet is shown in Fig. 79.3. The rate of hydrolysis is faster for folpet than for captan, whereas the reverse is true for thiol-mediated degradation.
1714
CHAPTER 79
Captan and Folpet
Table 79.1 Physical Properties of Captan and Folpet Parameter
Captan
Folpet
CAS number
133-06-2
133-07-3
Molecular weight
300.61
296.56
Formula A
C9HgCI3N02S
C9H4CI3N02S
FormulaB
C6Hg(C =OhN-SCCI3
C6H4(C=OhN -SCCl3
IUPACname
1,2,3,6-Tetrahydro- N -( trichloromethyl
N -(trichloromethyl thio) phthalimide
thio) phthalimide CA name
3a,4,7, 7a-Tetrahydro-2-
2-[(Trichloromethyl) thio]-lH-
[(trichloromethyl) thio]-IH-isoindole-
isoindol-I,3(2H)-dione
1,3(2H)-dione Physical form
Crystals
Crystals
Melting point
178°C
177°C
Solubility, water
3.3 mglliter at 25°C
I mglliter at 20°C
Solubility, acetone
3.0 gllOO ml
3.4 g/lOO ml
10gKow
2.35
2.85
The reaction of captan and folpet with cysteine results in the formation of thiazolidine-2-thione-4-carboxylic acid (TTCA; Lukens and Sisler, 1958a). This compound is seen in mammalian metabolism studies (DeBaun et aI., 1974) and has been suggested for use as a biological marker for human exposure assessment (Krieger and Thongsinthusak, 1993; van Welie et aI., 1991). The fate of captan and folpet in human and rabbit blood has been investigated (Crossley, 1967a, b). Crossley added captan and folpet to human blood and measured the decline of the parent with time and, concurrently, the increase of the imide ring (THPI or phthalimide). At initial concentrations of 1 Il-giml, captan degraded with a half-life of 18 s. The degradation of folpet was three times slower, with a half-life of 54 s. By measuring the generation of the imide rings, it was shown that the parent compounds actually degraded rather than complexed with blood constituents. These investigations were carried out at 22°C with unlabeled materials. Recent investigations of degradation rates employed radiolabe led captan and folpet and physiological temperatures (37°C).
As predicted by the QlO (Purves et aI., 1992), increased temperature results in a higher rate of degradation. The new data demonstrate that at physiological temperatures, captan degrades rapidly in human blood, having a t l / 2 of 0.97 s, whereas folpet degrades somewhat slower but still quite rapidly (i.e., t 1/ 2 = 4.9 s; Gordon et al., 2001). These data, which demonstrate the rapid degradation of the two compounds, are of course a critical component in any exposure assessment and risk evaluation modeling. With oral exposure, it is unlikely that captan, folpet, or thiophosgene would survive long enough to reach systemic targets such as the liver, uterus, or testes. With dermal exposure and subsequent low absorption, captan will be eliminated in less than 7 sand folpet in less than 35. This determination follows from the notion that compounds are considered to be essentially gone in seven half-lives (Medinsky and Klaassen, 1996). 79.2.3.3 Reaction with Proteins Investigations on the effects of captan and folpet with proteins generally have been carried out in vitro. Such studies identify potential interactions that may occur in the living animal; how-
Table 79.2 Rates of Chemical Reactions of Captan and Folpet Captan
Folpet
Reference
pH5
18.8 h
2.6 h
Cap tan: Pack (1987)
pH7
4.9h
l.1h
Folpet: Ruzo and Ewing
pH9
8.3 min
1.5 min
18 s, 22°C
51 s, 22°C
Captan: Crossley (1967a)
0.97 s, 37°C
4.9 s, 37°C
Gordon et al. (200 I)
I min
3min
Liu and Fishbein (1967)
Parameter Aqueous Hydrolysis
Reaction with blood thiols
Folpet: Crossley (1967b)
Degradation half-life (acid conditions to minimize hydrolysis) Decolorization of dithionitrobenzoic acid (DTNB)-thiol complex in blood
(1988)
79.2 Physical Properties and Chemical Reactions R- SCCI3 Captan/Folpet
Reaction with Thiols
tion was measured at 25°C and was 1.9 x 104 liter/(mol min) for captan and 1.5 x 104 liter/(mol min) for folpet. Other effects include the inhibition of Escherichia coli RNA polymerase (Elder, 1989), the inhibition of RNA synthesis by intact bovine nuclei (Elder, 1989), the inhibition of microsomal cytochrome P-450 benzphetamine N -demethylase and aniline hydroxylase after intraperitoneal dosing (Dalvi, 1988, 1989), the inhibition of the Ca2+ transport ATPase in human erythrocytes (Janik, 1986), and the inhibition of oxidative phosphorylation in rat liver mitochondria, correlated to mitochondrial swelling (Elder, 1989). Captan also disrupted the differentiation of cultured cells from the midbrains and limb buds of 34-36 somite rat embryos in vitro (Flint and Ortaon, 1984) and inhibited the attachment of tumor cells to polyethylene disks that were coated with concanavalin (Braun and Horowicz, 1983).
R'SSR' HCI
S 11
/C, Cl Cl Thiosphosgene Reaction with Thiols TTCA
+ 2HCl
o
2RSH
Reaction with sulfite
11 RSCZR
C
79.2.3.6 Thiophosgene
DMS-Acid [0]
2HCI
RSR + CS2
1715
)
DMS-O
Figure 79.3 General degradation scheme. For captan, R = THPI (Tetrahydrophthalimide); for folpet, R = PI (phthalimide); TTCA: thiazolidine2-thione-4-carboxylic acid; DMS-Acid: dithio-bis-methanesulphonic acid; DMS-O: monosulfoxide of dithio-bis-methanesulphonic acid.
ever, for cap tan and folpet, the rapid degradation of reactive species and the resultant limitation in exposure prevent many of these reactions from resulting in in vivo toxicological phenomena. FoIpet reacts with thiol-containing proteins (e.g., glyceraldehyde 3-phosphate; Siegel, 1971a), non-thiol-containing proteins (e.g., a-chymotrypsin; Siegel, 1971 b), and nuclear histones (Couch and Siegel, 1972, 1977). These reactions are often pH dependent. 79.2.3.4 Reaction with DNA The reactions of captan and folpet with DNA are not well characterized. Captan and folpet induce point mutations and clastogenic changes in vitro, but the mechanism by which such effects are induced has not been elucidated. When 35 S-captan was incubated with calf thymus DNA in buffer at pH 7.5 or 9.0, binding of appreciable amounts of radioactivity could not be demonstrated (Couch and Siegel, 1977). Captan reacts with guanine in vitro to produce 7-(trichloromethylsulfenyl) guanine (Elder, 1989) as reported by FAOIWHO (1990). In vivo DNA binding studies are discussed in Section 79.3.5.5. 79.2.3.5 Miscellaneous Chemical Reactions Because captan and folpet are reactive, there are unlimited opportunities for chemical reactions in isolation. Absent data that indicate exposure in vivo or relevance to the intact animal, these observations remain ancillary, reflecting their chemical reactivity, but having little bearing on mammalian toxicity. Captan and folpet react with p-nitrothiophenol via the thiol group (Liu and Fishbein, 1967). This differential rate of reac-
Thiophosgene (CAS 463-71-8) is a very short-lived compound that has a broad spectrum of reactions with a variety of functional groups (Sharma, 1978, 1986). Although this compound hydrolyzes at a slower rate than its oxygen analog, phosgene, the rate is sufficient to eliminate mutagenic activity in Salmonella typhimurium TA 100 when dimethyl sulfoxide (DMSO) is the solvent (Schuphan et aI., 1981). Thiophosgene is a toxicophore of captan and folpet, although its role in their fungicidal properties has been questioned (Lien, 1969). It is volatile and reacts with water to form carbonyl sulfide (COS) and two molecules of hydrogen chloride. The carbonyl sulfide then reacts with another water molecule to form hydrogen sulfide and carbon dioxide (Fig. 79.3). Thiophosgene has two reactive sites associated with the carbon atom. Whereas both chlorine atoms are electronegative, the carbon atom becomes positively charged, thus creating an electrophile. The reaction with cysteine is shown in Fig. 79.4. 79.2.4 METABOLISM The fate of captan and folpet in mammalian systems is determined by an amalgam of nonenzymatic chemical reactions with thiols and subsequent enzyme-mediated metabolism that predominately involve the generation of ring metabolites. In the intestine, both hydrolysis and thiol-mediated reactions occur. The rate of hydrolysis is particularly sensitive to pH, and the transition from the acid environment of the stomach to the neutral or basic conditions of the duodenum promotes the hydrolytic breakdown of these materials. These fungicides undergo a similar pattern of degradation (Fig. 79.3). The side chain is either fully mineralized or forms by-products such as TTCA with cysteine. The respective imides, THPI and phthalimide, are initially formed either through hydrolysis or through reaction with thiols. These are subsequently metabolized to secondary products; the THPI hexene structure of captan is more extensively metabolized than the phthalimide structure offolpet (Figs. 79.1 and 79.2).
1/16
CHAPTER 79
Captan and Folpet
o
79.2.4.2 Effect on Glutathione Levels in the Duodenum
11
Cl
+
Cl
'c/ 11
S
C-OH
c-c/
-
I
I
S, /NH
+ 2HCI
C 11
S Cysteine
Thiophosgene
Thiazolidine -2 - thione -4carboxylic Acid (TTCA)
Figure 79.4 Reaction of cysteine with thiophosgene.
79.2.4.1 Rat Metabolism Captan and folpet are rapidly eliminated when administered either orally or intraperitoneally. Multiple doses of captan or folpet do not alter subsequent excretory patterns, suggesting that liver enzymes are not induced by repeated exposure. This finding was expected because the parent molecules are not likely to reach the liver. There is no sex difference in the way these fungicides are metabolized. A lO-mg/kg dose of ringlabeled captan is rapidly excreted in the urine. After 24 h, approximately 75% of the administered dose is excreted in the urine and 6.5% is excreted in the feces. Nearly all radioactivity is excreted by 36 h (Trivedi, 1990a). Fourteen repeated single doses of 10 mg/kg followed by a dose of radiolabeled captan produced a similar excretory profile (Bratt, 1990). A dose of 6-mg/kg 3SS-captan given intraperitoneally to male rats was effectively eliminated within 72 h (Couch et aI., 1977). Captan at 500-mg/kg dose resulted in a similar profile except that relatively more material was excreted via the feces. In 96 h, 68.8 and 23.1 % was excreted via the urine and feces in males and 73.4 and 25.0% was excreted, respectively, in females (Trivedi, 1990b). Folpet demonstrates a similar pattern. Administration of 10 mg/kg results in approximately 96% of the radioactivity being excreted by 24 h (90% in urine; 6% in feces). With doses 50 times higher, only approximately 69% of the administered dose is cleared by 24 h (47% in urine; 22% in feces). The imide ring is relatively stable and is excreted along with additional ring metabolites. The side chain is unstable and reacts with thiols to form mineralized products such as C02, HCl, and H2S. In addition, products of the reaction also include TTCA, dithiobis(methanesulfonic acid) and its disulfide monoxide derivative. The reactions of captan and folpet are identical with regard to the -[trichloromethylthio] side-chain reactions. The THPI generated from captan is more easily metabolized than the phthalimide from folpet. This is due to the carbonyl groups of captan that draw electrons away from the hexene double bond, creating a 8+ charge at this site, thereby promoting substitutions. Administration of phthalimide results in metabolism to phthalamic acid C~79%, in females) and phthalic acid (~7%). Less than 1% of the original phthalimide is recovered in the urine (Chasseaud et aI., 1974). Phthalamic acid accounts for 80% of the original dose when 14C_[carbonyl] folpet is given to rats (Chasseaud, 1980).
Sulfhydryl groups are intimately involved with the degradation of captan and folpet. It is therefore of interest to see their effect on GSH. Swiss Webster mice fed captan at 4000, 8000, and 16,000 ppm for 35 days had GSH levels ("soluble thiols") elevated by day 1 (Miaullis et al., 1980). The percent increase over controls ranged from 146 to 227%. Gavage treatment at a relatively high dose of captan (2000 mg/kg) induced an increase in GSH levels that was observable within 2 h of treatment, whereas a smaller dose (20 mg/kg) induced a measurable increase at 4 h (Katz et aI., 1982; Sauerhoff et aI., 1982). Folpet induced increased GSH levels were demonstrated after both dietary administration and gavage (Chasseaud et aI., 1991). Folpet, administered by gavage (7.6, 72, and 668 mg/kg) initially induced a decrease (30 and 60 min), which subsequently rebounded to a higher than normal level. The decrease was statistically significant for the 72-mg/kg dose: levels of 76, 54,82, 155, and 130% (of control values) were observed at 0.5, 1, 2, 6, and 24 h, respectively. This rebound effect was also seen at 668 mg/kg: 72,52,94, 143, and 178% at the same time periods. Diethylmaleate produced a similar pattern of GSH loss and rebound. These data demonstrate that captan and folpet cause an initial lowering of GSH levels followed by an increase due to a homeostatic rebound. The generative process for GSH exceeds the loss, and a steady state of higher GSH levels is quickly reached. 79.2.4.3 Goat Metabolism Ring labeled 4C)captan was administered to goats in gelatin capsules three times per day for 4 days. The total daily dose equaled approximately 50 ppm (Cheng, 1980). Most of the radioactivity was excreted in the urine, and the next highest excretion was via the feces. Five biochemical reactions were noted from this study:
e
1. Cleavage of the N - S linkage in the captan molecule to form THPI, either by hydrolysis or reaction with SH compounds. 2. Ring hydroxylation of THPI to form 3-0H THPI. 3. Isomerization of the 3-0H THPI to form 5-0H THPI. 4. Epoxidation ofTHPI to form THPI-epoxide, which is subsequently hydrolyzed to form 4,5-diOH HHPI. 5. Hydrolysis of THPI and/or its hydroxylated derivatives to form their corresponding THP-amic acid derivatives.
e
When radiolabeled folpet 4C-labeled on the trichloromethylthio side chain (TCM)] was administered via capsules to goats, radioactivity was recovered as expired 14C02 (~31 %; reflecting the breakdown of the TCM), in the feces (~21 %), and in the urine (~1O%; Corden, 1997). Recovered 14C02 from the expired air (ca. 31 %) reflected the breakdown of the trichloromethylthio moiety to C02. In contrast, administration of ring-labeled folpet showed the same excretion pattern as discussed for captan [i.e., more excretion via the urine (~58%)
79.3 Toxicology
1717
than the feces (~35%), with the major urinary metabolite being phthalamic acid (~49% of the dose)].
normal sloughing of the stratum corneum serves to deplete the amount available for absorption.
79.2.4.4 Hen Metabolism
79.2.4.6 Human Metabolism
Laying hens metabolize captan in a similar way as mammals (Daun, 1988a, b). When tagged with 14C on the imide ring, the identified metabolites were THPI (15.8-68.9% of the tissue radioactivity), and 3-0H, and 5-0H-THPI, which represented 2.4-26% of the radioactivity. When tagged with 14C on the trichloromethylthio moiety TTCA, dithiobis-methanesulfonic acid and its monosulfoxide were seen. Parent captan was not present in the eggs or tissues.
Humans appear to metabolize captan in a similar manner to other mammals (Krieger and Thongsinthusak, 1993). Both THPI and TTCA have been recovered after oral and dermal dosing. Comparable human studies with folpet have not been conducted, but are expected to yield similar results except for the imide ring recovered.
79.2.4.5 Dermal Absorption
Captan and folpet are extensively altered in mammalian and avian systems through a combination of enzymatic as well as nonenzymatic chemical reactions. Two complementary processes, hydrolysis and thiol interactions, initially split the fungicides into their respective imide rings and trichloromethylthio complexes. Subsequent reactions, some of which may be enzymatic, produce a series of imide-based degradates and thiophosgene-mediated products. The reactions are rapid and nearly all material is eliminated from the animal within 2448 h; there is no accumulation of either imide or side chain. Urinary metabolites from the rings differ between captan and folpet, but those associated with the trichloromethylthio side chain are common [e.g., TTCA and dithiobis(methanesulfonic acid)]. Captan or folpet do not survive in the systemic circulation, thus limiting their primary effects to areas of initial contact. Due to this rapid elimination, meat, milk, or eggs from livestock that might have consumed feed with residues of captan or folpet present would be devoid of the parent materials.
Dermal absorption of captan and folpet, as subsequently noted, is estimated at no greater than 0.5% per hour. Captan penetration of skin has been measured in vitro, comparing rat to human, and in vivo, using rats. The rat study measured absorption at 1, 2, 4, and 8 hat 19.4 J.tg/cm2 (0.5 mg/kg) and 194 J.tg/cm2 (5 mg/kg; Adir et aI., 1982), and the data have been interpreted as indicating from 0.4% per hour absorption (Ghali, 1997) to 1.5% per hour (11.7% per 8 h; Thongsinthusak et aI., 1999). The 11.7% per 8-h rate has been noted as overly conservative because the test sites were not occluded, allowing contamination of the urine and feces samples. Additionally, the absorption rate did not consider the difference between rat and human skin permeability (Fletcher et aI., 1995). The in vitro ratlhuman comparisons showed that human skin was consistently less permeable than rat skin, but that the ratio of permeability was partly dependent on the concentration of captan applied and the solvent used. A study with 14C-folpet 50 WP in rats indicated a systemic absorption of 0.27% per hour (6.5% in 24 h; Wilson and Wright, 1990). This calculation was based on a least square analysis of 24-h urinary excretion at dose levels of 49, 460, and 4800 J.tg/cm2 (13.2, 3.5, and 1.3%, respectively), matching the excretion to the approximate dermal exposure of 2400 J.tg/cm2, based on the 50WP formulation concentration. There was rapid uptake of folpet into the skin and 95 and 94% for dose levels 4800 and 460 J.tg/cm2, respectively, were retained there at 24 h. The amount still in the skin after 24 h for the low dose was approximately 85% of the applied dose. A comparison with ring-labeled captan and side-chainlabeled folpet showed no difference in absorption between adult and young Fischer 344 rats, but a lesser amount of folpet was absorbed compared to captan (Shah et al., 1987). At 0.54 and 2.68 J.trnlcm 2, captan penetration in adult rats was 3.7 and 3.6%, whereas folpet penetration was 2.7 and 1.1 %, respectively. These data were interpreted by EPA to suggest 0.4% per hour (4.29% in 10 h) dermal absorption for folpet (Ghali, 1997; Levy et aI., 1997). These data show high dermal adsorption, but low penetration rates for captan and folpet. There are differing interpretations of these data, but it is reasonable to conclude that the hourly dermal penetration rate is no greater than 0.5%. The
79.2.5 SUMMARY
79.3 TOXICOLOGY 79.3.1 ACUTE TOXICOLOGY 79.3.1.1 Overview Both captan and folpet have low acute toxicity, except for the intraperitoneal route (Table 79.3). The reactivity to mucus membranes is high and the severe eye irritant finding is consistent with this property. When single 0.5-g doses are applied to the skin, the results show mild to low irritancy. Results of guinea pig sensitization studies give both positive and negative results; however, experience with handlers of these materials suggest that some persons (less than 10% of the population) are susceptible to sensitization. 79.3.1.2 Acute Oral Toxicity The low acute oral toxicity of captan and folpet reflects the rapid degradation of the fungicides once ingested and the absence of intact fungicide at sensitive biochemical targets. The LD50 values are above 5 g/kg for both technical and formulated
] 718
CHAPTER 79 Captan and FoIpet
Table 79.3 Acute Toxicity of Captan and Folper" Parameter
Captan
Reference
Folpet
Reference
Oral LDSO, rat
>5 glkg
Gaines and Linder
>5g/kg
Gaines and Linder
>10 g/kg
Ben-Dyke et al.
(1986)
(1986) 8 g/kg 50W fonnulation Intraperitoneal LDSO, rat Dennal LDSO,
40mglkg, M
(1970)
50W fonnulation
(1970)
Copley (1985)
48-52.5 mg/kg
Dickhaus and Heisler
>2000 mg/kg
Thoa and Redden
>5000 mg/kg
Korenaga (1982)
Severe
Thoa and Redden
Severe
V.S. EPA (1987)
(1983)
35 mg/kg, F (1995)
rabbit Irritation,
Ben-Dyke et al.
(1995)
eye Skin
Minimal
V.S. EPA (1975)
Minimal
V.S. EPA (1987)
Inhalation LCSO,
0.72-0.87 mglliter
Thoa and Redden
1.89 mglliter
Cracknell (1993)
Moderate
V.S. EPA (1987)
(1995)
rat,4h exposure Sensitization,
Moderate
guinea pig
Thoa and Redden (1995)
aM, males; F, females.
products, placing them in Toxicity Category IV for EPA regulatory purposes. The captan 50W formulation LD50 is 8.4 g/kg, whereas the comparable folpet formulation LDso is greater than 10 g/kg (Ben-Dyke et aI., 1970). Results from other investigators consistently show LDso values above 5 g/kg for both compounds (Boyd and Krijnen, 1968; Nelson, 1949). The imide ring degradates of captan and folpet, THPI and phthalimide, respectively, are stable compared to the parent compounds. Accordingly, some measure of their acute toxicity is in order. Both these compounds have low acute oral toxicity in mammals. The LDso of THPI is 2 g/kg (Cavalli, 1970); for phthalimide it is greater than 8 g/kg (U.S. EPA, 1974). In contrast to mammals, aquatic organisms are particularly sensitive to captan and folpet, and offer a useful test system to compare the toxicity of parent and degradate. Comparative toxicity in trout show THPI to be approximately 3500-fold less toxic than captan (captan LCso = 34 ppb versus THPI LCso > 120,000 ppb; U.S. EPA, 1999a). A similar comparison for folpet shows a 3267-fo1d difference (folpet LCso = 15 ppb versus phthalimide LCso = 49,000 ppb; U.S. EPA, 1999b). 79.3.1.3 Acute Intraperitoneal Toxicity
Intraperitoneal administered acute toxicity studies are not generally required for regulatory purposes because this route of entry affords little information for human risk assessment. The intraperitoneal route bypasses the intestine, although materials are still primarily absorbed by the portal system (Rozman and Klaassen, 1996). In the case of oral administration of captan or folpet, only trace amounts of parent material would enter the portal system and would be rapidly degraded. Intraperitoneal administration bypasses this degradation and thus affords an observation into the inherent toxicity of the materials.
Male and female SPF Wistar rats treated with 92.7% captan administered by injection in 1% methylcellulose had LDso values approximately 40 mg/kg for males and 35 mg/kg for females (Copley, 1985). This LDso, lower by over lOO-fold compared to oral administration, indicates the inherent toxicity of captan. It also demonstrates the effective barrier provided by the intestine . Male and female Wistar rats injected with 87.5% folpet in 1% methylcellulose had a 24-h LDso of 48.0-52.5 mg/kg. Deaths occurred between 24 hand 7 days, resulting in a 7-day LDso of 36-40 mg/kg. There was a steep dose-response curve: at 30 mg/kg, 1 in 10 deaths were seen, but at 60 mg/kg, lOin 10 deaths occurred by day 7 (Dickhaus and Heisler, 1983). The intraperitoneal LDso values for captan and folpet are similar and reflect the rule that governs their toxicological profile: hazard in the absence of exposure limits adverse effects. 79.3.1.4 Acute Dermal Toxicity
Captan and folpet pose little hazard of acute toxicity from dermal exposure. Limit doses of 2 g/kg are without effect in rabbits (Foster and Morgan, 1984; Gaines and Linder, 1986). A study with Phaltan (folpet 50%) showed that the LDso was greater than 22.6 g/kg (Kay and Calandra, 1960). 79.3.1.5 Eye Irritation
Captan and folpet irritate mucus membranes and there is the potential for damage when they contact the eyes. Bioassays, however, vary in their estimation of this hazard. Captan eye irritation studies show variable results. Minimal damage, as noted by no corneal or iris involvement, and low redness and swelling to the eyelids, has been reported (Harris, 1976). Conversely, severe damage, including corneal opacity, has also been reported
79.3 Toxicology (Rosenfeld, 1984). Washing the treated eyes after instillation of test material reduces irritation, as was observed in a study that employed a captan 50W formulation (Sauer and Seaman, 1980). An additional study employed a combination formulation that included captan (8%), folpet (44%), and captafol (8%), and showed conjunctival irritation, but no corneal involvement (Cisson et aI., 1983). Folpet Technical, in unwashed eyes, induced transient corneal opacity that progressed to vascularization of the cornea in two of six rabbits (Dreher, 1992a). A 100-mg instillation of Folpet Technical caused corneal opacity in some unwashed eyes and no opacity in eyes that were washed 30 s after instillation (Cisson et al., 1982). All eyes returned to normal by 7 days (washed) or 10 days (unwashed). Phaltan 500 Flowable formulation (50% folpet) instilled in rabbit eyes, followed by a 30 s wash after 30 s, resulted in no corneal opacity and minimal redness and swelling (Mercier, 1988). By 72 h, all swelling had subsided, but there was some residual redness and congestion.
79.3.1.6 Skin Irritation Both captan and folpet elicit very little irritation when applied as a single dose to either intact or abraded skin. Captan Technical applied to rabbit skin at 0.5 g showed no redness or edema at either 24 or 48 h for both intact and abraded test sites (Harris, 1976). Folpet Technical applied to rabbit skin at 0.5 g showed no redness or edema at observation periods up to 72 h (Rees, 1993). Doses offolpet as high as 22.6 g/kg produced only transient redness in rabbits (Kay and Calandra, 1960).
79.3.1.7 Acute Inhalation Toxicity Acute toxicity via the inhalation route of exposure varies somewhat with the specific formulation tested. For captan, the 4-h LCso was 1.21 mg/liter for males and 1.05 mg/liter for females (Cummins, 1995). An earlier study reported these values as 0.90 mg/liter for males and 0.67 mg/liter for females (Blagden, 1991). For folpet, the 4-h LCso for males and females was 1.89 mg/liter (95% confidence limits 1.47-2.31 mg/liter; Cracknell, 1993). The particle size mass median equivalent aerodynamic diameter was 4.6-5.2 ).Lm. Males were slightly more susceptible than females. The mortality for males was 0 in 5, 3 in 5, and 4 in 5 for dose levels 0.80, 1.60, and 1.99 mg/liter, respectively. Females at these respective doses showed mortality of 0 in 5, 1 in 5, and 1 in 5.
79.3.1.8 Skin Sensitization Guinea pig bioassays show that both captan and folpet have the potential to induce delayed contact hypersensitivity reactions. Captan Technical was positive in the Magnusson and Kligman maximization guinea pig assays (Dreher, 1992b). Folpet Technical was tested using the Magnusson and Kligman protocol in which a 10% w/v preparation in propylene glycol was injected by the intradermal route along with a 1: 1 preparation of Freund's Complete Adjuvant. Subsequent topical inductions were made with 50% w/v folpet. Challenge with either 10 or 50%
1719
w/v folpet in propylene glycol resulted in positive reactions (LSR, 1993).
79.3.1.9 Human Experience Reports of adverse effects are limited to incidences of skin irritation or sensitization reactions (Guo te aI., 1996; Peluso et aI., 1991; V.S. EPA, 1989). In human patch tests, 9 in 205 (4.4%) subjects showed a positive Draize reaction (Marzulli and Maibach, 1973),8 in 150 (5.3%) were sensitized by 1% topically applied captan (Jordan and King, 1977), and 24 in 200 were positive to chemicals in the fungicide category, predominantly the thiophthalimides (Lisi et aI., 1986). Additionally, a case of urticaria was associated with use of a captan-formulated product (Croy, 1973). However, a powder blush that contained 0.3% captan failed to sensitize any of 25 adult volunteers on which it was tested. This was in spite of the study design, which included repeated doses (five consecutive 300-mg induction exposures to the forearms) and occlusion of the application site for 48 h after each dose. The individuals were challenged 10 days later and all individuals were negative (Ivy Research Laboratories, 1981). Regular exposure to captan, formulated in a shampoo at 7%, appears to be well tolerated (Guo, 2001). Although there is potential for eye irritation, based on laboratory studies, there are no reports in the literature of adverse effects (NLM, 2001). Likewise, eye injuries in agricultural workers who carry out reentry activities do not appear to be problematic (Krieger, 2001).
79.3.1.10 Summary Adverse skin reactions to captan and folpet due to delayed contact hypersensitivity are possible for mixers, loaders, and applicators, and may occur in low incidence. The potential for eye irritation exists, but extensive use experience suggests this problem is minimal. There is little acute risk from oral ingestion or dermal exposure to either product. The World Health Organization has classified folpet as unlikely to present an acute hazard in normal use (FAOIWHO, 1996).
79.3.2 SVBCHRONIC TOXICITY Mechanistic studies in mice have focused on changes to the duodenum and illuminate the mode of action for tumor formation. These studies confirm the irritant properties of captan and folpet, the effects of irritation to the duodenum, and the reversibility of these effects upon cessation of treatment. Other observations in the mouse include reduced weight gain and depressed food intake at high doses; this is also seen as a general secondary effect in rat studies. Observations in rats also include hyperkeratosis and acanthosis of the esophagus and stomach, particularly for folpet. Dogs do not tolerate capsuleadministered captan or folpet well; emesis is always seen. Rabbits administered folpet dermally show marked skin irritation, and rats respond to repeated dermal application of folpet with severe skin irritation.
1720
CHAPTER 79
Captan and Folpet
79.3.2.1 Mice Mechanistic Studies Male CD-l mice (four or five per group) treated with 3000-ppm captan for 1, 3, 7, 14, or 28 days showed shortened duodenal villi due to damage by captan. This effect was observable in the crypts within 3 days of treatment initiation (Tinston, 1996). Immature cells were observed at the villi tips from day 7 onward, indicating a higher turnover of cells. There was some focal gastritis and parakeratosis noted in one mouse. When captan is fed to mice at 6000 ppm for 28-90 days, villus atrophy occurs, together with a crypt hypertrophy and crypt cell hyperplasia (Tinston, 1995). CD-l mice administered captan for 56 days as 0, 400, 800, 3000, or 6000 ppm were evaluated for proliferative changes in the duodenum (Tinston, 1995). An assessment of the duodenum was made using histopathology and a bromodeoxyuridinelabeling index to measure crypt cell proliferation. Captan induced hyperplasia of the crypt cells, an increase in the crypt celllabeling index, and an increase in the number of cells in the crypt cell population. At 3000 ppm, the villus-to-crypt height ratio decreased from 5.4 in males and 5.9 in females to 1.4 and 2.6, respectively. These observed changes are consistent with an irritant action of captan on the duodenum. The no observed effect limit (NOEL) for duodenal hyperplasia is 400 ppm. Male CD-l mice treated with 5000-ppm folpet for 28 days (Waterson, 1995) were observed to have proliferative changes in the duodenum proximal to the pyloric sphincter. An inflammatory response, similar to that noted with captan, was not seen. Treatment of CD-l mice with folpet at 0, 150,450, and 5000 ppm for 28 days resulted in duodenum proliferative effects (Milburn, 1997). Villi length was reduced and crypt compartments were expanded, reducing the villi-to-crypt ratio. The NOEL for hyperplasia in the duodenum for males was 450 ppm (69 mg/kg per day) and the NOEL for females was 150 ppm (29 mg/kg per day).
in females (291 mg/kg per day). Irritation to the esophagus and forestomach occurred at all doses and in both sexes. A variety of hematological and clinical chemistry changes were noted, but the incidence and pattern did not indicate a clear target. Folpet Technical admixed in the diet at 0, 300, 100, 3000, or 10,000 ppm and fed to Sprague-Dawley rats for 13 weeks, followed by a 2-week recovery showed similar signs of irritation in the forestomach, primarily at 10,000 ppm, but no irritation of the esophagus (Reno et aI., 1981). Following a 2-week recovery period during which the rats received control diet, the forestomach histology returned to normal. Folpet Technical fed to B6C3Fl mice for 28 days at levels of 0, 1000, 5000, and 10,000 ppm induced a reduced body weight gain in the top two doses (Crown, 1981).
Dermal A 28-day rat study with Folpet Technical applied in mineral oil at 0, 1, 10, 20, and 30 mg/kg per day to the backs of Sprague-Dawley rats 6 h per day, five days per week. All dose levels elicited irritation that was more severe in males than females and resulted in decreased weight gain (Dougherty, 1988). The irritation was so severe in the 30-mg/kg per day male group, that application was terminated after 10 days. All adverse effects noted were related to the skin irritation induced by folpet. Inhalation Captan has been tested in Wistar rats by nose-only inhalation at nominal dose levels of 0.1, 0.5, 5, and 15 !-ig/liter for 13 weeks (Hext, 1989). There were deaths in males at the high dose and dose-related effects on the larynx (e.g., squamous metaplasia, squamous hyperplasia, vacuolar degeneration of squamous epithelium). The no observed effect concentration (NOEC) for toxic effects (other than generalized irritation) was 0.6 !-ig/liter (measured). 79.3.2.3 Rabbits
Subchronic Study B6C3Fl mice were fed folpet at 0, 1000, 5000, or 10,000 ppm for 4 weeks (Rubin, 1981). There was reduced food intake and body weight gain at 5000 ppm. 79.3.2.2 Rats Oral Wistar rats treated for 4 weeks with captan at 0, 2000, 4000, 8000, or 12,000 ppm were observed to have a doserelated decrease in body weight gain and food intake (Til and Beems, 1979). At the top two doses, there were increases in basophilia of hepatocytes in the periportal area of the liver accompanied by increases in relative liver weight. A similar finding was observed in the females at the next lower dose (4000 ppm). The relative organ to body kidney weights were statistically increased at all doses. Folpet Technical admixed in the diet at 0, 2000, 4000, and 8000 ppm, and fed to Fischer rats for 13 weeks produced a treatment and dose-related decrease in body weight gain and hyperkeratosis/acanthosis of the esophagus and nonglandular stomach (Sela, 1982). The NOEL for decreased weight gain was 2000 ppm in males (136 mg/kg per day) and 4000 ppm
A 21-day dermal study with captan in rabbits at 0,12.5,110, or 1000 mg/kg per day (6-h exposure per day) resulted in a dose-related desquamation of the skin by day 21, erythema and edema at the high dose, and acanthosis and hyperkeratosis of the treated skin at all doses (John son, 1987).
79.3.2.4 Dogs Beagles, two per sex per treatment group, were administered captan at 0, 30, 100, 300, 600, or 1000 mg/kg per day for 4 weeks. The results included treatment-related emesis and a dose-related decrease in food intake and body weight gain (Blair, 1987). There was an increase in relative liver weight in males at 600 and 1000 mg/kg per day and relative kidney weight in females at 1000 mg/kg per day. Some fatty changes were seen in the kidney and liver of one male at 1000 mg/kg per day. Folpet administered to two beagle dogs per sex per group at 0, 20, 60, 180, and 540 mg/kg per day for 4 weeks induced emesis (Daly, 1983). Food intake and body weights were reduced in
79.3 Toxicology
a dose-related manner. There were, however, no histopathologic changes noted. Folpet was administered to four beagles per sex at 0, 790, 1800, and 4000 mg/kg per day for 13 weeks with gelatin capsules (Barel et aI., 1985). Daily doses of 4 g/kg were well above the maximum tolerated dose and resulted in severe deterioration of the males, all of which were killed for humane reasons. One of the females at the high dose was also terminated in moribund condition. Vomiting and diarrhea were noted clinical signs and both food intake and body weight gains were reduced in a doserelated fashion. Treatment resulted in irritation of the gastric mucosa, atrophied testes, thyroid degeneration, and muscular dystrophy.
1721
Table 79.4 Captan Mechanistic Study Designa Time (months) Dose (ppm) 0
3
6
9
12
18
20
S
S
S
S
S
S
RS
RS
6000
T,S
6000
T
T,S
6000
T
T
T
T,S
RS
6000
T
T
T
T
T,S
6000
T
T
T
T
T
RS T,S
aReproduced with permission from Pavkov and Thomasson (1985). T = treatment; S = sacrifice; RS = recovery sacrifice.
79.3.2.5 Miscellaneous Studies Rats and mice administered 3000-ppm captan had reduced immune function as measured by sheep red blood cell antibody formation after treatment for 42 days (LaFarge-Frayssinet and Decloitre, 1982). Captan was reported to suppress both B- and T-cell function in mice. Wistar rats had depressed lymphocyte count and lower relative thymus weight after 3 weeks of dietary administration of captan at 1000 ppm [50 mg/kg body weight (bw) per day], initiated as weanlings (Vos and Krajnc, 1983). Additionally, rats administered pre- and postnatal captan at 750 and 2000 ppm (37.5 and 100 mg/kg bw per day) showed a decrease in secondary IgG response to tetanus toxoid in the high-dose animals (Vos and Krajnc, 1983). Evaluation of these studies by Joint Meeting on Pesticide Residues (JMPR) concluded that captan may be an immunodepressant (FAOIWHO, 1990). These dose levels, however, are high and may not be relevant to anticipated human exposure scenarios.
79.3.3 CHRONIC TOXICITY One mechanistic study was conducted in mice with captan. These data are also relevant for folpet because both share a common mechanism of toxicity. Duodenal tumors in mice were seen in both chronic and oncogenic studies (discussed later). Chronic administration of folpet produced evidence of irritation to the esophagus and stomach in rats.
79.3.3.1 Mice In a mechanistic study, male CD-1 mice were administered captan at dietary doses of 0 and 6000 ppm for 3, 6, 9, 12, 18, or 20 months (Pavkov and Thomasson, 1985). Mice were examined at the end of each dosing period and, in addition, various recovery periods were evaluated (Table 79.4). One group dosed with 6000 ppm for 6 months was held for an additional6-month recovery period; another was held for a 12-month recovery period. One group dosed with 6000 ppm for 12 months was followed after a 6- and 8-month recovery. The most characteristic pathologic findings consisted of necrotizing and proliferative changes in the nonglandular portion of the stomach (after 3 months), dilation of the small intestine, and focal epithelial
hyperplasia in the proximal part of the small intestine. Focal epithelial hyperplasia was also found in controls, but the incidence was lower compared to that of the treated animals, and the localization of these foci was more caudal than was the case for captan-administered mice. Diffuse hyperplasia was found only in treated mice and was not considered prerequisite for the development of focal hyperplasia. Adenomas and adenocarcinomas also developed in the small intestines of treated mice with localization in the proximal 7 cm of the small intestine; the area of localization was the same as for the focal hyperplasia. Removal of captan from the diet resulted in a significant reduction in the incidence of focal epithelial hyperplasia as compared to the incidence in concurrent lifetime-treated mice, and was no greater than that in concurrent controls. The incidence of neoplasia, however, in mice in the recovery group was not significantly different from that of concurrent lifetime-treated mice, but increased in mice treated for 6 months with a recovery period of 6 months and in mice treated for 12 months with a recovery period of 6-8 months, respectively, when compared to controls. The latter increase was not found in mice treated for 6 months with a recovery period of 12 months.
79.3.3.2 Rats Rats were administered folpet at dietary concentrations of 0, 250, 1500, and 5000 ppm (Crown et aI., 1989). Body weight gain and food intake were decreased at 5000 ppm. The incidence and severity of diffuse hyperkeratosis in the esophagus and nonglandular epithelium of the stomach were increased in both sexes at 5000 ppm. The stomach was also affected at 1500 ppm. The NOEL was 250 ppm (12 and 15 mg/kg per day, males and females, respectively). A folpet combined chronic toxicity/oncogenicity study from which the EPA derived a NOEL for use in chronic dietary risk assessment employed dietary dose levels of 0, 200, 800, or 3200 ppm (Cox et aI., 1985). The EPA selected the 200-ppm level (9 mg/kg per day) as the NOEL for conducting chronic dietary risk assessments, based on hyperkeratosis/acanthosis and ulceration/erosion of the nonglandular stomach at 800 ppm (equivalent to 35 mg/kg per day; U.S. EPA, 1999b).
1722
CHAPTER 79
Captan and Folpet
79.3.3.3 Dogs
Dogs were treated with captan at 0, 12.5,60, and 300 mg/kg per day for 1 year (Blair, 1988). Only the high-dose animals differed from control in increased incidence of emesis and soft stool, increased relative liver weight, and decreased total serum protein and albumin. There were two I-year dog studies conducted with folpet: the first at 0, 10,60, and 1401120 mg/kg per day (Daly and Knezevich, 1986) and the second at 0,325,650, and 1300 mg/kg per day (Waner, 1988). In the first study, the 140 mglkg per day was reduced to 120 mg/kg per day on day 50 due to unacceptable decreases in body weight gain and food intake. A NOEL was selected for the study at 10 mg/kg per day based on lowered body weight gain and food intake at 60 and 120 mg/kg per day. There were no clinical signs of toxicity noted, but clinical chemistry values showed a treatment-related decrease in total plasma protein parameters and cholesterol. Organ weights were not affected by treatment nor was there evidence of macroscopic or microscopic changes as a result of treatment. In the second study, folpet induced incidences of diarrhea, vomiting, and salivation that were associated with reduced food intake and reduced body weight gain. Testes weights were reduced in males administered 1300 mg/kg per day when compared to controls on an absolute basis, but were similar to controls when measured on a relative body weight basis. The NOEL for this study was 325 mg/kg per day, based on decreased body weight gain. The WHO acceptable daily intake (AD!) is 0.1 mg/kg per day for both captan (FAOIWHO, 1990) and folpet (FAOIWHO, 1996). 79.3.4 DEVELOPMENTAL AND REPRODUCTIVE TOXICITY 79.3.4.1 Developmental Studies Rats Captan administered to Sprague-Dawley CD rats at 0, 18, 90, and 450 mg/kg per day resulted in decreased maternal weight gain and decreased food consumption at the high dose (Rubin, 1987). There were no effects on postimplantation loss or fetal survival. Fetal body weight was reduced and the incidence of "small" fetuses «3.0 g) was increased at the high dose. There were no increases in incidences of treatment-related malformations. The incidence of minor skeletal variations, including the presence of a fourteenth (lumbar) rib, incomplete fusion of vertebral hemicentra fusion, and reduced ossification of the pubes was increased at 450 mg/kg per day. The NOELs for maternal and developmental toxicity in this study were 18 and 90 mg/kg per day. In another study, folpet was administered by gavage to Sprague-Dawley CD rats at 0, 150, 550, or 2000 mg/kg per day from gestation days 6-15. Maternal toxicity in the form of decreased food intake and body weight gain was observed at the mid- and high dose. Fetuses showed slight developmental retardation at 150 mg/kg per day, suggesting the NOEL was slightly below this level (Rubin, 1983). Pups from rats treated with 400
mg/kg per day from gestation days 8-10 were normal (Kennedy et aI., 1968) as were rats treated with 360 mg/kg per day from gestation days 6-19 (Hoberman et aI., 1983). Rabbits Captan did not induce any teratogenic effects when administered to New Zealand White (NZW) rabbits at 0, 10, 40, and 160 mg/kg per day from gestation days 7-19 (Rubin and Nyska, 1987). The highest dose was toxic to both dams and fetuses. An increased incidence of minor skeletal variations was seen at this dose. In another study, captan was administered to New Zealand White rabbits at 0, 10, 30, or 100 mg/kg per day (Tinston, 1991). The developmental NOEL was 10 mg/kg per day based on increased postimplantation loss, reduced mean fetal weight, and increased skeletal defects in fetuses (27 presacral vertebrae) at the maternally toxic dose of 30 mg/kg per day. Folpet was tested in both Dutch Belted and NZW rabbits for potential developmental toxicity. The original studies (Fabro et aI., 1966; Kennedy et aI., 1968; McLaughlin et aI., 1969) were performed at high doses ranging from 75 to 150 mg/kg per day during gestation days 7-12, 6-16, or 6-18. These studies consistently demonstrated the absence of adverse effects. One study reported five incidences of hydrocephaly at doses that were maternally toxic (Feussner et aI., 1984; one at 20 mg/kg per day and four at 60 mg/kg per day). A second study in which doses of folpet were "pulsed" failed to replicated this finding (Feussner, 1985). The most recent study employed doses of 10,40, and 160 mg/kg per day, and confirmed the absence of folpet-induced developmental effects (Rubin, 1985b). Although a weight-of-evidence (WOE) analysis concluded folpet is not a developmental toxin, the EPA has assigned this compound an FQPA uncertainty factor of 3 based on the initial Feussner study. Although the EPA pointed to one study where hydrocephaly occurred at maternally toxic doses of 20 (one instance) and 60 mg/kg per day (three fetuses in two litters; Feussner et aI., 1984), a second "pulse dose" study (Feussner, 1985) failed to replicate this finding and other developmental studies in NZW rabbits showed no evidence of teratogenicity or hydrocephaly (Fabro et aI., 1966; Kennedy et aI., 1968; McLaughlin et aI., 1969; Rubin, 1985b). The EPA cited the Feussner study as the basis for assigning an FQPA three-fold uncertainty factor (U.S. EPA, 1999b), although a WOE analysis concluded that folpet is not a selective developmental toxin (Neal, 2000). Mice CD-1 mice administered fo1pet by gavage, subcutaneously (100 mg/kg per day), or by inhalation at (624 ).lg/m3) showed no developmental abnormalities (Courtney et aI., 1983). BL6 mice treated subcutaneously and orally with folpet at 100 mg/kg per day and AKR mice treated subcutaneously at 100 mg/kg per day were judged to have not adverse developmental findings (Bionetics Research Laboratories, 1968). Other Species Captan is not teratogenic in beagles when administered in the diet at 60 mglkg per day either through-
79.3 Toxicology out gestation or throughout gestation plus lactation (Kennedy et aI., 1975b). Folpet was studied in Rhesus and stumptailed macaques as part of research on thalidomide (Vondruska, 1969). There were no malformations with folpet at doses up to 75 mg/kg per day. Thalidomide at 10 mg/kg per day produced limb defects. 79.3.4.2 Reproductive Studies In a three-generation study, COBS CD rats were treated with captan at 25, 100, 250, and 500 mg/kg per day (Schardein et aI., 1982). Nonreproductive parental toxicity was seen at 100 mg/kg per day and above in the absence of reproductive effects. Pup weights were lower by 7% compared to controls at 25 mg/kg per day. A subsequent one-generation study at 0, 6, 12.5, and 25 mg/kg per day showed no effect on pup weights at 25 mg/kg per day. The NOEL selected by the EPA for use in risk assessment was 12.5 mg/kg per day, based on the weight gain depression in the three-generation study (Ghali, 1997; Schardein and Aldridge, 1982). The reference dose employed by the EPA, 0.13 mg/kg per day, is based on this NOEL and the use of a 100-fold safety factor. Folpet administered to rats at 0, 250, 1500, and 5000 ppm showed diffuse hyperkeratosis of the nonglandular epithelium of the stomach (Rubin, 1986). The NOEL for this study was 250 ppm, which averaged 24 mg/kg per day. There were no reproductive effects noted. Other two- or three-generation studies in rats with folpet also showed no adverse reproductive effects (Hardy, 1985; Kennedy, 1967). 79.3.5 MUTAGENICITY 79.3.5.1 Overview The issue of mutagenicity is controversial. Throughout this chapter, the rapid degradation of captan and folpet in living systems has been central to understanding their toxicology. The pattern of mutagenicity is consistent with this degradation and
provides examples of how such degradation diminishes adverse effects. Interest in the mutagenicity of captan and folpet has produced many mutagenicity studies. In vitro studies showed clear evidence that both are mutagenic, captan is more potent compared to folpet, and their activity is inversely proportional to the presence of thiols in the reaction vessels. Once thresholds for complete degradation are reached, in vitro activity is abolished. The large reserves of thiols present in the intact animal and the near instantaneous reaction of captan and folpet with the thiols serves to ensure complete elimination of captan and folpet before they reach sensitive DNA targets. The net result of this rapid degradation is the absence of mutagenicity in vivo. The mechanism by which captan and folpet effect their mutagenicity is not clear; however, data suggest that thiophosgene, in addition to the parent compounds, is mutagenic (Arlett et aI., 1975). For in vitro systems, both frame-shift and base-pair substitutions are seen. Cytogenetic effects are seen in vitro, but positive results are not as ubiquitous as point mutations. These clastogenic effects are reduced when enzyme enhanced rat liver extract (S-9) is present and are generally absent in vivo. The weight of evidence shows that although captan and folpet possess inherent mutagenic potential, they are not mutagenic in vivo. 79.3.5.2 Mutations In Vitro Assays Table 79.5 shows results from representative in vitro assays with S. typhimurium, strain TA 100, a prokaryote organism. The greater potency of captan relative to folpet is noted. Other S. typhimurium strains showed similar results. A mutation index (the ratio between induced versus spontaneous revertants) of 7.3 for captan and 6.3 for folpet was seen for strain 104 (Barrueco and de la Pena, 1988). Positive findings were generally seen with strains 98, 1535, 1537, and 1538 (Carere et aI., 1978; Shiau et aI., 1981; Shirasu et aI., 1976), but negative findings were seen with strain 1536 (Shiau et aI., 1981). Where there were marginally positive results with captan, folpet was usually negative. Strain WP2 try-hcr+ of E. coli
Table 79.5 Prokaryote Reverse Mutation: Salmonella typhimurium, Strain TA lOO Compound
Resultsa
Reference
Captan
26.7 revertants per 108 cells/nmol, -S-9
Shiau et al. (1981)
++ Folpet
@
50 fig/plate (167 nmol), +S-9
+ @ 50 fig/plate (167 nmol), +S-9 7.7 revertants per 108 cells/nmol, -S-9 +
@
50 fig/plate (167 nmol), -S-9
-
@
50 fig/plate (167 nmol), +S-9
Captan
26 revertants/nmol, -S-9
Folpet
8 revertants/nmol, -S-9
Captan
93.7 revertants/nmol, -S-9
Folpet
15.0 revertants/nmol, -S-9
Shiau et al. (1981)
De Flora et al. (1984) Moriya et al. (1983)
a +S-9: Rat liver homogenate included for "metabolism" of test material. -S-9: Incubation without rat liver
homogenate.
1723
1724
CHAPTER 79
Captan and Folpet
Table 79.6 In Vitro Eukaryote Mutation Assays Assay
Compound
Chinese hamster
Captan
Results
Reference
Positive only in the absence of serum from the culture media
ArIett et al.
Mean number of resistant colonies: 0.3 and 0.6 at 5 and
V79IHgprt
(1975)
10 ).tg/ml captan; vapor emitted from sodium bicarbonate activated captan impregnated on filter paper above the test system also induced mutations Chinese hamster
Captan and
CHOlHgprt
Folpet
Mouse
Captan
Both compounds were positive in the absence of S-9
O'Neill et al.
Positive in the absence of S-9
OberIyeta!.
(1981)
lymphoma L5178yrrK
(1984)
showed a strong response to captan and a negative response to folpet (Nagy et aI., 1975) or, where both were positive, the revertants per plate were greater for captan than for folpet (Shirasu et aI., 1976). Both were positive with the WP2 try-hcrstrain (Nagy et aI., 1975; Shirasu et aI., 1976) as well as other tests with the WP2 strain (Bridges et aI., 1972; Simmon et aI., 1976). Tests with Bacillus subtilis strains TK 6321 and 5211 were positive for both compounds: captan showed a greater mutagenic response than folpet (Shiau et aI., 1981). Captan also induced point mutations in Aspergillus nidulans (MartinezRossi and Azevedo, 1987). Assays with eukaryote organisms such as Chinese Hamster cells and mouse lymphoma cells are shown in Table 79.6. These data show that captan and fo1pet induce mutations when measured in vitro. THPI was tested with S. typhimurium strains TA 98, 100, 1535, and 1202 as well as E. coli WP2 uvrA, and
was negative (Carver, 1985). Phthalimide is inactive in S. typhimurium as well (Rideg, 1982). Effect of Exogenous Thiols on Mutagenicity Assays The rat liver S-9 fraction is added to in vitro systems to simulate the metabolic capability of intact organisms. In this way compounds that are mutagenic only after they are metabolized by cell enzyme systems are detected. Metabolism, however, appears to play no role in the expression of mutagenicity of captan or folpet; on the contrary, the addition of S-9 serves to diminish mutagenic potency. The reduced mutagenic activity following the addition of S-9 is an example of the general phenomenon of thiol-re1ated degradation of captan and folpet. The presence of sufficient thiols abolishes mutagenic activity. The addition of S-9 or rat blood prior to the addition of captan or folpet reduces or abolishes activity (Table 79.7).
Table 79.7 Comparison of Captan and Folpet Mutagenicity with the Addition of Exogenous Proteina Captan
Folpet
Component
(rev/plate)
(rev/plate)
Commentb
None
3200
1320
Captan is more active than folpet
30
50
S-9 decreases activity
111
60
S-9 fraction decreases activity
19
21
Cysteine abolishes activity (control
32
19
Blood abolishes activity
268
219
6
8
S9 abolishes activity (control
31
35
S-9 fraction decreases activity
4
6
6
14
Strain and Dose E. coli
WP2hcr
S-9
0.I5).tM/plate
S-9 fraction
(45 ).tg/plate)
20-mM
rev/plate < 30)
cysteine Rat blood
S. typhimurium
None S-9
Captan is more active than folpet rev/plate <17)
TA 1535
S-9 fraction
0.15 ).tM/plate
20 mM
(45 ).tg/plate)
Cysteine abolishes activity
cysteine Rat blood
Blood abolishes activity
a Reproduced from Moriya et al. (1978). bCaptan (0.1 ml of 1.5 ).tMlml) or folpet (0.2 ml of 0.75 ).lMIml) was incubated for 10 min at 37°C with 0.5 ml of one of the following: S-9 (containing 0.3 ml S-9/ml), S-9 fraction (S-9 mix minus cofactors), 20-mM cysteine, or rat blood diluted twice with phosphate buffer or water as control. After incubation the tester strains (0.1 ml) and agar (2 ml) were added to the test tubes and plated out. Revertants/plate were read after incubation at 37°C for 2 days.
79.3 Toxicology
1725
Table 79.8 In Vivo Mutagenicity Assays Results
Reference
Assay
Compound
Somatic cell mutation
Captan
Negative, oral
Nguyen (1981)
(mouse spot test)
Captan
2.2% frequency after
Imanishi et al. (1987)
intraperitoneal dose of 15 mg/kg Somatic mutation and
Folpet
Negative, oral
Moore (1985)
Captan
Negative
Mollet and Wurgler (1974)
Captan
Negative
Simmon et al. (1977)
Captan
Negative
Folpet
Negative
recombination (Drosophila SMART test) Mouse heritable translocation assay Drosophila sex-linked recessive lethal assay
Captan
Negative
Captan
Weakly mutagenic
Folpet
Weakly mutagenic
Folpet
Negative
Mouse dominant
Captan
Negative
lethal assay
Folpet
Negative
Rat dominant lethal assay
Captan
Negative
Folpet
Negative
Captan
Negative
Folpet
Negative
Kramers and Knaap (1973) Mollet (1973) Valencia (1981) Vogel and Chandler (1974) Jorgenson et al. (1976) Kennedy et al. (1975a) Rideg (1982) Epstein et al. (1972)
Folpet
Negative
Captan
Negative
Simmon et al. (1977)
Captan
Positive
Collins (1972a)
Folpet
Positive
Collins (I 972b)
Folpet
Negative
Bradfield (1980)
When cysteine is added to either captan or folpet in varying ratios, the mutagenic activity declines as the ratio increases from O.S to 2.S. At a ratio of S-!l-m cysteine to l-!l-m captan or folpet mutagenic activity is abolished (Moriya et al., 1978). Glutathione provided similar protective actions when added to assay vessels in ratios of 1 or higher compared to the fungicide (Rideg, 1982). Adverse toxicity as well as mutagenicity in Chinese hamster V79 cells is reduced when 10% fetal calf serum is used in the standard V79IHgprt assay (Arlett et al.,
to assay vessels serves to detoxify captan and folpet as it does in the mutation assays. This action is expressed by a decrease in cytotoxicity with a resulting increase in tolerated dose levels. At some point it is expected that the threshold for detoxification is exceeded and the remaining captan or folpet can act to affect the chromosomes. These data are mixed in that some positive and some negative findings are reported.
1975).
In Vivo Assays Table 79.10 lists representative in vivo cytogenetic studies with captan or folpet. Micronucleus assays were conducted in CD-1 mice with both compounds and yielded negative results. Captan was administered at 40, 200, and 1000 mgikg (Jacoby, 1985b) and folpet was administered at 10, SO, and 2S0 mgikg (Jacoby, 1985a). Chlorambucil, the positive control, resulted in a significant increase in micronuclei. These negative results contrast with positive findings by Chinese investigators who treated mice with captan (96.S% purity) at 10, SO, 100, 400, and 800 mgikg by gavage and reported a dose-related increase in micronuclei (Feng and Lin, 1987). These same investigators also reported effects on chromosomal aberrations at 400 mgikg and above. The results remain unexplained.
In Vivo Assays Armed with knowledge of how these compounds degrade, it is not surprising that mutagenicity is absent in vivo (Table 79.8). 79.3.5.3 Cytogenetic Effects The effects on chromosomes mirror the pattern of activity for mutations: clastogenic findings are seen in vitro but are generally absent in vivo.
In Vitro Assays Table 79.9 lists representative in vitro cytogenetic studies with captan or folpet. The addition of S-9
1726
CHAPTER 79
Captan and Folpet
Table 79.9 In Vitro Cytogenetic Assays Assay
Compound
Results
Reference
Chinese hamster V79
Captan
Positive for sister chromatid
Tezuka et al. (1980)
exchange and chromosomal aberrations Chinese hamster lung
Captan, Folpet
Positive in the absence of S-9
Ishidate et al. (1981)
Folpet
Positive, but higher
Loveday (1989)
fibroblasts Chinese hamster ovary
(CHO)
concentrations required with S-9
Human blood Iymphocytes Human lymphoid cell
Captan
Negative
Folpet
Negative (5
Captan
Positive in the absence of S-9
Sirianni and Huang
Captan
Negative
Sasaki et al. (1980);
Captan
Positive
PiIinskaya (1983) ~g,
Bootman et al. (1987)
2-h exposure)
line
(1978)
Human diploid fibroblast cell line
Tezuka et al. (1978)
Human embryonic lung
Legator (1969)
and rat kangaroo cell lines
The work by Chidiac (Chidiac, 1985; Chidiac and Goldberg, 1987) provides valuable information with regard to the genotoxicity of these compounds. The basis for this work drew upon evidence of mutagenicity and the tumorigenic effect captan has on the mouse duodenum. It was postulated that evidence of cytogenetic damage would be seen in the duodenum after exposure to captan. This mouse bioassay was validated with known carcinogens and noncarcinogens. Nuclear aberrations (NA) consisted of micronuclei and apoptotic bodies in the crypt cells of the duodenal epithelium. X-irradiation, 1,2-dimethylhydrazine, benzo(a)pyrene (B(a)P), and N-methyl-N-nitrosourea (MNU) induced tumors in the small intestine. Each led to a dose-related increase in the incidence ofNA 24 h after administration to mice. Benzo(e)pyrene and methylurea, which are noncarcinogenic structural analogs of B(a)P and MNU, did not induce NA. Cells of the duodenum were harvested and examined for the presence of NA
after a variety of captan dose regimens. Captan as well as THPI consistently failed to induce NA. Captan was administered to male CD-l mice using a number of regimens, including a single bolus dose of 4000 mg/kg, dietary dose levels of 4000 and 16,000 ppm, and five repetitive doses totaling 5000 mg/kg (Table 79.11). In all cases, including pretreatment with L-buthionin-S,R-sulfoximine (an inhibitor of glutathione synthesis), the investigators noted an absence of the expected signs of DNA damage. Folpet was recently tested in a study that replicated the Chidiac experimental design (Gudi and Krsmanovic, 2001). Mice were administered five consecutive daily oral doses of folpet at 2000 mg/kg per day. Nuclear aberrations in the duodenal crypt compartment were absent in folpet-treated mice, whereas mice administered a single dose (65 mg/kg) of dimethylhydrazine showed both apoptotic cells and micronuclei in the crypts.
Table 79.10 In Vivo Cytogenetic Assays Assay
Compound
Micronucleus
Captan
Chromosomal aberration
Results
Reference
Negative
Jacoby (I 985b)
Positive
Feng and Lin (1987)
Folpet
Negative
J acoby (l985a)
Captan
Negative
Tezuka et al. (1978)
Negative
Fry and Fiscor (1978)
Negative
Chidiac and Goldberg
(see Table 22)
Heritable translocation
(1987)
Positive
Feng and Lin (1987)
Folpet
Negative
Esber (1983)
Captan
Negative
Jorgenson et al. (1976)
79.3 Toxicology
1727
Table 79.11 Nuclear Aberration Study with Captana Treatment
Dose levels
Results
Single bolus dose
o and 4000 mg/kg o and 4000 mg/kg
Negative
Single bolus dose after
Negative
pretreatment with BSOb Dietary administration, 7 days
0, 8000, and 16,000 ppm
Five daily doses
Single
Negative
Cumulative
0
0
20
100
200
1000
1000
5000c
aReproduced with permission from Chidiac and Goldberg (1987). bBSO: L-buthionin-S,R-su1foximine, an inhibitor of glutathione synthesis. cDoses: Day 1,2000 mg/kg; day 2, dosing suspended due to toxicity; days
79.3.5.4 Dominant Lethal Assays
Dominant lethal assays have generally been negative (Table 79.7). However, positive findings for both compounds have been reported (Collins, 1972a, b). In spite of these positive findings, it appears that the compounds do not induce dominant lethal effects. This conclusion is based on (1) the absence of positive micronucleus assays [with the noted exception of Feng and Lin (1987)]; (2) the absence of adverse effects in two-generation rat reproductive studies; (3) the lack of negative dominant lethal effects in other studies [captan: Kennedy et al. (1975a), Rideg (1982), Shirasu et al. (1978), Tezuka et al. (1978); folpet: Bradfield (1980), Calandra (1971), Kennedy et al. (1975a), Rideg (1982)], and (4) the consistency of the findings with the rapid degradation of these compounds. A less than I-s (captan) or less than 5-s (folpet) half-life in blood argues against the possibility of parent molecules reaching the testes. Thiophosgene is considered to be more reactive than captan or folpet, but also would not reach the testes. 79.3.5.5 DNA Interaction
Captan was negative for inducing unscheduled DNA synthesis (UDS) in human diploid fibroblasts in vitro (Mitchell, 1975). It was also negative for UDS in primary liver cells (Probst et aI., 1981; Rocchi et aI., 1980). The nature of the captan and folpet molecules imparts difficulties in conducting in vivo DNA binding studies. Two approaches, using radiolabeled test material, have sought to determine if captan covalently binds with duodenal DNA of the mouse. In both cases, the trichloromethylthio side chain was labeled because it is the chemically active portion of the molecule and is expected to participate in DNA binding if it occurs. The first study used 14C-captan (Selsky and Matheson, 1981); the second study used 35S-captan (Provan et aI., 1995). When the carbon atom of the trichloromethylthio moiety is labeled, it enters the C-l carbon pool via C02 that is formed as the molecule degrades. As such, a low level of ubiquitous labeling appears throughout the mouse. When the sulfur atom
3-5,
Negative
1000 mg/kg.
of the trichloromethylthio moiety is labeled, sulfur exchange occurs, resulting in a low level of incorporation of 35S into proteins. Histones, in turn, are associated with DNA and result in a low level of associated radioactivity (Provan et aI., 1995). Investigators have concluded that covalent binding of captan to DNA has not been demonstrated (Pritchard and Lappin, 1991; Provan et aI., 1992, 1993; Selsky and Matheson, 1981). An experimental design that "proves the negative," however, has not been achieved and the EPA holds that in vivo DNA binding has not been ruled out (Hsu and McCarroll, 1998). The polyps, adenomas, and adenocarcinomas that develop in the mouse duodenum as a result of continuous oral administration of captan or folpet arise from the crypt cell compartment. Figure 79.5 depicts a much simplified anatomy of the duodenum. Should mutagenicity play a role in the development of these tumors, it must be consistent with this anatomy and the nature of the chemical reactions associated with these compounds. There are two factors that suggest mutagenicity cannot be involved in the etiology of these tumors: first, nearly all absorption takes place through the villi; second, the degradation rates of captan and folpet prevent them from reaching the crypt compartment through diffusion. Material that is absorbed through the villi enters blood or lacteal vessels and is transported away from the crypts. Crypt cells receive blood supply from arterial vessels rather than the portal system. The remaining molecules that start to diffuse down to the crypt compartment must first pass through mucus and then diffuse through approximately 16 epithelial cells before reaching the stem cells located in position T4 from the base of the crypt (Potten and Loeffier, 1990). Mutational events in cells distal to the stem cells (some of which may still be dividing) are of no import because these cells migrate up the villi and are shed within 2-4 days. The duodenal mucosal cells are rich in glutathione, having a concentration of approximately 8 mmol in CD-l mice (Chasseaud et aI., 1991); thus, the degradation of parent molecules is promoted. Whereas the half-lives of captan and folpet are very short, the exponential loss of captan virtually eliminates all molecules in short order.
1728
CHAPTER 79
Captan and Folpet
Ml,,,",. ,''''" r
> 99% Absorption through villi
Villus
Crypt
llnsnl Cell Location (T4)
+
___
'--.1..--::::.- To Portal Circulation Arterial Supply Figure 79.5
Schematic of duodenal villi and crypts.
In summary, captan and folpet are chemically active molecules that can induce mutations and cytogenetic effects if they are positioned to interact with sensitive targets. The very nature of this reactivity coupled with mammalian anatomy, however, precludes such interaction in vivo. This conclusion is supported by the weight of evidence (including chemical fate), although instances of positive results in vivo have been reported. Captan and folpet are judged not to act as mutagens or genotoxins in the intact animal.
79.3.6 CARCINOGENICITY 79.3.6.1 Overview Carcinogenicity joins mutagenicity as a controversial topic with cap tan and folpet. Both compounds induce the development of duodenal tumors in mice when fed at high doses. This observation is cited by governmental agencies and has become the focus of regulatory actions. Rodent bioassay data are robust: there is a treatment and dose relationship of tumor incidence in mice, but such a relationship is absent in rats. Rat studies, however, have shown increased incidences of some tumors, but these are judged to be not treatment related (Gordon et aI., 1994). This finding is not embraced by the EPA (Quest et aI. , 1993 ; U.S. EPA, 1999a, b). At issue is the regulatory classification of captan and folpet. The current B2 classification held by the EPA characterizes these fungicides as "probable human carcinogens" (U.S. EPA, 1984b, 1986a, 1989). The framework around which the EPA analyzes captan and folpet carcinogenicity data is the 1986 carcinogen risk assessment guidelines (U.S. EPA, 1986c). The proposed new cancer guidelines (U.S. EPA, 1996) provide an opportunity for fresh evaluation that incorporates both mechanistic and chemical property data. Studied in
this light, both captan and folpet are judged not to pose a carcinogenic risk to humans and should be classified as "not likely to cause tumors at low doses." The practical meaning of such a classification is that captan and folpet are not human carcinogens.
79.3.6.2 Mouse Bioassays An early captan study combined both gavage and dietary administration (Innes et al. , 1969). These investigators dosed neonatal F, hybrid mice by gavage with 215 mg/kg per day for 3 weeks and followed by dietary administration of 560 ppm for 18 months. This study was negative. However, the dietary concentration, in hindsight, appears to be below the threshold necessary for tumor induction. The National Cancer Institute administered captan to B6C3FI mice at 8000 and 16,000 ppm for 80 weeks (NCI, 1977). Duodenal tumors were evident at the high dose (3 in 46 males; 3 in 48 females). There was 1 in 43 males at the 8000 ppm that also had a tumor. Two other studies confirmed the treatment relationship of captan and duodenal tumors (Daly and Knezevich, 1983; Wong et al. , 1981). The tumor incidence is shown in Table 79.12. The NOEL for duodenal tumors in mice (based on proliferative changes in the duodenum) is 400 ppm. Captan has also been evaluated by intraperitoneal and dermal administration. A study that treated two different "strain N ' mice intraperitoneally with captan (along with 64 other chemicals previously tested by the National Cancer Institute) indicated a slight increase in lung tumors in males in one strain (Maronpot et al., 1986), but the significance of these data was questioned due to lack of interlaboratory consistency and lack of correlation to the standard rodent bioassays (FAOIWHO, 1990). A dermal study using the two-stage carcinogenesis model concluded that captan was neither a complete skin car-
79.3 Toxicology
1729
Table 79.12 Captan Duodenal Tumor Incidence in Micea Dose (ppm and mg/kg/day) 0
100
400
800
6000
8000
10,000
16,000
0
15
61
123
925
NC
NC
NC
Reference
Males Adenoma
2/91
3/83
0/93
lI87
4/84
Carcinoma
0/91
0/83
0/93
0/87
2/84
Adenocarcinoma
0/9
Daly and Knezevich (1983)a
5/46
3/43
Wong et al. (1981)h
Duodenal neoplasms
NCI (1977)
2/74
20/73
21172
39/75
0
100
400
800
6000
8000
10,000
16,000
0
18
70
142
1043
NC
NC
NC
Females Adenoma
3/85
lI82
lI83
7/81
3/91
Carcinoma
0/85
0/83
0/83
0/81
lI91
Adenocarcinoma
0/9
Daly and Knezevich (1983)0 0/49
3/48
Wong et al. (1981)h
Duodenal neoplasms
NCI(1977)
2/72
24/78
19/76
26/76
aNC: not calculated. bIncidence reported reflects pathology reevaluation of slides (Robinson, 1993). cThe total tumor incidence combines both benign and malignant tumors.
cinogen nor a promoter, although at high doses (450 mg/kg three times per week for 3 weeks, followed by croton oil factor A I three times per week for 51 weeks) there was some evidence it may act as a weak initiator (Antony and Mehrota, 1994). Tissue damage rather than mutagenic effect might account for this finding, however, because the control, DMSO, did not replicate the irritation effects of captan. The carcinogenic effect of folpet on the mouse duodenum is similar to that of captan (Table 79.13). The first two bioassays had doses of 1000, 5000, and 10,000 ppm (Rubin, 1985a), and 1000,5000, and 12,000 ppm (Wong et aI., 1982). In both cases there was a low incidence of tumors at 1000 ppm. This finding triggered a third study with lower doses (East, 1994). The NOEL for tumors in the third study was established at 450 ppm. Although the primary site of gastrointestinal tumors is the duodenum in mice, a low incidence of tumors is seen in the stomach with folpet. There were some tumors noted with captan, but the incidence was low, not dose related, and not obviously treatment related. The differential aqueous stability in acid conditions of the stomach between captan and folpet may account for this finding. Cap tan elicits effects in the stomach, but these are restricted to polyp formation. The blockage from the stomach to the duodenum seen in some mice that resulted from the presence of polyps and tumors located just after the pyloric sphincter was suggested as a contributing cause of stomach tumors (Nyska et aI., 1990). This blockage was thought to result in an increased concentration of folpet and folpet degradates in the stomach. Stomach tumors were evident, however, where no blockage was apparent and thus argue against this hypothesis (East, 1994).
79.3.6.3 Rat Bioassays In contrast to mice, there is no consistent tumor response across studies with rats (Table 79.14). Evaluation of captan tumor incidence data for kidney and uterine tumors using appropriate statistics and proper tumor grouping shows no treatment effect (Foster and Elliott, 2000; Gordon et aI., 1994). It is unlikely the kidney tumors are related to treatment with captan because there is no increase in malignant tumors (carcinomas), there is a small increase (in a single animal) in benign tumors (adenomas) only, there is no statistically significant increase or trend in kidney adenomas, and the finding of kidney adenomas is seen in one out of four rat bioassays with captan and one out of seven bioassays with both captan and folpet. It is unlikely the uterine sarcoma tumors are related to treatment with captan because there is no statistical significance when tumors and polyps are considered together, a consideration dictated by the etiology of uterine sarcomas (Leininger and Jokinen, 1990). The four tumors noted are comprised of three different cell types and this finding was not consistent with the other bioassay results. In evaluating this study, the JMPR found "no other effects" in addition to depression of food intake and body weight gain at 2000 ppm and a slight increase in relative liver weight in males (FAOIWHO, 1990). With folpet, the study director concluded that the incidence for mammary glands and thyroid tumors were not related to treatment (Crown et al., 1985, 1989). The EPA, however, continues to cite the rat tumor data as supportive of their B2 classification of captan (US. EPA, 1985a) and folpet (U.S. EPA, 1995b).
1730
CHAPTER 79 Captan and Folpet
Table 79.13 Folpet Duodenal Tumor (Adenoma/Carcinoma) Incidence in Mice Dose (ppm and mg/kg/day)
Males
Females
0
150
450
1000
1350
5000"
0
16
47
93 h
151
502 b
0/52
4/52
17152
1187
2/61
8/67
1282
Reference
38171
Wong et al. (1982)
Rubin (1985a)
25/52
East (1994)
0/48
0/42
0
150
450
1000
1350
5000"
0
16
51
96h
154
515 b
1151
2/52
10/52
0188
1163
7/67
0/49
12,000
0/44
0/89
0/96
10,000"
0/49
10,000"
12,000 1284 Rubin (1985a)
19/52
38/73
Wong et al. (1982) East (1994)
1150
aDose levels for the Rubin study were 5000 and 10,000 ppm for the first 21 weeks and then adjusted down to 3500 and 7000 ppm. bWong et al. (1982).
79.3.6.4 Comparison of Rat and Mouse Response to Folpet A stark difference between mice and rats exists when comparing their tumor response to captan and folpet. All strains of mice tested show a treatment and dose-related incidence of duodenal tumors. All strains of rats show neither duodenal tumors nor proliferative changes. This suggests that the physiology of the mouse and rat differ in specific toxicokinetic and/or toxicodynamic ways that account for this difference. A series of comparative studies in the CD-l mouse and Sprague-Daw ley rat were conducted with folpet at 50 and 5000 ppm in an attempt to uncover the reason or reasons for this difference (Chasseaud et aI., 1991). Areas investigated included respective milligrams per kilogram per day doses, transit times through the gut, glutathione changes with dosing, effects on enzymes, pH changes, GSH levels, and binding of 14C-folpet to intestinal components. There were a variety of quantitative differences between rats and mice; however, none appeared to account for the qualitative difference seen in tumor response. Had a "smoking gun" been elucidated, that is, had the precise basis for the rat being refractory and the mouse being susceptible been determined, then this factor in humans could be compared with these two species. 79.3.6.5 Relevance of Mutagenicity to Mouse Thmors and Human Risk Assessment The presence or absence of a mutagenic component in the etiology of mouse duodenal tumors determines the paradigm used to assess risk to humans. Mutagenic carcinogens are thought to confer some increased risk at all dose levels, whereas chemicals that induce tumors through epigenetic means have thresholds. The paradigm for the mutagenic (nonthreshold) carcinogens assumes an increased risk at any dose above zero and is reflected by such mathematical paradigms as the linear multistage model
t,
(Pi tot and Dragan, 1996) that determines the q the slope of the 95% upper bound of the curve that describes increased risk. The procedure for assessing risk for threshold carcinogens is the MOE paradigm. MOE calculates the difference between the no effect level (NOEL) in rodent bioassays and the expected exposure to humans. An MOE of 100 or greater (NOEL/exposure > 100) is generally considered to provide sufficient safety. Section 79.3.5 concluded that captan and folpet are not in vivo mutagens. This view was also reflected by the EPA in its PD4 document on captan: "The studies on the mutagenicity of captan lend significant support to its classification as an oncogen, but there is little or no risk of its producing mutagenic effects in humans" (U.S. EPA, 1989, p. 8126). Absent this mutagenic component, a plausible alternative explanation for these tumors must be advanced. This explanation must account for the development of tumors without requiring genetic damage such as alkylation of DNA or cross-linking of DNA to proteins. Mechanistic studies have provided these data. 79.3.6.6 Mode of Action Leading to Duodenal Thmors in the Mouse There is a preponderance of data that point unambiguously to a proliferation-based nongenotoxic mode of action for captan and folpet. This mode of action is consistent with the chemical and physical properties of captan and folpet, and the effect these compounds produce with high dietary exposure in the mouse. A key component of this mode of action is that it is thresholdbased; that is, at dietary doses below the threshold, tumors will not develop. The physical and chemical properties include the instability of captan and folpet in aqueous solution at physiological pH, the reaction of captan and folpet with thiols, the generation of thiophosgene from both hydrolysis and thiol interactions, the transient nature of thiophosgene due to its chemical reactivity,
79.3 Toxicology
1731
Table 79.14 Rat Bioassays with Captan and Folpet Test material
Experimental design
Findingsa
Reference
Captan
Osborne-Mendel
Negative
NCI (1977)
Negative (uterus)
Til et al. (1983)
Negative (kidney)
GoldenthaI et al. (1982)
Negative
Hazleton (1956)
Negative
Crown et al. (1989)
Negative (thyroid, testes)
Cox et al. (1985)
Negative (mammary glands,
Crown et al. (1985)
0,2525,6060 ppm (TWA)b mg/kg/day not calculated Wistar (Cpb:WU) 0, 125,500,2000 ppm 0, 6.25, 24, 98 mg/kg/day Charles River CDl 0, 500, 2000, 5000 ppm 0, 25, 100, 250 mg/kglday 0, 1000, 5000, 10,000 ppm mg/kglday not calculated Folpet
Fischer (chronic toxicity study) 0, 250, 1500, 5000 ppm 0, 12.5,75,250 mg/kglday CD 0, 200, 800, 3200 ppm 0, 10,40, 160 mg/kg/day Fischer 0, 500, 1000, 2000 ppm
thyroid, lymphoma)
0, 25, 50, 100 mg/kglday
aThe EPA notes the incidence of tumors (in tissues) "associated" with treatment in organs listed in support of B2 cancer classification for both captan and folpet (Quest et aI., 1993). Weight of evidence analysis shows captan is not a rat carcinogen (Foster and EIliott, 2000) nor is folpet (Study Director conclusions). bTWA: time weighted average.
and the comparatively low toxicity of THPI and phthalimide. The mode of action for mouse duodenal tumors must be consistent with these properties and effects. It must also be supported by generally accepted principles of carcinogenicity. Mouse duodenal tumors develop with oral administration above a threshold if maintained for at least 6 months (Pavkov and Thomasson, 1985). Histopathological analysis shows that tumors arise from the crypt compartment and show a continuum from hyperplasia to polyps, adenomas, and adenocarcinomas (Tinston, 1995, 1996). Histologic and proliferation studies have characterized the changes to the duodenum with exposure to captan and show two sequential events (AlIen, 1994; Foster, 1994). First, epithelial cells that comprise the villi are damaged by exposure to captan and sloughed off into the intestinal lumen at an increased rate. The villi height is shortened. Second, basal cells in the crypt compartment that normally divide at a rate commensurate with the normal loss of villi cells from the tips of the villi increase their rate of proliferation to a hyperphysiologic state. Crypt depth is subsequently increased and the villi-to-crypt ratio (measured by their respective sizes) decreases. A small number of transformed cells exist in the duodenum as evidenced by a low incidence of duodenal tumors in bioassay control (Tables 79.12 and 79.13) and historical control mice (Bomhard and Mohr, 1989; Chandra and Frith, 1992; Lang, 1995; Maita et aI., 1988; Ward et aI., 1979). It is postulated
that these transformed cells are subject to proliferative pressure and, as a result of this continued pressure for at least 6 months, progress to tumors. The basis for this postulation is the body of data that show abnormally high cell proliferation, which is not carcinogenic per se, but does play a role in tumor development (Butterworth et aI., 1992; Ledda-Columbano et aI., 1989; Pitot et aI., 1991). The role of proliferation in thyroid tumors is well established (Chhabra et aI., 1992) and the influence of proliferation on initiated liver cells is also known (Solt et aI., 1977). Classically, the two-stage carcinogenesis model in the skin points to the importance of sustained proliferation in the promotion of initiated cells to tumors (Berenblum and Armuth, 1977). In addition to promoting the clonal expansion of nascent tumor cells in situ, abnormally high proliferation may increase fixation and expression of premutagenic DNA lesions, increase the number of spontaneously initiated cells during replication, perturb checkpoints in the cell cycle leading to mutagenic events, and increase the number of spontaneously initiated cells by blocking cell death/elimination (Ledda-Columbano et aI., 1989). Thus, there are two avenues for duodenal tumors to develop: promotion of nascent tumors cells and initiation of normal basal cells through disruptions in normal DNA replication. The progression to tumors under this mode of action is depicted schematic ally in Fig. 79.6. A genetic component is nei-
1732
CHAPTER 79
Captan and Folpet
ormal Duodenum
Lo~s
Villi
ell
r) pt
ell Proliferation
Enlarged rypt H) pcrpla tic r) pts Dose > SO O1gfkglds)
RC010\'al of aptan/Folpct
~
Rapid Reeo\ cry
omlal Duodenum
ontinued Irritation
-------------;~~ Adenoma
Adenocarcinoma
Proliferalh e Pressure on . pontaneou I)-tran formed 'cll ;" ';111.
Figure 79.6
Mode of action for captan and folpet in the mouse duodenum.
ther required nor plausible. Thresholds have been established for the initial cellular response to cap tan or foIpet administration: villi damage and crypt cell hyperplasia. The NOELs for captan and folpet are similar: 400 ppm (60 mg/kg per day) for captan and 450 (69 mg/kg per day; males) or 150 ppm (29 mg/kg per day; females) for folpet. Administration of captan or folpet below these thresholds will not lead to tumors, because the basis for tumor progression (hyperphysiologic cell division rate) is absent. This mode of action requires that the appropriate paradigm for assessing carcinogenic risk in humans is margin of exposure not linear low-dose extrapolation (qi). 79.3.6.7 Epidemiology In a limited retrospective cohort mortality study, 138 workers in a captan manufacturing plant who were employed for a minimum of 3 months during a 23-year period beginning in 1954 were followed for 30 years (Palshaw, 1980). These workers were exposed to captan at estimated air concentrations ranging from 0.83 mg/m3 for THPI operators to 1.54 mg/m 3 for captan operators. Other workers had little or no exposure (originally ranked as 0, 1,2, or 3 for none, low, moderate, or high, respectively). These data showed there were no increased deaths that resulted from cap tan exposure. 79.3.6.8 Summary Captan and folpet at sufficiently high doses act locally on the duodenal mucosa and result in damage to villi. Epithelial cells
of the villi are lost and the homeostatic feedback mechanism increases cell proliferation in an attempt to make up this loss. Transformed cells that reside in the crypt compartment are sensitive to this proliferative pressure and are promoted to frank tumors (Fig. 79.6). This mode of action has no mutagenic component and has a clear threshold for the first event that leads to tumors: increased proliferationlhyperplasia of the duodenal crypt compartment.
79.4 COMMON MECHANISM OF TOXICITY 79.4.1 CAPTAN AND FOLPET Captan and folpet show obvious similarities in structure and effects. The Food Quality Protection Act (V.S. Congress, 1996) formally recognized the existence of such similarities and mandated that the EPA consider common mechanisms of toxicity when conducting risk assessments. The EPA issued guidance on how to determine the presence of a common mechanism for two or more pesticides (V.S . EPA, 1999c). Their criteria include structure, adverse effects, and mode of action. Captan and folpet share sufficient common characteristics to conclude that they have a common mechanism of toxicity (Bemard and Gordon, 2000). This finding is specific to the key toxicological endpoint, duodenal tumors in mice, but may apply as well to other nonspecific endpoints. The finding that a common mech-
1733
79.4 Common Mechanism of Toxicity
anism of toxicity exists for captan and folpet is supported by the following determinations:
1. Structural similarity. The active side chains, -SCCb, are identical. 2. Site of action. Toxicity is expressed at the site of contact for both chemicals (that is, they are local irritants as opposed to systemic toxicants). 3. Reactivity with thiols. Both react with thiols to produce similar degradates. Differences in rates of reaction are attributable to physical/chemical properties of the two compounds and do not serve to diminish their commonality. 4. Mechanism of pesticidal action. Toxicity to fungi is mediated through reactions with both soluble and insoluble thiols in fungal conidia. These same reactions account for expression of the common toxic endpoint in mammals. 5. Common toxic end point. Gastrointestinal tumors in mice that generally are specific to the duodenum. 6. Mode of action. Both captan and folpet express their common toxic endpoint through a nongenotoxic compensatory proliferation mechanism. 7. Specificity of action. For both materials, the majority of tumors appear in the duodenum, but with folpet, some tumors are noted in the stomach. The hydrolytic rate of folpet is approximately 8 times faster than that of captan at pH 5 and may promote the presence of active metabolites in the acid environment of the stomach. Tumors are restricted to the mouse; rats are refractory. 8. Other toxic endpoints. Both captan and folpet show a similar pattern of toxicity for mutagenicity and skin sensitization. Both compounds show nonspecific secondary endpoints such as developmental toxicity manifested as decreased fetal weights and ossification defects at maternally toxic doses. Finding a common mechanism of toxicity for captan and folpet will influence the way the EPA regulates these two fungicides under FQPA. It also will afford toxicologists an opportunity to integrate data from the individual compounds to generate a more robust data base, which is particularly valuable for evaluation of noncarcinogenicity in rats and elucidation of the mode of action of carcinogenicity in mice. 79.4.2 CAPTAFOL
Captafol (CAS 2939-80-2, Fig. 79.7) differs from captan and folpet in a number of areas. The side chain differs in structure as well as chemical activity. The two-carbon tetrachloro moiety of captafol is able to produce an episulfonium ion that can act as a systemic alkylating agent (Fig. 79.8). This ion, absent with captan and folpet, is able to enter the systemic circulation and may be carcinogenic (Williams, 1992). The spectrum oftumors in rodent bioassays is broad and affects both mice (Ito et aI., 1984) and rats (Nyska et al., 1989; Quest et al., 1993), whereas the tumor spectrum of captan and folpet is narrow, focusing on
0:;° 0
I
Cl Cl I
I
N-S-C-C-H I
I
Cl Cl
Captafo[ CH],
°
9
Cl \\ ",,0 I N -s- N -S-C-CI
CH,'
P
CH] Tolylfluanid
Dichlofluanid
Figure 79.7
CaptafoI, dichlofluanid, and tolylfluanid.
the mouse duodenum (Gordon et aI., 1994). Mutagenic results in some assays show a differing pattern of activity. For example, when tested in S. typhimurium, TA 102 and TA 104, captan was negative in strain TA 102 and positive in strain TA 104, whereas captafol was negative for TA 104 and positive for TA 102 (Barrueco and de la Pena, 1988). In S. typhimurium strains TA 100, TA 98, TA 1535, TA 1537, and TA 1538 as well as E. coli strain WP2 her, captan and folpet were positive in all systems, whereas captafol was positive only in WP2 her and was "doubtful" in TA 100 (Moriya et aI., 1983). Two results follow from the finding that captafol does not share a common mechanism of toxicity with captan and folpet. First, under FQPA, residues will not be combined for a cumulative risk assessment. Second, the "structural similarity" (Quest et aI., 1993) of captafol should not be referenced when evaluating the carcinogenicity of captan or folpet. The first point is moot, because captafol is not registered in the United States; the second point avoids confounding comparisons. 79.4.3 DICHLOFLUANID AND TOLYLFLUANID
Dichlofluanid (CAS 1085-98-9) and tolylfluanid (CAS 73127 -1) do not share a common mechanism of toxicity with captan or folpet with regard to mouse duodenal tumors, principally because they do not induce these tumors. Both compounds have a fluorine atom substituted for one of the three chlorine atoms on the trichloromethylthio moiety (Fig. 79.7). They differ from one another by the addition of a methyl group on
o
~° I
C~l
Cl Cl I
I
N-S-C-C-H I
I
Cl Cl
\
I
V
I
N-S-C- -H -
/ t
tl~1Cl
Captafol
Figure 79.8
Episulfonium ion fonnation by captafol.
\
cr+
N-S-CCl:1
/ \CHCI /
Episulfonium ion
1734
CHAPTER 79 Captan and Folpet
the benzene ring. Like captan and folpet, these compounds react with sulfuydryl groups (Schuphan et aI., 1981). The monofluorodichloromethylthio moiety conveys more chemical reactivity to the parent as measured by the reaction rate with 4-nitrothiophenol compared to the trichloromethylthio moiety. The reaction rate of dichlofluanid is over twice that of captan and folpet, but the trichloro dichlofluanid analog is less reactive than either captan or folpet. Dichlofluanid and its bis(fluorodichloromethyl) disulfide degradate were reported to be negative for mutagenicity in S. typhimurium TA 100, whereas the bis-(trichloromethyl) disulfide from captan and folpet was positive (Schuphan et aI., 1981). The presence of the fluorine atom apparently lessens the mutagenicity of these compounds. Thiophosgene and its monofluorine analog are postulated to be degradates of dichlofluanid. Either compound reacts with cysteine to form TTCA in a similar way as it is formed with captan or folpet. These compounds have been reviewed by the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and the WHO Expert Group on Pesticide Residues (FAOIWHO, 1984,1989). Dichlofluanid was negative for carcinogenicity when tested in mice at 5000 ppm. The levels that cause no toxicological effect in rats and dogs are 500 (30 mg/kg bw per day) and 1000 ppm (25 mg/kg bw per day), respectively. For tolylfluanid, the levels that cause no toxicological effect in rats and dogs are 300 ppm (15 mg/kg bw per day) and 12.5 mg/kg bw per day, respectively. The absence of duodenal tumors in mice suggests that the ability to induce these tumors is not a general property of the chloroalkylthio fungicides.
79.5 HUMAN RISK ASSESSMENT 79.5.1 CANCER In contrast to the relatively high background duodenal tumor incidence seen in mice, the incidence in humans (Parkin et aI., 1992) and rats (Goodman et aI., 1979; Maekawa et aI., 1983; Maita et aI., 1987; McMartin et aI., 1992) is low. This suggests that humans are closer to rats; that is, humans are refractory to tumors with captan or folpet because the number of transformed cells in situ is low. Nonetheless, prudence dictates that humans be considered similarly to the mouse for risk assessment purposes. The no effect levels for duodenal crypt cell proliferation, the prerequisite for tumor formation, are 400 ppm for captan and 150-450 ppm for folpet. The approximate equivalent doses are 30-60 mg/kg per day. For this assessment we used a NOEL of 50 mg/kg per day. Humans are exposed to captan and folpet predominantly by two routes: oral and dermal. Exposure via the oral route occurs through consumption of food that contains residues; exposure via the dermal route occurs through the use of products that contain these fungicides. Exposure from food is low and there are no contributions from water. Milk, which is both aqueous
based and metabolic ally produced, was shown to have no captan or degradates present. A national milk survey for captan that was conducted over the course of 1 year and analyzed 224 samples from a statistically derived paradigm across four regions of the United States (North East, North Central, West, and South) found no detectable levels (LOQ = 0.005 ppm) of captan, THPI, 3-0H-THPI, or 5-0H-THPI (Slesinski and Wilson, 1992). Exposure to oral residues only is considered relevant for human cancer risk assessment. Dermal exposure is not relevant for human cancer risk assessment, because dermal contact does not result in systemic exposure and captan has been found not to be a skin carcinogen (Antony and Mehrota, 1994). For both captan and folpet, the EPA has calculated the estimated exposure for cancer risk purposes as 0.00005 mg/kg per day (U.S. EPA, 1999a, b). The MOE for each of these fungicides, based on a NOEL for duodenal crypt cell proliferation of 50 mg/kg per day is 1,000,000. These MOEs suggest virtually no risk of cancer to persons who consume produce treated with either captan or folpet. It is unlikely that both compounds would be present on the same commodity at the same time because the uses of captan in the United States do not overlap those of folpet. Additionally, normal agronomic practice usually relies on one or the other, not both. Nonetheless, if the expected residues are combined for a cumulative risk assessment, the MOE is still satisfactory. This analysis shows that humans are not at risk for duodenal tumors from these fungicides. This level of risk would be characterized as "not likely at oral low doses" and "not likely by dermal exposure" according to the proposed cancer risk assessment guidelines (U.S. EPA, 1996). In contrast to the EPA B2 classification, the practical meaning of this assessment, based on the mode of action of captan and folpet is that they are not human carcinogens. This assessment is particularly relevant for reentry workers such as strawberry harvesters who might be exposed to captan residues. 79.5.2 NON CANCER For noncancer risks, captan and folpet present an interesting challenge for risk assessors. The transient nature of these molecules coupled with their inherent low toxicity make it difficult to assign meaningful endpoints. Noncancer endpoint risk characterization requires the selection of relevant endpoints for nondietary and dietary exposure, and that NOELs be determined for both acute and chronic exposure. Nondietary exposure, in turn, comprises dermal exposure (including eye exposure) and inhalation. Three nondietary hazards associated with captan and folpet that are relevant to human safety are skin sensitization, eye irritation, and lung irritation. Only one of these, skin sensitization, appears to effect persons who come in contact with these materials. The incidence of sensitization reactions is below 10% in trials with captan and well below this incidence in actual use
79.6 Conclusion (Krieger, 2001). Systemic toxicity from dermal exposure is not possible due to the labile nature of these molecules in the blood. Skin irritation from single instance contact with captan or folpet is not expected. Repetitive dermal exposure, however, might induce progressive skin irritation, although it is not evident with people who repeatedly use a shampoo containing 7% captan (Guo,2001). Inhalation is a potential avenue for adverse effects, although the absence of adverse reports suggest that this is not an issue. The AIHGH has assigned a threshold limit value of 5 mg/m3 for captan (ACGIH, 1998) and the same value has been suggested as appropriate for folpet (Seifried, 1996). For acute dietary risk, the EPA has used the NOEL from developmental studies for both captan (V.S. EPA, 1995a) and folpet (Levy et aI., 1997). For captan, this is 10 mg/kg per day, based on effects at 30 mg/kg per day (a maternally toxic dose) in a rabbit study. For folpet, this is 10 mg/kg per day, based on effects at 20 mg/kg per day in a rabbit study. This "default" selection is not ideal because the NOEL is based on multiple doses, it is based on effects on the fetus not the individual, and it is specific for a subgroup (women of childbearing age) that comprises only part of the general population. A meaningful acute dietary risk assessment is dependent on an appropriately designed single exposure oral toxicity study; such data are not currently at hand. Captan acute dietary exposure at the 99th percentile is estimated for the general V.S. population at 0.009512 mg/kg per day (Kidwell and Watters, 1999); the exposure for folpet is estimated at 0.00046 mg/kg per day (Guo, 2001; Petersen, 1997). The EPA estimates these acute dietary exposures at 0.036 mg/kg per day for captan at the 99.9th percentile (V.S. EPA, 1999a) and at 0.001532 mg/kg per day for folpet at the 99th percentile (V.S. EPA, 1999b). For chronic EPA estimates, the dietary exposure for the general population in the Vnited States is at 0.000664 mg/kg per day for captan and at 0.000053 mg/kg per day for folpet (V.S. EPA, 1999a, b). For chronic dietary risk assessment, the captan NOEL of 12.5 mg/kg per day and the folpet NOEL of 9 mg/kg per day are used. Margins of exposure (NOEL -7- exposure) for captan are 18,825 and for folpet are 169,811. The WHO AD! is 0.1 mg/kg per day for both captan (FAOIWHO, 1990) and folpet (FAOIWHO, 1996). This is approximately equal to the EPA's cPAD for captan, 0.13 mg/kg per day, and the EPA's PAD (without the threefold FQPA safety factor) for folpet, 0.09 mg/kg per day.
79.6 CONCLUSION Captan and folpet are structurally similar molecules that act through a common mechanism with regard to their ability to induce duodenal tumors in mice. The mode of action has been elucidated for these tumors and is dependent on irritation to and cell loss from the intestinal villi, followed by a compensatory increase in proliferation within the crypt compartment.
1735
This proliferative pressure, with time, promotes transformed cells that are normally resident in situ. The mode of action is not dependent on a mutagenic component nor are mutations within basal cells of the crypts a plausible occurrence. Captan and folpet are, however, mutagenic when tested in a variety of in vitro systems, and this observation has challenged investigators to solve the paradox that exists between in vitro and in vivo test results. The solution to this question is the finding that these compounds degrade extremely rapidly when thiols are present. In human blood, captan's t 1/ 2 is less than 1 sand folpet's is less than 5 s. Thiophosgene, the reactive degradate that is formed from the trichloromethylthio side chain, reacts not only with thiols, but with other functional groups as well and is also rapidly lost. The import of this rapid degradation is that systemic exposure to captan, folpet, or their common degradate, thiophosgene, is absent. This, along with the low estimated dermal absorption rate of 0.5% per hour, assures that adverse systemic risk in agricultural workers is absent. Local effects due to irritation, however, may occur. These include eye and skin irritation, skin sensitization, and irritation of the airways. Oral exposure at sufficient doses will irritate the mucus membranes of the gastrointestinal tract. Systemic effects noted in laboratory studies such as depressed weight gain or delayed development of fetuses and pups are secondary effects that result from the primary irritation of the gastrointestinal tract. Thus, captan and folpet, when used in the agricultural setting are characterized as follows: • They have low acute toxicity. • They are not carcinogenic, mutagenic, or teratogenic. • They are neither selective developmental toxins nor are they reproductive toxins. Relevant hazards are the following: • Irritation of mucus membranes. • Sensitization after repeated exposure. • Irritation of the skin after repeated exposures (specifically for folpet). • Irritation of the airways. These products have been in use for over 50 years and experience shows that eye irritation and sensitization reactions, particularly with reentry operations, are not problematic. In addition, a limited survey of persons using a 7% captan-based shampoo indicates that repeated use does not cause skin irritation or skin sensitization reactions. The risks of captan and folpet are low; the benefits to the agricultural community are high. Captan and folpet remain valuable fungicides.
1736
CHAPTER 79
Captan and Folpet
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U.S. EPA (1984b). "Weight of Evidence Classification of Captan as B2." Report 46, Fed. Reg. 46294, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1985a). "Captafol: Notice of Special Review." Report 49, Fed. Reg. 1l03, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1985b). "Captan Position Document 2/3." Report 50, Fed. Reg. 25885, Office of Pesticides and Toxic Substances, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1986a). "Classification of Folpet as B2." Report 51, Fed. Reg. 33992, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1986b). "Guidance for the Reregistration of Pesticide Products Containing Captan as the Active Ingredient (EPA Case Number 0120)." U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1986c). "Guidelines for Carcinogen Risk Assessment." Report 51, Fed. Reg. 33992-34003, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1987). "Guidance for the Registration of Pesticide Products Containing Folpet as the Active Ingredient (Case Number 0630)." U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1989). "Captan; Intent to Cancel Registrations; Conclusion of Special Review (PD4)." Report 54, Fed. Reg. 8116-8150, Office of Prevention, Pesticides and Toxic Substances, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1995a). "The HED Chapter of the Re-registration Eligibility Decision Document (RED) for Captan." Health Effects Division, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1995b). "The HED Chapter of the Re-registration Eligibility Decision Document (RED) for Folpet." Health Effects Division, Office of Pesticide Programs, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1996). "Proposed Guidelines for Carcinogen Risk Assessment." Report 61, Fed. Reg. 17960-18011, U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1999a). "Captan: Reregistration Eligibility Decision (RED)." Report 738-R-99-015, Prevention, Pesticides and Toxic Substances (7508C), U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (1999b). "Folpet: Reregistration Eligibility Decision (RED)." Report 738-R-99-011, Prevention, Pesticides and Toxic Substances (7508C), U.S. Environmental Protection Agency, Washington, DC. U.S. EPA (I 999c). "Guidance for Identifying Pesticide Chemicals that Have a Common Mechanism of Toxicity, for Use in Assessing the Cumulative Toxic Effects of Pesticides." Report 6055, U.S. Environmental Protection Agency, Washington, DC. Valencia, R. (1981). "Mutagenesis Screening of Pesticides Using Drosophila." Report 600/1-81-017, U.S. Environmental Protection Agency, Washington, DC. van Welie, R., T. H., van Duyn, P., Lamme, E. K., Jager, P., van Baar, B. L. M., and Vermeulen, N. P. E. (1991). Determination of tetrahydrophthalimide and 2-thiothiazolidine-4-carboxylic acid, urinary metabolites of the fungicide captan, in rats and humans. Int. Arch. Occup. Environ. Health 63(3), 181-186. Vogel, E., and Chandler, J. L. R. (1974). Mutagenicity testing of cyclamate and some pesticides in Drosophila melanogaster. Experientia 30(6), 621-623. Vondruska, J. F. (1969). "Teratologic Investigation of Captan in Macaca mulatta (Rhesus monkey) and Macac arctoides (Stumptailed macaque)." Report M5519, Industrial Bio-Test Laboratories, Inc. (MRID 00043398). Vos, J. G., and Krajnc, E. I. (1983). Immunotoxicity of pesticides. Dev. Sci. Practice Toxicol. 11, 229-240. Waner, T. (1988). "Folpan: Chronic Oral Study in Beagle Dogs for 52 Weeks." Report MAK.062IFOL, Life Science Research Israel, Ltd., Ness Ziona, Israel. Ward, J. M., Goodman, D. G., Squire, R., Chu, A., and Linhart, M. S. (1979). Neoplastic and nonneoplastic lesions in aging (C57BLl6N x C3H1HeN)FI (B6C3F 1) mice. J. Natl. Cancer Inst. 63, 849-854.
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CHAPTER 79
Captan and Folpet
Waterson, L. (1995). "Folpet: Investigation of the Effect on the Duodenum of Male Mice after Dietary Administration for 28 Days with Recovery." Report MBS 45/943003, Huntingdon Research Centre Ltd. (MRID 44286303). Williams, G. M. (1992). DNA reactive and epigenetic carcinogens. Exp. Toxico!. Pathol. 44, 457-464. Wilson, A., and Wright, A. (1990). "A Study of Dermal Penetration of Carbon 14-Folpet in the Rat." Report MAG/IIPH, Toxicol Laboratories, Inc. (MRID 42122018).
Wong, Z. A., Bradfield, L. G., and Akins, B. J. (1981). "Lifetime Oncogenic Feeding Study of Captan Technical (SX-944) in CD-l Mice (ICR Derived)." Report SOCAL 1150, Chevron Environmental Health Center, Richmond, CA (MRID 00068076). Wong, Z. A., Eisenlord, G. H., and MacGregor, J. (1982). "Lifetime Oncogenic Feeding Study ofPhaltan Technical (SX-946) in CD-l (ICR Derived) Mice." Report SOCAL 1331, Chevron Environmental Health Center, Richmond, CA (MRlD 125718).
CHAPTER
80 Mammalian Toxicokinetics and Toxicity of Chlorothalonil P. P. Parsons Syngenta
80.1 IDENTITY AND USES OF CHLOROTHALONIL Chlorothalonil is a halogenated benzonitrile fungicide with broad spectrum activity against vegetable, ornamental, orchard, and turf diseases. It was first registered for use as an agrochemical in the United States in 1966. Chlorothalonil is available in a wide variety of formulations including suspension concentrates, wettable powders, and water dispersible granules. Its mode of fungicidal action is to bind to sulfhydryl groups of amino acids, proteins, and peptides and, in doing so, it ties up free glutathione in fungal cells, thereby blocking glycolytic and respiratory enzyme pathways. This action prevents the ability of fungal cells to infect plants and results in death of the fungus. Chlorothalonil's multisite mode of action has meant that no significant problem with fungal resistance has been encountered. In addition to its use as an agricultural fungicide, chlorothalonil also has wider biocidal applications, for example in paints and lubricating fluids.
Chemical class Empirical formula Synonyms and trade names
halogenated benzonitrile CCl4N2 Bravo®, Daconil®, Tuffcide®, Acticide®
80.1.1.3 Physical Properties Physical appearance Solubility (at 25°C)
Vapor pressure Molecular weight Melting point 10gPo/w
grey/white crystalline solid, odorless in pure form practically insoluble in water (0.6-0.8 mgll) xylene-80 gll acetone-20 g/l cyclohexane-30 g/l 7.62 x 10- 5 Pa at 25°C 265.9 250-251°C 2.94 at 25°C
80.2 MAMMALIAN TOXICOKINETICS 80.1.1 PHYSICAL AND CHEMICAL PROPERTIES
An overview of the available metabolism and pharmacokinetic data for chlorothalonil has been published by Wilkinson and Killeen (1996).
80.1.1.1 Structure CN
80.2.1 ORAL ADMINISTRATION
C I * " " " Cl
~
Cl
CN
Cl
80.1.1.2 Chemical Identity Common name CAS No. EINECS Chemical name Handbook of Pesticide Toxicology Volume 2. Agents
chlorothalonil 1897-45-6 217-588-1 2,4,5,6-tetrachloroisophthalonitrile
In rats, given a single, oral low dose (1.5 mg/kg) of chlorothalonil, around 20-22% of the absorbed dose is excreted in bile and around 10% in urine (Marciniszyn et al., 1985a, b, 1986a). At higher doses (200 mg/kg) a considerably lower proportion (8%) of the absorbed dose is excreted in bile, indicating that this is a saturable process. These data indicate that overall absorption from the G.!. tract is in the order of 30-32% of the administered dose. The majority of radiolabel is excreted in feces with at least 80% of administered dose excreted by this route within 96 h. Approximately 90% of the administered
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CHAPTER 80
Chlorothalonil
dose was excreted within 24-48 h although excretion was less rapid at doses of 50 mg/kg and above. Highest tissue concentrations were observed in the kidney, approximately 0.1 % of the dose. A similar metabolic profile was seen on repeated dosing and there was no evidence for bioaccumulation (Savides et aI., 1986a, b). Thiol-derived metabolites were identified in urine. Following administration of similar doses of chlorothalonil to germ-free rats, only 3% of the dose appeared in urine with lower proportions excreted as thiol-derived metabolites, indicating that gut microflora may play a role in the disposition and metabolism of chlorothalonil in the rat. Bile cannulation studies have confirmed that chlorothalonil undergoes enterohepatic circulation in the rat (Marciniszyn et aI., 1986b). In dogs, approximately 6% of an oral dose of 50 mg/kg was excreted within 48 h (1 % in urine and 5% in bile). As with the rat, absorption and subsequent excretion were rapid with around 89% of an administered dose recovered within 48 h. The extent of urinary thiol-derived metabolite excretion in dogs was lower than that seen in rats. At necropsy at 48 h approximately 0.1 % of the administered dose was present in the liver and kidneys with <0.01 % in other tissues (Savides et aI., 1995). Limited data for the monkey show that, following a single oral dose of 50 mg/kg, 1.8-4.1 % of the dose appeared in urine with very low levels of thiol-derived metabolites appearing in urine. Fecal excretion predominated with around 92% of the dose eliminated via this route over 96 h. Absorption and excretion were rapid and there was no evidence of bioaccumulation (Savides et aI., 1990). There are limited data concerning the disposition and excretion of chlorothalonil in the mouse with no metabolism data in this species. Low levels of radioactivity were found in the tis-
sues and urinary excretion indicated that at least 10% of the dose was absorbed with the majority (70-80%) of the dose excreted in faeces (Ribovich et aI., 1982). Comparison of the differences in urinary metabolite excretion profile between species suggests that the ability to excrete thiol-derived metabolites may be correlated with the observed species differences in susceptibility to renal toxicity. Mechanistic studies have been conducted to determine if a relationship exists between the ability to excrete urinary thiolderived metabolites of chlorothalonil and the potential to induce renal toxicity. Inhibition of y-glutamyltranspeptidase using Acivicin (Savides et aI., 1985) and renal organic anion transport using probenecid (Marciniszyn et aI., 1986b) decreased urinary thiol-derived metabolite excretion in rats. Administration of the mono glutathione conjugate to rats was shown to produce a qualitatively similar pattern of metabolite excretion to that seen following administration of chlorothalonil itself (Mead et aI., 1987a, b). These studies indicate that excretion of thiol-derived metabolites of chlorothalonil requires glutathione conjugation and then subsequent enzymatic processing of glutathione-derived conjugates that are selectively accumulated within the kidney. By analogy with other chemicals that undergo extensive glutathione conjugation, it is reasonable to presume that metabolism proceeds via cysteine conjugates and N -acetyl cysteine conjugates ("mercapturates") as outlined in Fig. 80.1. In conclusion, data from a variety of species demonstrate that, following oral administration, chlorothalonil is rapidly absorbed with fecal excretion predominating. The toxicokinetic profile is similar on repeated dosing with no evidence for bioaccumulation. Glutathione conjugation plays a central role
Thil-derived metabolites in plasma
1
Di- and tri-glutathione
""ju;"" ¥
Chlorothalonil & metabolites excreted in feces
i,
Enterohepatic re-circulation
b,l,l~///"
Thiol-derived metabolites in kidney
1
Thiol-derived metabolites excreted in urine
Figure 80.1 Diagram illustrating proposed metabolism of chlorothalonil following oral administration to rats. Broken lines indicate multi stage events involving several enzymatic steps and transport processes.
80.3 Acute Toxicity
in the metabolism of chlorothalonil and subsequent complex metabolic processing of these conjugates results in selective renal uptake and urinary excretion of thiol-derived metabolites. Knowledge from the metabolism of other chemicals that undergo extensive glutathione conjugation implicates a role for mercapturic acid-mediated metabolism for chlorothalonil.
80.2.2 DERMAL ADMINISTRATION Studies have been conducted to determine the nature of metabolites appearing after dermal administration of chlorothalonil to the rat and monkey. Separate in vitro and in vivo studies have been conducted to determine the extent to which chlorothalonil undergoes percutaneous absorption.
80.2.2.1 Urinary Metabolite Profile Following Dermal Exposure Limited studies have been conducted in the rat and the monkey to investigate the profile of urinary metabolite excretion after dermal administration of chlorothalonil. In rats, a maximum of 3.1 % of the applied dose was excreted in urine of which around 0.1-1 % constituted thiol-derived metabolites (Savides et aI., 1989). Fecal excretion of radiolabel was similarly low. A high proportion (20--40%) of administered dose was retained in skin at the application site. The low level of thiol-derived metabolites excreted in urine following dermal administration of chlorothalonil may explain the absence of renal toxicity seen in the subchronic toxicity studies using this route. In monkeys, only 1.2% of the dose was excreted in urine and similar amounts in feces. Around 2-4% of the dose was retained in skin at the application site after a 48 h exposure and no thiol metabolites could be detected in urine (Magee et aI., 1990).
80.2.2.2 Percutaneous Absorption A number of percutaneous absorption studies have been conducted with chlorothalonil using both in vitro and in vivo approaches. These studies used either acetone or formulation blank as the vehicle. In Vitro Studies Two in vitro studies have been conducted using human abdominal epidermis. In one study, the chlorothalonil was used either neat (no vehicle) or as a solution in acetone (Ward, 1989a). The mean absorption rate observed in this study over a 48 to 55 h exposure interval was 0.034 ± 0.020 J.lg/cm 21h, equating to 0.085% of the applied dose. No radiolabel was detected in the receptor chamber fluid until 48 h after application using either neat or diluted material. In a separate study (Ward, 1989b), chlorothalonil was applied as a suspension in a commercial formulation base either as a concentrate or as a spray-strength dilution under both occluded and nonoccluded conditions. The mean absorption rate per hour was higher with the occluded application than with the nonoccluded application (0.18 vs 0.005 J.lg/cm 21h). Approximately 0.094% of the applied dose
1745
was absorbed over a 10 h period with the non occluded application. These data indicate that the absorption of chlorothalonil through the human epidermis is considerably lower than 1% of the applied dose. In Vivo Studies A study was conducted in the rat which investigated the percutaneous absorption of chlorothalonil using acetone and a blank commercial formulation as dosing vehicles (Andre et aI., 1991a). Absorption values ranging from 6 to 26% were obtained taking into account the radiolabelled material bound to skin at the application site. A more realistic indication of systemic absorption is obtained when skin bound material is discounted. Using this approach, percutaneous absorption at 10 h postapplication in formulation blank was 2.2, 3.6, and 0.4% of the applied using applications of 0.1, 0.5, and 5.0 mg/kg, respectively. These values are consistent with the absorption profile seen in vitro. Thus, the available data indicate that chlorothalonil is a poor skin penetrant. This view is supported by the estimation of percutaneous absorption by comparison of the ability of chlorothalonil to induce toxicity within the key target organ (kidney) following dermal and oral administration to rats. Such a comparison is based on the findings of the 21 day dermal toxicity study (see Section 80.5.2) with the interim findings after a similar dosing period in a 90 day subchronic oral toxicity study (see Section 80.5.1). In these respective studies, the NOAELs for the induction of renal tubular hyperplasia were 600 mg/kg/d (the highest dose) and 3 mg/kg/d. Toxicokinetic studies in rats indicate that approximately 32% of an oral dose of chlorothalonil is absorbed from the G.!. tract. Therefore, the NOAEL for renal hyperplasia after 6 weeks in the 90 day oral rat study of 3 mg/kg/d equates to a systemic dose of 0.96 mg/kg/d (i.e., 0.32 x 3). Since the application of 600 mg/kg/d of chlorothalonil to rat skin did not result in any kidney toxicity, it can be deduced that that approximately 0.16% of the applied dermal dose was absorbed systemically (i.e., 0.96/600 x 100%).
80.3 ACUTE TOXICITY 80.3.1 ORAL Chlorothalonil is not acutely toxic by the oral route having a maximum lethal dose (MLD) > 5000 mg/kg in the rat with no deaths at 5000 mg/kg in carbosa methyl cellulose (Moore, 2000). The only clinical signs of toxicity were congenital staining, soft feces, and/or occurred on the day of dosing and the day following dosing.
80.3.2 DERMAL The dermal MLD was> 10,000 mg/kg in rabbits with no deaths observed at this dose. In this study, chlorothalonil was applied
1746
CHAPTER 80 Chlorothalonil
to abraded skin for 24 h (Shults et aI., 1981b). Slight edema and yellow discoloration were observed at the application site and eye irritation was also seen. At necropsy, pale areas were observed in the liver. Chlorothalonil is not systemically toxic by the dermal route. 80.3.3 INTRAPERITONEAL The intraperitoneal MLD has been estimated to be 3.2 mg/kg in the rat. Clinical signs of toxicity were not reported. Pathological findings were generally consistent with injection of an irritant substance into the peritoneal cavity with chronic fibrinous peritonitis, enlarged and congested mesenteric lymph nodes, intestinal edema, and red foci on the kidneys and lungs (Wazeter and Lucas, 1971). 80.3.4 INHALATION Acute inhalation exposure (whole body) of rats to an atmosphere containing chlorothalonil dust resulted in a 4 h MLC of 0.1 mg/l (Shults et aI., 1993). Mortality was dose-related with deaths at 0.08, 0.14, and 0.21 mg/l (4, 6, and 9/10 animals, respectively) occurring from 10 minutes to 2 days postexposure. Clinical signs of toxicity included gasping, eye closure, and exaggerated breathing observed during exposure. Recovery was evident from day 4 onward and the majority of animals were normal by day 9. Congestion of the lungs and white frothy fluid in the trachea were observed in those animals that died with no abnormal pathology in surviving animals. The MMAD was 2.5-3.6!l-m with 25% of particles <2 !l-m in diameter. In a separate study (Shults et aI., 1981c), rats were exposed (whole body) to a dust atmosphere containing chlorothalonil at concentrations of 0.07-0.22 mg/l for 4 h. The MMAD was 1.35-5.5 !l-m with 90% of particles of < 10 !l-m in diameter. The MLC was estimated to be 0.09 mg/l with a dose-related incidence of mortality (1120 at 0.7 mg/l vs 19120 at 0.22 mg/l). Clinical signs of toxicity included rales and a bloody nasal discharge. Pulmonary congestion was observed in animals that died during the study. Hepatic necrosis and deposition of eosinophillic material were observed at the top dose. A high incidence of respiratory mycoplasmosis was seen in all animals in this study. Although the study is compromised by the presence of concomitant infection, the 4 h MLC was in agreement with that observed in the study above. It is concluded that chlorothalonil is toxic by inhalation causing death by asphyxiation secondary to pulmonary edema. In standard regulatory studies, the clinical signs of toxicity and pathological findings are consistent with exposure to a substance that causes pulmonary irritation to the lungs and respiratory tract. 80.3.5 SKIN IRRITATION Technical chlorothalonil is not a skin irritant in rabbits (Shults et aI., 1981d). In this study, a prolonged exposure period of 24 h
was used and effects were studied using both abraded and intact skin. Although the study conditions were designed to maximize the potential to induce skin irritation, the only effects observed were isolated signs of mild irritation. In contrast, prolonged and/or repeated dermal exposure to chlorothalonil has been shown to produce significant signs of skin irritation in acute and subchronic dermal toxicity studies in the rat and rabbit (see Section 80.5.2). It is concluded that, while chlorothalonil is not a skin irritant in standard studies for assessment of this endpoint, it does display potential to cause skin irritation in other dermal toxicity studies involving prolonged or repeated dermal application. Chlorothalonil has been shown to cause dermal reactions in humans who have been occupationally exposed to chlorothalonil, although it has not always been apparent if this constitutes and irritation or sensitization response (see Section 80.11.1).
80.3.6 EYE IRRITATION In a standard rabbit eye irritation study, chlorothalonil caused irreversible ocular lesions in rabbits with corneal opacity persisting for up to 14 days postinstillation. Effects also persist in the iris and conjunctiva (Wilson, 1977a). A further four studies have been conducted with technical chlorothalonil and each of these studies demonstrated irreversible corneal opacity (Francis et aI., 1973; O'Meara and Laveglia, 1995; Wilson, 1977b, c). The severity and persistence of these eye lesions indicates that chlorothalonil has potential to cause serious damage to eyes. The effect of washing the eyes postcompound instillation has not been investigated using the technical material itself, although studies with high strength formulations have shown that postinstillation washing ameliorates these effects. This observation is relevant to the recommended treatment of individuals following accidental ocular exposure. Experience from accidental human exposure indicates that chlorothalonil causes ocular pain and is also irritating to the human eye (see Section 80.11.2). However, the severe and irreversible eye lesions seen in the rabbit have not been documented in humans. 80.3.7 SUMMARY OF ACUTE TOXICITY Chlorothalonil has very low acute toxicity by the oral and dermal routes although it is very toxic by inhalation. Many of the effects seen following acute exposure are consistent with irritation at the initial site of contact. However, chlorothalonil was not irritating to skin when tested in a standard skin irritation study although dermal irritation has been observed in acute and subchronic toxicity studies in the rat and rabbit, indicating the potential for chlorothalonil to cause skin irritation following repeated or prolonged exposure. Chlorothalonil causes irreversible and severe ocular lesions in rabbits. The acute toxicity of chlorothalonil is summarized in Table 80.1.
80.5 Subchronic Toxicity Table 80.1 Acute Toxicity of Chlorothalonil Species
Result
Oral toxicity
Rat
MLD> 10,000 mg/kg
Dermal toxicity
Rabbit
MLD> 10,000 mg/kg
Inhalation toxicity
Rat
4 h MLC 0.1 mg/l
Rat
4 h MLC 0.09 mg/l
Study
Intraperitoneal toxicity
Rat
1 h MLC 0.52 mg/l
Rat
MLD 3.2 mg/kg
Rat
MLDSmg/kg
Mouse
MLD 12mg/kg
Skin irritation
Rabbit
aNot a skin irritant
Eye irritation
Rabbit
Irreversible corneal opacity
Rabbit
Irreversible corneal opacity
Rabbit
Irreversible corneal opacity
Rabbit
Irreversible corneal opacity
Rabbit
Irreversible corneal opacity
a Potential
skin irritant with prolonged or repeated dermal exposure.
80.4 SENSITIZATION 80.4.1 SKIN SENSITIZATION The potential for chlorothalonil to induce skin sensitization has been investigated in guinea pigs using a number of different study designs (see Table 80.2). Although these studies provide an inconsistent profile with regard to skin sensitization potential, the data are supportive of the view that chlorothalonil is a skin sensitizer of relatively low potency. It appears that a concentration of at least 40% w/v technical chlorothalonil is required for the induction of sensitization in guinea pigs. It is concluded that chlorothalonil is a weak skin sensitizer in guinea pigs. In addition to these data in animals, information is available concerning the potential of chlorothalonil to cause skin sensitization in humans following dermal exposure in the occupational setting (see Section 80.lLl).
80.4.2 RESPIRATORY SENSITIZATION There are no animal data concerning the potential for chlorothalonil to induce respiratory sensitization.
80.5 SUB CHRONIC TOXICITY 80.5.1 ORAL In rats, dietary administration of chlorothalonil for 28 days caused clinical signs of toxicity, decreased body weight, and decreased hematological parameters at doses of ?:,375 mg/kg/d. One death occurred at the top dose (1500 mg/kg/d). Absolute kidney weight was increased at ?:,175 mg/kg/d. No effects were observed at 80 mg/kg/d but this cannot reliably be considered
1747
as a NOEL as histopathological examination was not conducted in this study and longer term studies suggest that hyperplasia of the forestomach and proximal tubular epithelium can occur at this dose level (Wilson et aI., 1982b). The principal lesions observed following dietary administration of chlorothalonil to rats and mice for up to 90 days were hyperplasia and hyperkeratosis of the forestomach and hyperplasia of the proximal tubular epithelium of the kidney (Shults et aI., 1983, 1985; Wilson et aI., 1983a, b, 1984, 1985a, b). No treatment-related mortality was observed in these studies at doses up to 1500 mg/kg/d and there were only limited clinical signs of toxicity. Renal hyperplasia only occurred at a low incidence at the top dose tested in male mice with no effect in females. Other effects included decreased plasma ALT activity and increased kidney weight. The overall NOELs in these studies were 1.5 mg/kg/d for rats and 2.8 mg/kg/d for mice with respective LOELs of 3.0 and 9.2 mg/kg/d. The NOEL for hyperplasia of the forestomach was 3 mg/kg/d in both species with LOELs of 10 mg/kg/d in the rat and 9 mg/kg/d in the mouse. The forestomach lesions were fully reversible after a 13 week recovery period in rats but were not investigated in mice. The NOAEL for proximal tubular hyperplasia after 6 weeks dosing was 3 mg/kg/d in rats with a clear increase in the incidence of hyperplasia at ?:, 40 mg/kg/d in animals necropsied at 13 weeks. Proximal tubular hyperplasia was evident in 2/5 animals at 6 weeks at 10 mg/kg/d. In male mice, the NOEL for renal hyperplasia was 48 mg/kg/d with a LOEL of 130 mg/kg/d. Further investigative studies have been conducted in the rat. Vacuolar degeneration has been shown to be present in the proximal tubular epithelium after only two daily doses of chlorothalonil at 175 mg/kg/d (Ford et aI., 1988). In 28 day (Hironaka et aI., 1996) and 90 day (Mizens et aI., 1996) studies, chlorothalonil has also been shown to increase cell proliferation in the forestomach (BrdU) and proximal tubule (PCNA) of rats. Significant increases in labelling indices were evident from day 7 of dosing through to days 28 or 90 respectively at doses of ?:,15 mg/kg/d. The NOEL for increased labelling indices in both tissues was 1.5 mg/kg/d. Similar morphological lesions were seen in the renal proximal tubular epithelium following gavage administration of equivalent doses of the monoglutathione conjugate of chlorothalonil (150 mg/kg/d) or parent compound (75 mg/kg/d), although in these animals no hyperplastic changes were seen in the forestomach (Mead, 1987b). This finding implicates glutathione conjugation in the metabolism of chlorothalonilinduced renal toxicity and demonstrates that it is the parent compound that causes toxicity in the rodent forestomach. In dogs, administration of chlorothalonil for 90 days caused a decrease in body weight gain at 150 and 500 mg/kg/d with one death seen at the top dose (Fillmore et aI., 1993). There was some indication of a decrease in mean body weight throughout the study in both sexes at 15 mg/kg/d, but this was not considered to be toxicologically significant. Some changes in clinical chemistry were seen at these dose levels with plasma ALT decreased at all dose levels. The NOAEL was 15 mg/kg/d
1748
CHAPTER 80 Chlorothalonil Table 80.2 Summary of Skin Sensitization Studies Conducted in Guinea Pigs Induction
Challenge
Test method (date)
concentration
concentration
Conclusion
Reference
9 induction Buehler
10% in water
10% in water
Not a sensitizer
CTLlC/3873
100%
10% in saline
aSensitizer
Shults and Wilson,
100%
100%
aNot a sensitizer
CTLlC/3573
i.d.---O.5% in acetone
0.001% in acetone
aNot a sensitizer
Tucker, 1986
0.0125% in
Sensitizer
Tucker, 1986
Sensitizer
Tucker, 1986
(1974) 9 induction Buehler (1985) 9 induction Buehler
1985
(1985) Maximization (1986)
topical-1 % in acetone
Maximization (1986)
i.d.-5% in propylene glycol
acetone
topical-l % in acetone 20 induction open
0.01,0.1, or 1%
cutaneous (1986) lO-induction
0.0025, 0.0075 or 0.025%
100%
100%
Not a sensitizer
Wilson et aI., 1982a
Acetone vehicle
Acetone vehicle
Equivocal
Wilson et aI., 1988a
Buehler (1982) Maximization (1988) Buehler
Acetone vehicle
Acetone vehicle
Sensitizer
Wilson et aI., 1988b
Maximization
Acetone vehicle
Acetone vehicle
Sensitizer
Wilson et aI., 1988c
a Result
confirmed by rechallenge.
as effects on ALT were not considered adverse. No lesions were observed in the stomach or the kidney, indicating that the dog has a different toxicity profile to rodents.
at all dose levels, indicating the potential for chlorothalonil to cause dermal irritation with repeated exposure. 80.5.3 INHALATION
80.5.2 DERMAL
Subchronic (21 day) dermal toxicity studies have been conducted in rats and rabbits. In rabbits, no systemic effects were observed that were associated with chlorothalonil administration at doses up to 50 mg/kg/d (Shults et aI., 1986). However, histopathological examination revealed evidence of parasitic infection in all animals which compromises the value of this study. The NOEL for local effects in the skin was 0.1 mg/kg/d based on the observation of erythema and skin thickening at ::::2.5 mg/kg/d. In rats, decreases in body weight gain were observed over the 21 day study period along with simultaneous decreases in food consumption (Mizens et aI., 1986a). The decrease in weight gain was more prominent early in the study with increasing weight gain observed toward the end of the study. Plasma ALT activity was decreased at all dose levels. No histopathological effects were observed in the kidney, indicating that the NOEL for renal effects following dermal application was > 600 mg/kg/d. Erythema, hyperkeratosis, and squamous epithelial cell hyperplasia where observed at the site of application
No repeat-dose inhalation toxicity studies have been conducted with technical chlorothalonil. However, the symptoms and pathological findings seen after acute inhalation exposure are confined to the respiratory tract with no evidence of systemic toxicity. It is therefore anticipated that toxicity in a repeat-dose inhalation study would be expected to manifest as local irritation with the NOAEL being driven by site-of-contact toxicity within the lungs and respiratory tract as opposed to systemic toxicity. The subchronic toxicity of chlorothalonil is summarized in Table 80.3.
80.6 CHRONIC TOXICITY The chronic oral toxicity of chlorothalonil has been determined in the rat and mouse with two long-term studies available in each species. The initial studies on chlorothalonil (Wilson et aI., 1983c, 1985c) were conducted at relatively high dose levels and failed to demonstrate NOELs, so the studies were repeated (Wilson et aI., 1987, 1989) using lower doses. The
80.7 Genotoxicity
1749
Table 80.3 Summary of Subchronic Toxicity Data Duration
NOEL
Species
Route
(days)
(mglkgld)
LOEL (mglkg/d)
Rat
Oral diet
90
None
40 (hyperplasia of forestomach
Rat
Oral met
90
1.5
3
3
10 (hyperplasia of forestomach)
and kidney)
10 Mouse
Oral met
90
(t kidney weight,
,} ALT)
40 (hyperplasia of kidney)
(t kidney weight,
3
9
3
9 (hyperplasia of forestomach)
48 15
,} ALT)
130 (hyperplasia of kidney)
Dog
Oral--capsule
90
Rabbit
Dermal
21
0.1
Rat
Dermal
21
None
60 local effects
600
No renal histopathology
50
150 (,} weight gain) 2.5 local effects 2.6 no systemic effects
observed at top dose-600
findings in the chronic rat and mouse studies were consistent with those seen in the subchronic studies with hyperplasia of the forestomach and proximal tubular epithelium being the most prominent effects. The hyperplastic changes in the proximal tubular epithelium were associated with an increase in absolute kidney weight. The NOELs for hyperplasia and hyperkeratosis of the forestomach were 1.8 and 1.6 mg/kg/d for rats and mice respectively and the NOELs for hyperplastic changes in the proximal tubular epithelium were 1.8 mg/kg/d for rats and 5.4 mg/kg/d for mice. In addition to the hyperplastic lesions in the kidney and forestomach, some minor changes were seen in clinical chemistry, hematology, and urinalysis. There was no evidence of systemic toxicity in organs other than the kidney and forestomach. Although changes occurred in some organ weight: body weight ratios, these were attributable to decreases in body weight. In beagle dogs, I year administration of chlorothalonil caused a significant decrease in body weight gain at 500 mg/kg/d with increases in absolute liver and kidney weights at :::150 mg/kg/d (Mizens and Laveglia, 1994). No histopathological findings were seen in association with the organ weight changes. Changes in clinical chemistry were seen but only at one time point (27 weeks) and with no relationship to dose. Increased pigmentation of the kidney was observed at 150 and 500 mg/kg/d. The severity of this effect was similar at both dose levels and minimal pigmentation was seen in control and low dose animals. Due to its presence in control animals and lack of relationship to dose, the renal pigmentation was not considered to be of toxicological significance. Moreover, this observation was clearly distinct from the hyperplastic lesion seen in the renal proximal tubular epithelium of rodents. This indicates that there are species-specific differences in susceptibility to the renal toxicity of chlorothalonil.
Although dogs do not possess an anatomical equivalent of the rodent forestomach, the stomachs from this study were examined to determine if there was any evidence of cell proliferation using PCNA labelling. There was no increase in labelling index in any treated group when compared to concurrent controls, providing reassurance that chlorothalonil is not toxic to the gastric mucosa of dogs. The overall NOAEL for this study was 150 mg/kg/d. The chronic toxicity findings are summarized in Table 80.4.
80.7 GENOTOXICITY 80.7.1 IN VITRO GENOTOXICITY STUDIES
An extensive range of tests have been conducted to assess the genotoxic potential of technical chlorothalonil. In vitro studies were mostly negative including Ames tests using renal and hepatic metabolic activation systems. Up to 17 metabolites of chlorothalonil have also been tested and shown to give gave negative results using rat kidney S9. Two nonstandard bacterial DNA repair assays have been performed, one of which was positive and the other negative. A significant increase in the incidence of chromosomal aberrations has been observed in Chinese hamster ovary cells although this was only evident in the absence of auxiliary metabolic activation (Mizens et al., 1986b). In the absence of S9, increases in structural aberrations over control values were observed only at the top two concentrations which approached cytotoxic levels. These data indicate that chlorothalonil has clastogenic activity in vitro in the absence of S9. The effect was not observed in the presence of S9, even at dose levels some 20 times higher. However, given the lack of genotoxicity observed in other in vitro test systems (e.g., the Ames test) and the known
1750
CHAPTER 80 Chlorothalonil Table 80.4 Summary of Chronic Toxicity Data
NOAEL
LOEL
Duration of study
Species
(mg/kg bwtlday)
(mg/kg bwtlday)
2 years
Rat
None
40
2 years
Rat
3.8
15 (kidney tumors)
1.8
3.8 (kidney hyperplasia)
1.8
3.8 (stomach mucosal tumors)
1.8
3.8 (stomach squamous cell tumors)
2 years
Mouse
None
2 years
Mouse
5.4 23 1.6
1 year oral
Dog
150
reactivity of chlorothalonil, a rationale for such a profile is possible. Chlorothalonil can be viewed as a reactive molecule insofar as it is known to be reactive towards thiol (-SH) groups. It can be considered as a soft electrophile with a preference for sulphur nucleophiles rather then nitrogen/oxygen nucleophiles. Such chemicals tend to show reactivity toward protein (contains critical S electrophiles) rather than toward DNA (contains critical 0 and N nucleophiles). The profile of activity in genotoxicity assays is that such material displays negative findings in the Ames test but apparent positive findings in the in vitro cytogenetics assay, usually in the absence of exogenous metabolic activation. The activity of chlorothalonil in the IVC assay is likely to be through reactivity with protein (not DNA), and with the protein dependency of the chromosomal structure allowing visualization as a structural aberration. Such an activity would not be expected to produce genotoxicity in vivo, as reaction with inter- and intracellular thiols would dissipate the activity. This is supported by the observation that the in vitro activity of chlorothalonil in the IVC assay is removed by the addition of S9, and also by the fact that in vivo metabolism/distribution studies have confirmed that chlorothalonil reacts very rapidly with thiols. The clastogenic response observed in vitro in the IVC assay is therefore considered to be of no real significance regarding possible genotoxicity in vivo especially when considered in light of the findings of the in vivo cytogenetic studies with chlorothalonil. 80.7.2 IN VIVO GENOTOXICITY STUDIES
Several in vivo bone marrow cytogenetic studies have been conducted with chlorothalonil using single or repeated dosing schedules in three different species (rat, mouse, and Chinese hamster). Most of these studies used high doses of chlorothalonil (up to 5000 mg/kg) which resulted in some mortality. All of these in vivo cytogenetic studies were negative indicating that the clastogenicity seen with chlorothalonil in vitro
125 23 (kidncy hyperplasia) lOO (stomach mucosal tumors) 5.4 (stomach squamous cell tumors) 500 (decreased body weight gain)
is not manifest in vivo. Further reassurance for lack of genotoxic activity in vivo comes from a study which demonstrated that intraperitoneal administration of radiolabelled chlorothalonil to the rat did not result in any labelled material binding covalently to rat kidney DNA. It is therefore concluded that chlorothalonil is not genotoxic in vivo. The results of the genotoxicity studies conducted with chlorothalonil are summarized in Table 80.5.
80.8 CARCINOGENICITY Treatment-related increases in the incidence of renal tubular adenoma and carcinoma were observed in rats and male mice (Wilson et aI., 1983c, 1985c, 1987, 1989). Squamous cell adenomas and carcinomas were also observed in the forestomach of both species. In dogs, there was no evidence of neoplastic development nor was there any evidence for the occurrence of preneoplastic lesions in the kidney or stomach after administration of chlorothalonil for up to 1 year. The NOEL for mucosal cell tumors of the forestomach was 1.8 mg/kg/d in rats and 23 mg/kg/d in mice. In rats, the NOEL for tumors of the proximal tubular epithelium was 3.8 mg/kg/d in males and 15 mg/kg/d in females. In the first mouse carcinogenicity study, renal tubular adenomas and carcinomas were observed at all dose levels in males including the lowest dose of 125 mg/kg/d. In a subsequent study using the same strain of male mice, there were no treatmentrelated increases in the incidence of renal tumors up to a dose level of 100 mg/kg/d. Table 80.6 summarizes the key NOAELs observed in these studies.
80.8.1 MODE OF CARCINOGENIC ACTION
A mechanistic interpretation for the carcinogenicity of chlorothalonil has been published by Wilkinson and Killeen (1996).
80.8 Carcinogenicity
1751
Table 80.5 Summary of Genotoxicity Data Result
Reference
Test
Study type
In vitro
Ames test (hepatic activation)
Negative (± activation)
Banzer and Kouri, 1977
Ames test (renal activation)
Negative (± activation)
Jones et ai., 1984
Chromosomal aberration assay
Negative (+ activation)
Mizens et ai., 1986b
(CHO)
Positive (- activation)
Gene mutation assay (Chinese
Negative (± activation)
Kouri,1977
hamster V-79 cells & mouse fibroblast BALB/3T3 cells) DNA repair test (S. typhimurium)
Positive
Auletta and Kouri, 1977
DNA repair test (B. subtilus)
Negative
Shirasu, 1978
Cell transformation assay
Negative
Price and Bailee, 1979
Negative (rat)
Killeen and Siou, 1983
(F1706 P95 and H4536 P97 cells)
In vivo
Micronucleus test
Negative (mouse) Negative (Chinese hamster) Killeen, 1983
Negative (rat)
Chromosome aberration test
Negative (mouse) Negative (Chinese hamster) Siou et al., 1985
Chromosome aberration test
Acute study-negative
(Chinese hamster)
Subchronic study-
Covalent binding to DNA
Negative (rat)
Savides et al., 1987
Chromosome aberration test
Negative (Chinese
Proudlock, 1995
negative
hamster)
Table 80.6 Summary of Carcinogenicity Findings NOAEL
LOEL
Duration of study
Species
(mglkg bwtlday)
(mg/kg bwtlday)
2 years
Rat
None
40
2 years
Rat
3.8
IS (kidney tumors)
1.8
3.8 (kidney hyperplasia)
1.8
3.8 (stomach mucosal tumors)
2 years
Mouse
None
2 years
Mouse
5.4
I year oral
Dog
23 150
SO.S.1.1 Forestomach Thmors Repeated administration of chlorothalonil causes hyperplasia in the forestomach of rats and mice. The data are consistent with a temporal sequence of events starting with increased cell proliferation, multifocal ulceration and erosion of the forestomach mucosa, regenerative hyperplasia and hyperkeratosis, ultimately progressing to the formation of gastric tumors within the forestomach. Clear thresholds have been demonstrated for the induction of hyperplasia and neoplasia in both species. The
125 23 (kidney hyperplasia) 100 (stomach mucosal tumors) 500 (decreased body weight gain)
fact that chlorothalonil is not genotoxic provides reassurance that these tumors occur as a secondary consequence of local irritation within the rodent forestomach. Oral subchronic studies with the monoglutathione conjugate of chlorothalonil, failed to induce any toxicity in the rat forestomach, indicating that it is parent chlorothalonil, and not a metabolite, that is the toxic agent at this site. In contrast to the findings in the rat, there was no evidence of preneoplastic stomach lesions in dogs orally administered
1752
CHAPTER 80 Chlorothalonil
chlorothalonil for up to 1 year at dose up to 500 mg/kg/d, a dose level considerably higher than that which causes hyperkeratosis and hyperplasia in the rodent forestomach (approximately 4 mg/kg/d in the rat). The absence of any evidence of increased cell proliferation in the dog stomach was confirmed by PCNA labelling of tissue obtained at the termination of this study. The absence of stomach lesions in the dog is attributable to the anatomical differences between rodents and dogs in that dogs do not possess a forestomach. Similarly, humans are like the dog in that they do not possess an anatomical equivalent of the rodent forestomach. It is therefore concluded that the rodent forestomach tumors induced by chlorothalonil are not indicative of a carcinogenic risk to humans. 80.8.1.2 Renal Thmors The experimental data show a temporal sequence of events that lead to the formation of renal tumors in rats and male mice. Vacuolar degeneration of the renal proximal tubular epithelium has been shown to occur after two daily doses of chlorothalonil and cytotoxicity and degeneration of tubular epithelial cells can be seen after only 2 days treatment. Continued administration of chlorothalonilleads to the development of a regenerative hyperplasia within the renal proximal tubular epithelium. Continued regenerative hyperplasia ultimately results in progression of the kidney lesion to tubular adenoma and carcinoma. Thus, the data clearly show that initial cytotoxicity and regenerative hyperplasia within the proximal tubular epithelium are essential prerequisites for subsequent tumor development. Clear thresholds have been demonstrated for this nongenotoxic secondary mode of action which is a direct consequence of chronic stimulation of cell proliferation. Doses of chlorothalonil below the threshold for the induction of these preneoplastic lesions would not be expected to be carcinogenic. Studies have been conducted investigating the role of the glutathione conjugation pathway and subsequent formation of urinary thiol-derived metabolites in renal tumor formation. The central role of glutathione in the metabolism and subsequent toxicity of chlorothalonil has been shown by studies with a monoglutathione conjugate of chlorothalonil. Knowledge of the metabolism of other chemicals that undergo extensive glutathione conjugation implicates a role for mercapturic acidmediated metabolism for chlorothalonil. Maneuvers that inhibit key enzymes in this metabolic process, such as inhibition of the activity of y-glutamyltranspeptidase or the renal organic anion transporter, decrease the urinary excretion of thiol-derived metabolites in the rat. Administration of a monoglutathione conjugate to rats caused similar lesions in the kidney to parent chlorothalonil although no effects were seen in the forestomach. Studies undertaken in vitro using isolated kidney mitochondria have shown that respiration is inhibited in the presence of synthetic mono- and dithiol conjugates derived from chlorothalonil (Andre et aI., 1991 b; Savides et al., 1988). Furthermore, a correlation appears
to exist between the interspecies differences in susceptibility to renal toxicity and the differences in capacity to produce these thiol-derived metabolites as rats excrete more thiolderived metabolites than dogs. The proposed mode of action for the induction of renal toxicity in rodents is outlined in Fig. 80.2. It is concluded that chlorothalonil is a nongenotoxic kidney carcinogen in rats and mice and the NOAELs observed for both tumors and the precursor lesions indicate that it is appropriate to assume that a threshold exists for carcinogenicity. The species differences in metabolism are reflected in the different toxicity profiles seen in rodents and dogs and suggest that the dog is the most appropriate species for human health risk assessment and that these tumors are highly unlikely to develop in humans exposed to chlorothalonil.
Parent ChlorothaloniI
CI~CI CIYCN Cl
• •
GS-JrCI Gs-JrSG GS CIYCN
Biliary excretiY
Absorption Conjugation to GSH
GS
,
CIYCN
,
GIT- -
_> ,
Plasma metabolites
Liver Plasma
Cysteine con~t:J"
~ase
N-acety7
Thiol-derived conjugates
Kidney
)11'
Renal Toxicity
Urinary excretion Figure 80.2 Schematic outlining potential pathways of chlorothalonil metabolism in the rat that lead to formation of toxic metabolites within the kidney. Following absorption from the gastrointestinal tract, chlorothalonil is conjugated to glutathione in the liver. Further metabolic processing results in the formation of cysteine conjugates that may be detoxified via N -acetylase or activated to toxic thiol-derived species. GSH = glutathione, GIT = gastrointestinal tract.
80.10 Investigative Toxicity Studies
80.9 REPRODUCTIVE TOXICITY 80.9.1 DEVELOPMENTAL TOXICITY
The potential for chlorothalonil to induce developmental toxicity has been investigated in the rat and rabbit. In rabbits, maternal toxicity was evident at 20 mg/kg/d with body weight loss and decreased food consumption observed at this dose (Wilson et aI., 1988d). One death occurred in each of the mid and high dose groups. There were no adverse effects on the fetus and no treatment-related effects on the incidence of skeletal or visceral malformations. The NOEL for maternal toxicity was 10 mg/kg/d and the NOEL for developmental toxicity was 20 mg/kg/d. In rats, chlorothalonil was maternally toxic at 400 mg/kg/d with mortality, decreased body weight gain, and decreased food consumption at this dose (Mizens et aI., 1983). Food consumption was significantly decreased at all doses during days 6-9 of gestation and at the top dose on days 9-15. Food consumption returned to normal values on cessation of treatment with a compensatory increase seen at 25 and 100 mg/kg/d. There was a significant increase in the incidence of postimplantation loss due to early embryonic death at 400 mg/kg/d with a corresponding decrease in viable litter size. One rat at this dose level had reabsorbed 16 out of 17 implantation sites. However, exclusion of this animal from the statistical analysis still resulted in a significant increase in postimplantation loss compared to concurrent and historical controls. The NOELs for maternal and developmental toxicity were 100 mg/kg/d. It is concluded that chlorothalonil is not a developmental toxicant when tested up to doses that cause significant maternal toxicity and maternal death. 80.9.2 FERTILITY
In a two-generation reproductive toxicity study (Lucas et at., 1990), chlorothalonil caused a dose-related decrease in body weight gain which was evident at all doses in FO and Fl parental generations although achieving statistical significance at 1500
and 3000 ppm (68 and 145 mg/kg/d). No mortalities or clinical signs oftoxicity were observed in this study. Hyperplasia of the forestomach and kidney was observed at all doses in both parental generations with more marked effects in the Fl generation. Thus, a NOEL could not be established for parental toxicity. There were no adverse effects on reproductive performance or development including fertility indices, gestation length, litter size, number of live pups and stillborn pups, and pup survival. No gross malformations were observed which could be considered as treatment-related. There was a significant decrease in mean pup body weight on day 21 postpartum at 1500 (68 mg/kg/d) and 3000 ppm (145 mg/kg/d). This was only seen in the FIb litter at 1500 ppm but was seen consistently across all litters at 3000 ppm. Therefore, NOAEL for fetotoxicity is considered to be 1500 ppm (68 mg/kg/d). The NOEL for reproductive performance was 145 mg/kg/d with no effects at the top dose. It is concluded that chlorothalonil is not a reproductive toxicant. There was no evidence of reproductive toxicity in the absence of maternal toxicity. Therefore, the data are consistent with the view that the fetus and developing animal are not uniquely sensitive to chlorothalonil. Table 80.7 presents a summary of the reproductive toxicity studies. The key NOAELs for all toxicological endpoints are summarized in Table 80.8.
80.10 INVESTIGATIVE TOXICITY STUDIES 80.10.1 ACUTE EFFECTS ON HEPATIC AND RENAL GLUTATHIONE CONTENT
This study was designed to investigate and compare the time course effect of the acute oral administration of chlorothalonil on hepatic and renal glutathione (nonprotein sulfhydryl) content (Sadler and Ignatoski, 1985). At 5000 mg/kg chlorothalonil caused an decrease in body weight gain and liver weight which were evident 18 h after treatment. Within 9 h of treatment, hepatic glutathione levels were decreased and renal glutathione
Table 80.7 Summary of Reproductive Toxicity of Chlorothalonil NOAEL Study
Species
(mglkg bwtlday)
Developmental
Rat
Maternal and developmental toxicity-lOO (decreased maternal weight
Developmental
Rabbit
Maternal-IO (decreased weight gain at 20 mg/kg/d)
Two-generation
Rat
gain and increased incidence of resorptions at 400 mglkg/d) Developmental-20 (top dose) reproduction
1753
Parental-none (renal and forestomach lesions at all doses; LOAEL was 23 mglkg/d) Developmental--{j8 (decreased pup weight at day 21 at 145 mg/kg/d) Reproductive-145 (no effects at top dose)
1754
CHAPTER 80
Chlorothalonil
Table 80.8 Summary of Key Toxicological Endpoints Endpoint
Study 90 day rat diet
90 day dog
LOEL (mg/kg/d)
Increased kidney weight
1.5
Hyperplasia of forestomach
3
10
3
10
40
Increased kidney weight
3
9
Hyperplasia of forestomach
3
9
Renal hyperplasia
48
130
Decreased body weight gain
15
150
Decreased body weight gain
150
500
Renal hyperplasia 90 day mouse diet
NOEL (mglkg/d)
(capsule) I year dog (capsule) 2 year rat diet
2 year mouse oral
Renal hyperplasia
1.8
Renal tumors
3.8
Forestomach hyperplasia
1.8
Forestomach tumors
1.8
Renal hyperplasia
5.4
Renal tumors
99 (top dose)
3.8 15 3.8 3.8 23 None
Forestomach hyperplasia
1.6
5.4
Forestomach tumors
1.6
5.4
(squamous cell) 21 day dermal-rat
Rat reproductive
Local effects in skin
None
60 (lowest dose)
Systemic toxicity
None-but NOEL
60 (23% decrease
Parental NOEL (renal and
for renal
in body weight gain
hyperplasia of 600
at lowest dose)
None
23 (lowest dose)
forestomach hyperplasia)
toxicity
Developmental (decreased
68
145
100
400
100
400
10
20
20
None
pup weight at day 21) Developmental
Rat: Maternal (decreased weight
toxicity
gain) Developmental (increased resorptions) Rabbit: Maternal (decreased weight gain) Developmental
levels were elevated. The depletion of hepatic glutathione is considered a direct consequence of glutathione conjugation within the liver utilizing tissue resources. The increase in renal glutathione content is more difficult to explain but may be a consequence of urinary excretion of glutathione conjugates. 80.10.2 EFFECT OF DIETARY VS GAVAGE DOSING ON RENAL TOXICITY
IN THE RAT A study was conducted that was designed to compare the early morphological changes in the rat kidney following oral
administration of chlorothalonil by gavage with those following dietary administration (Ford et aI., 1988). Chlorothalonil was administered by gavage at 175 and 1750 mgikg in the diet (equivalent to 88 mgikg bw). Effects in the kidney were determined at 24, 48, 73, and 96 h postadministration. At the 48 h postadministration time point, vacuolar degeneration of the proximal tubular epithelium was observed in 2/3 animals that were gavaged. After 96 h all animals (gavage and diet) exhibited vacuolar degeneration of proximal tubular epithelium although the incidence of affected tubules was higher in gavaged animals than in those administered chlorothalonil in diet.
References
80.11 HUMAN DATA Most information concerning the effects of chlorothalonil in humans has been obtained from exposures arising in the manufacture and production and chlorothalonil. Health screening programs in such facilities have shown that the majority of effects documented following exposure to chlorothalonil were attributable to the irritant nature of the substance and included irritation to the skin, eyes, and respiratory tract. 80.11.1 DERMAL EFFECTS
There are a number of case reports in the published literature documenting occupational dermatitis in workers exposed to technical chlorothalonil. Skin reactions have also been documented in patch test studies with human volunteers. The main criticism of these reports is that they do not clearly discriminate between effects that may be a consequence of skin irritation and those that may represent a true sensitization response. Nevertheless, the weight of evidence suggests that chlorothalonil is a weak skin sensitizer in humans. Special skin surveys were conducted at the main chlorothalonil manufacturing plant to compare dermal findings in 1978 and 1979. In 1978,60% of the employees had some type of skin abnormality including 19 cases of contact dermatitis (McAmis, 1994a). In 1979, following the initiation of improved industrial hygiene measures in late 1978, there were no cases of contact dermatitis and only 21 % of the workers had some kind of skin abnormality. The most common abnormality was skin drying which was seen in 19 of the 26 employees with skin abnormalities (Chelsky, 1980a, b). A delayed irritant dermatitis has been documented which may occur up to 72 h after exposure and, although photosensitization reactions may occur, they are very rare events. 80.11.2 OCULAR EFFECTS
A review of clinical cases from employees exposed to chlorothalonil, at a packaging plant where the exposure was described as infrequent and light, was conducted in 1990 (Chelsky, 1990a). The purpose of this review was to assess the effects of chlorothalonil on the human eye. All of the ocular exposures to chlorothalonil involved intense pain with mild to moderate conjunctivitis and irritation of the corneal surface. Ocular edema was also seen in more extensive exposures. With lesser exposures, complete recovery occurred within 24 h. Recovery took slightly longer with following extensive exposure. In no instance was corneal opacity observed. 80.11.3 RESPIRATORY EFFECTS
Where respiratory effects have been noted these are generally consistent with the irritant properties noted in animal studies, although of a much less severe nature. In a review of medical records from workers at an independent facility used to
1755
grind technical chlorothalonil (Chelsky, 1990b, 1992; McAmis, 1994b), it was noted that, even in a workplace described as "dusty" and with workers wearing little protective clothing, ocular and dermal effects predominated, although a few cases of nasal and pharyngeal pain, burning, and soreness were noted. In a study conducted at another manufacturing facility, workers exposed to chlorothalonil showed a lower forced expiratory volume and higher incidences of nose and throat irritation, coughing phlegm, and shortness of breath than reference workers (Huang et aI., 1995). Thus, the information available from animal and human exposure indicates that chlorothalonil is irritating to the respiratory tract; a finding that is entirely consistent with the local site-of-contact toxicity seen in other epithelial tissues. 80.11.4 CLINICAL CASES AND POISONING INCIDENTS
Considering that chlorothalonil has been a commercial fungicide for over 25 years, there have been few reports of adverse effects in humans resulting from its use. The majority of the reported human effects have been related to the irritant properties of chlorothalonil. Of the reported skin effects, contact dermatitis is the most frequent diagnosis and this finding is almost exclusively in individuals exposed to chlorothalonil for prolonged periods (over 8 h) in an occupational environment. In summary, there have been a few reports in the literature of humans suffering adverse health effects following exposure to chlorothalonil (McAmis, 1995). Considering that chlorothalonil has been marketed for more than 25 years as a fungicide in agriculture, forestry, nursery plants, paints, and stains, the reports of adverse effects are very rare. The reported effects are associated with the irritant properties of the technical material.
REFERENCES Andre, J. c., et al. (1991a). "Comparison of the Effects of Dose Level and Vehicle on the Dermal Absorption of 14C-Chlorothalonil by Male Rats." Unpublished Syngenta study, Rep. 1698-88-0007-AM-001. Andre, J. C., et al. (199Ib). "Evaluation of Mitochondrial Function in the Presence and Absence of Sulfur-Containing Analogs of Chlorothalonil." Unpublished Syngenta study, Rep. 3113-88-0107-AM-001. Auletta, A., and Kouri, R. (1977). "Activity of DTX-77-0033 in a Test for Differential Inbibition of Repair Deficient and Repair Competent Strains of Salmonella typhimurium." Unpublished Syngenta study, Rep. OOO-STX-770033-002. Banzer, C. B., and Kouri, R. E. (1977). "Activity of Chlorothalonil in the Salmonella Microsomal Assay for Bacterial Mutagenicity." Unpublished Syngenta study, Rep. 000-STX-77-003S-001. Chelsky, M. (l980a). "Special Skin Surveys of Green's Bayou Plant Employees, 1978 and 1979." Confidential company medical report originally generated by Diamond Shamrock Corp. Chelsky, M. (l980b). "Skin Rashes among Green's Bayou Plant Employees." Confidential company medical report originally generated by Diamond Shamrock Corp. Chelsky, ~. (1990a). "Study of Chlorothalonil Plant Workers 1990. Evaluation of Potential for Persistent Effects on Eyes of Workers." Confidential company medical report originally generated by Diamond Shamrock Corp.
1756
CHAPTER 80
Chlorothalonil
Chelsky, M. (1990b). "Annual Employee Health Screening Greens Bayou Plant, 1986-1990. Special Reference to Chlorotbalonil Workers and tbe Respiratory System." Confidential company medical report originally generated by Diamond Shamrock Corp. Chelsky, M. (1992). "Annual Employee Healtb Screening Reports, Greens Bayou Plant, 1986-1991. Special Reference to Chlorotbalonil Workers and the Respiratory System." Confidential company medical report originally generated by Diamond Shamrock Corp. Fillmore, G., et al. (1993). "A 90-Day Oral Toxicity Study in Dogs witb Chlorothalonil." Unpublished Syngenta study, Rep. 521O-92-0l03-TX-003. Ford, W. H., et al. (1988). A 4-Day Study in Rats with Technical Chlorothalonil." Unpublished Syngenta study, Rep. 1095-86-0091-TX-002. Francis et al. (1973). "Daconil Technical Air Milled, Eye Irritation in the Albino Rabbit." Unpublished Syngenta study, Rep. 7948-95-3. Hironaka, M., et al. (1996). "Analysis of Hyperplastic Changes in tbe Stomach and Kidney of Male Rats after 28-day Induction by Chlorothalonil Technical." Unpublished Syngenta study, Rep. 3561. Huang, J., et al. (1995). Respiratory effects and skin allergy in workers exposed to tetrachloroisophthalonitrile. Bull. Environ. Contam. Toxieol. 55, 320-324. Jones, R. E., et al. (1984). "SalmonellalMammalian-Microsome Plate Incorporation Assay (Ames Test) with and without Renal Activation witb Technical Chlorothalonil." Unpublished Syngenta study, Rep. 694-5TX-84-0064-002. Killeen, J. C, Jr. (1983). "Research on the Possible Mutagenic Potentiality of Chlorothalonil by the Detection of Chromosomal Alteration in the Rat (Rat, Mouse and Hamster)." Unpublished Syngenta study, Rep. 000-5TX81-0025-001. Killeen, J. C., Jr., and Siou, G. (1983). "The Micronucleus Test in tbe Rat, Mouse and Hamster Using Chlorothalonil." Unpublished Syngenta study, Rep.000-5TX-81-0024-004. Kouri (1977). Lucas, E, et al. (1990). "A Two Generation Reproduction Study in Rats with Technical Chlorotbalonil." Unpublished Syngenta study, Rep. 1722-870121-TX-003. Magee, T. A., et al. (1990). "Study to Evaluate the Urinary Metabolites of Chlorothalonil Following Dermal Application to Male Rhesus Monkeys." Unpublished Syngenta study, Rep. 3382-89-0214-AM-001. Marciniszyn, J. P., et al. (1985a). "Pilot Study of the Biliary Excretion of Radioactivity Following Oral Administration of Chlorothalonil ( 14 C_ DS-2787) to Sprague-Dawley Rats." Unpublished Syngenta study, Rep. 633-4AM-83-0062-002. Marciniszyn, J. P., et al. (1985b). "Study of the Distribution of Radioactivity Following Oral Administration of 14C-Chlorothalonil ( 14 C-SDS-2787)." Unpublished Syngenta study, Rep. 631-4AM-84-0078-002. Marciniszyn, J. P., et al. (1986a). "Study of the Biliary Excretion of Radioactivity Following Oral Administration of 14C-Chlorothalonil ( I4 C-DS-2787) to Male Sprague-Dawley Rats." Unpublished Syngenta study, Rep. 633-4AM85-0012-002. Marciniszyn, J. P., et al. (1986b). "Pilot Study of the Effect of the GammaGlutamyl Transpeptidase Inhibitor, AT-125 on the Metabolism of 14C_ Chlorothalonil." Unpublished Syngenta study, Rep. 1376-86-0072-AM002. McAmis, R. J. (1994a). "Review of Dermal Chlorothalonil Exposures in Humans." Confidential company medical report originally generated by Diamond Shamrock Corp. McAmis, R. J. (1994b). "Review of Respiratory Chlorothalonil Exposures in Humans." Confidential company medical report originally generated by Diamond Shamrock Corp. McAmis, R. J. (1995). "Diagnosis of Poisoning, Specific Signs of Poisoning, Clinical Tests." Confidential company medical report originally generated by Diamond Shamrock Corp. Mead, R. L., et al. (1987a). "Analysis of Urine Samples from a 90-Day Feeding Yes No Study in Rats with tbe Monoglutatbione Conjugate of Chlorothalonil (T-117-11)." Unpublished Syngenta study, Rep. 1108-850078-TX-006.
Mead, R. L., et al. (1987b). "Analysis of Urine Samples from a 90-Day Feeding Study in Rats with Chlorothalonil (T-l17-11)." Unpublished Syngenta study, Rep. 1115-85-0079-TX-005. Mizens, M., and Laveglia (1994). "A Chronic (12-Month) Oral Toxicity Study in Dogs with Technical Chlorothalonil." Unpublished Syngenta study, Rep. 92-0457. Mizens, M., et al. (1983). "A Teratology Study in Rats with Technical Chlorothalonil." Unpublished Syngenta study, Rep. 517-5TX-82-00l1-003. Mizens, M., et al. (1986a). "A 21-Day Repeated Dose Dermal Toxicity Study Rats witb Technical Chlorothalonil." Unpublished Syngenta study, Rep. 6859-96-0113-TX-02. Mizens, M., et al. (1986b). "In Vitro Chromosomal Aberration Assay in Chinese Hamster Ovary (CHO) Cells witb Technical Chlorotbalonil." Unpublished Syngenta study, Rep. l109-85-0082-TX-002. Mizens, M., et al. (1996). "A 90-Day Pilot Study for the Evaluation Proliferation in tbe Kidneys of Male Rats Following the Oral Administration of Technical Chlorothalonil." Unpublished Syngenta study, Rep. 6704-96OOlO-TX-003. O'Meara, H. 0., and Laveglia, J. (1995). "Eye Irritation Study in Albino Rabbits with Technical Chlorothalonil." Unpublished Syngenta study, Rep. 6300-95-0083-TX -00 I. Moore (2000). Price, P., and BaIlee, D. (1979). "Analyses of Samples from Cell Transformation Studies for 2,4,5,6-Tetrachloroisophthalonitrile (Chlorothalonil, DS2787) and 4-Hydroxy-2,5,6-Trichloroiso-Phthalonitrile (DS-3701) (DTX77-0037 and DTX-77-0041)." Unpublished Syngenta study, Rep. 041-5TX79-0021-001. Proudlock, R. J. (1995). "In Vivo Bone Marrow Chromosomal Analysis in Chinese Hamsters Following Multiple Dose Administration of Technical Chlorothalonil." Unpublished Syngenta study, Rep. 6005-94-0047-TX-003. Ribovich, M. L., et al. (1982). "Balance Study of the Distribution of Radioactivity Following Oral Administration of 14C-Chlorothalonil ( 14 C-DS-2787) to Male Mice." Unpublished Syngenta study, Rep. 613-4AM-82-0178-001. Sadler, E. M., and Ignatoski, J. A. (1985). "Time Course of the Acute Effect of Technical Chlorotbalonil on Hepatic and Renal Glutatbione Content in Male Rats." Unpublished Syngenta study, Rep. 751-5TX-85-0032-001. Savides, M. C., et al. (1985). "Pilot Study for tbe Determination of the Effects of Probenecid Pre-treatment on Urinary Metabolites and Excretion of 14C-Chlorothalonil ( 14 C-SDS-2787) Following Oral Administration to Male Sprague-Dawley Rats." Unpublished Syngenta study, Rep. 621-4AM85-0035-001. Savides, M. C., et al. (1986a). "Study of tbe Distribution of Radioactivity Following Repeated Oral Administration of 14C-Chlorothalonil to Male Sprague-Dawley Rats." Unpublished Syngenta study, Rep. 1173-84-0079AM-003. Savides, M. C, et al. (1986b). "Identification of Metabolites in Urine and Blood Following Oral Administration of 14C-Chlorothalonil to Male Rats: Effects of Multiple Dose Administration on the Excretion of Thiol Metabolites in Urine." Unpublished Syngenta study, Rep. 62 1-4AM-83-006 1-002. Savides, M. C, et al. (1987). "Determination of tbe Covalent Binding of Radiolabel to DNA in the Kidneys of Male Rats Administered 14C-Chlorotbalonil ( 14 C-SDS-2787)." Unpublished Syngenta study, Rep. 1173-86-0096-AM002. Savides, M. C, et al. (1988). "A Study to Evaluate the Effects of SulfurContaining Analogs of Chlorothalonil on Mitochondrial Function." Unpublished Syngenta study, Rep. 1479-87-0037-AM-001. Savides, M. C., et al. (1989). "Study to Determine the Metabolic Pathway for Chlorothalonil Following Dermal Application to Rats." Unpublished Syngenta study, Rep. l625-87-0057-AM-001. Savides, M. C, et al. (1990). "Study to Evaluate tbe Urinary Metabolites of Chlorothalonil from Male Rhesus Monkeys." Unpublished Syngenta study, Rep. 3349-89-0179-AM-001. Savides, M. C., et al. (1995). "Study to Determine the Extent and Nature of Yes No Biliary Excretion of Chlorothalonil and/or Metabolites in the Dog." Unpublished Syngenta study, Rep. 5521-93-0319-AM-001. Shirasu, Y. (1978). "Mutagenicity Testing on Daconil in Microbial Systems." Unpublished Syngenta study, Rep. 000-5TX-61-0002-001.
References
Shuits and Wilson (1985). "Dennal Sensitisation Study (Closed Patch Repeated Insuit) in Guinea Pigs with Chlorothalonil 90DG Fonnulation." Unpublished Syngenta study, Rep. 707-5TX-84-0I26-002. Shuits, S. K., et al. (198Ia). "Acute Oral Toxicity (LDsO) Study in Rats with Technical Chlorothalonil." Unpublished Syngenta study, Rep. 296-5TX-800092-002. Shuits, S. K., et al. (l98Ib). "Acute Dennal Toxicity (LDsO) Study in Albino Rabbits with Technical Chlorothalonil." Unpublished Syngenta study, Rep. 296-5TX -80-0093-002. Shults, S. K., et al. (l98Ic). "Acute Inhalation Toxicity Study (Four Hour Exposure) in Rats with Technical Chlorothalonil (SDS-2787)." Unpublished Syngenta study, Rep. 296-5TX-80-0096-002. Shults, S. K., et al. (198Id). "Primary Dennal Irritation Study in Albino Rabbits with Technical Chlorothalonil." Unpublished Syngenta study, Rep. 2965TX -80-0094-002. Shuits, S. K., et al. (1983). "A 90 Day Feeding Study in Mice with 2,4,5,6-Tetrachloroisophthalonitrile (Chlorothalonil)." Unpublished Syngenta study, Rep. 6l8-5TX-83-0007-004. Shuits, S. K., et al. (1985). "Histopathologic Re-evaluation of Renal Tissue from a 90-Day Feeding Study in Mice with Technical Chlorothalonil." Unpublished Syngenta study, Rep. 753-5TX-85-0053-002. Shults, S. K., et al. (1986). "21-Day Repeated Dose Dennal Toxicity Study in Albino Rabbits with Technical Chlorothalonil." Unpublished Syngenta study, Rep. 754-5TX-85-0023-007. Shuits, S. K., et al. (1993). "Acute (Four-Hour) Inhalation Toxicity (LCso) Study in Rats with Hammer-Milled Technical Chlorothalonil." Unpublished Syngenta study, Rep. 5290-92-0l60-TX-002. Siou, G., et al. (1985). "Acute and Subchronic In Vivo Bone Marrow Chromosomal Aberration Assay in Chinese Hamsters with Technical Chlorothalonil." Unpublished Syngenta study, Rep. 625-5TX-83-00l4-003. Tucker, S. B. (1986). "Skin Sensitisation Studies with Chlorothalonil Conducted at the Department of Occupational Dennatology, University of Texas." Unpublished Syngenta studies, Reps. 5TX-84-0023, 5TX-84-0027, 5TX-84-00I2, 1094-84-00l2-DA002, 5TX-84-0044, 5TX-84-0076, and 5TX-84-0045. Ward, R. J. (1989a). "Chlorothalonil: In Vitro Absorption from Technical Material through Human Epidennis." Unpublished Syngenta study, Central Toxicology Laboratory, Rep. CTLIPI2640. Ward, R. J. (1989b). "Chlorothalonil: In Vitro Absorption from Bravo 720 Formulation through Human Epidennis." Unpublished Syngenta study, Central Toxicology Laboratory, Rep. CTL1P12880. Wazeter and Lucas (1971). "Acute Intraperitoneal Toxicity (LDSO) in Male Albino Rats of Technical Chlorothalonil." Unpublished Syngenta studies, Rep. 000-5TX -7 I -0006-00 1. Wilkinson, C. F, and Killeen, J. C. (1996). A mechanistic interpretation of the oncogenicity of chlorothalonil in rodents and an assessment of human relevance. Regulatory Toxicol. Pharmacol. 24, 69-84. Wilson, P. D. (l977a). "Primary Eye Irritation Study in Rabbits." Unpublished Syngenta study, Rep. DTX-77-0075.
1757
Wilson, P. D. (l977b). "Primary Eye Irritation Study in Rabbits." Unpublished Syngenta study, Rep. DTX-77-0059. Wilson, P. D. (1977c). "Primary Eye Irritation Study in Rabbits." Unpublished Syngenta study, Rep. DTX-77-0069. WiIson, N. H., et al. (l982a). "Dennal Sensitisation Study in Hartley-Derived Guinea Pigs with Technical Chlorothalonil." Unpublished Syngenta study, Rep. 394-5TX-81-0132-002!7020. Wilson, N. H., et al. (1982b). "Four Week Dietary Range-Finding Study in Rats with Technical Chlorothalonil." Unpublished Syngenta study, Rep. 0995TX-8I-0174-003. Wilson, N. H., et al. (1983a). "A 90-Day Toxicity Study of Technical Chlorothalonil in Rats." Unpublished Syngenta study, Rep. 099-5TX-80-0200006. Wilson, N. H., et al. (I 983b). "A Subchronic Toxicity Study of Technical Chlorothalonil in Rats." Unpublished Syngenta study, Rep. 562-5TX-810213-004. Wilson, N. H., et al. (1983c). "A Chronic Dietary Study in Mice with Technical Chlorothalonil." Unpublished Syngenta study, Rep. 108-5TX-79-0102-004. Wilson, N. H., et al. (1984). "A Subchronic Toxicity Study of Technical Chlorothalonil in Rats (Electron Light Microscopy of Kidneys)." Unpublished Syngenta study, Rep. 562-5TX-8-02I3-004-001. Wilson, N. H., et al. (1985a). "Histopathologic Re-evaluation of Renal Tissue from a 90-Day Toxicity Study in Rats with Technical Chlorothalonil." Unpublished Syngenta study, Rep. 753-5TX-85-0055-002. Wilson, N. H., et al. (1985b). "Histopathologic Re-evaluation of Renal Tissue from a Subchronic Toxicity Study of Technical Chlorothalonil in Rats." Unpublished Syngenta study, Rep. 753-5TX-85-0056-002. Wilson, N. H., et al. (1985c). "A Tumourgenicity Study of Technical Chlorothalonil in Rats." Unpublished Syngenta study, Rep. 099-5TX-80-0234008. Wilson, N. H., et al. (1987). "A Tumourgenicity Study of Technical Chlorothalonil in Male Mice." Unpublished Syngenta study, Rep. 1099-84-0077TX-006. Wilson, N. H. (l988a). "Guinea Pig Maximization Test with Technical Chlorothalonil (T-117-II)." Unpublished Syngenta study, Rep. 1094-84-0044-TX001. Wilson, N. H. (I 988b). "Guinea Pig Epicutaneous Test Involving Chlorothalonil in Acetone." Unpublished Syngenta study, Rep. 1094-84-0045-TX001. Wilson, N. H. (1988c). "Guinea Pig Maximization Test with Technical Chlorothalonil (T-117 -11 )." Unpublished Syngenta study, Rep. 1094-84-0076-TX001. Wilson, N. H., et al. (l988d). "A Teratology Study in Rabbits with Technical Chlorothalonil." Unpublished Syngenta study, Rep. I 544-87-0060-TX-002. Wilson, N. H., et al. (1989). "A Tumourgenicity Study of Technical Chlorothalonil in Rats." Unpublished Syngenta study, Rep. 1102-84-0J03-TX007.
CHAPTER
81 Dialkyldithiocarbamates (EBDCs) Susan Hurt, Janet Ollinger Rohm and Haas Company
Gail Arce Griffin LLC
Quang Bui Cerexagri, Inc.
Abraham J. Tobia Aventis
Bennard van Ravenswaay BASFAG
81.1 CHEMISTRY AND FORMULATIONS
EBDCs can also be sold as a premix with various blending partners.
Ethylenebisdithiocarbamates (EBDCs) are a group of fungicides that have been used widely throughout the world since the 1940s to protect a wide variety of crops against fungal disease. There are five members of the class, specifically mancozeb, maneb, metiram, zineb, and nabam. All members have an ethylenebisdithiocarbamate backbone, with different metals associated with the individual compounds. The structure of each compound is shown below. Mancozeb: Maneb: Metiram: Zineb: Nabam:
[-MnSC(:S)NHCH2CH2NHC( :S)S- ]xZny, where xjy = 11 [-MnSC(:S)NHCH2CH2NHC(:S)S- ]x [[ -(NH3)Zn-S-C(:S)NHCH2CH2NHC(:S)S- 13 -S-C(:S)NHCH2CH2NHC(:S)S- ]x [-ZnSC(:S)NHCH2CH2NHC(:S)S- ]x [NaSC(:S)NHCH2CH2NHC(:S)SNa]
The molecular weights of the individual EBDCs are mancozeb-271, maneb-265, metiram-1088.6, zineb-275, and nabam-256. At this time mancozeb, maneb, and metiram are the most widely used EBDCs. Zineb is used to a lesser degree and nabam is no longer used in agriculture. Thus, this chapter will focus on mancozeb, maneb, and metiram. The EBDCs are sold as wettable powder, dry flowable (also called water dispersable granules), and flowable formulations. Handbook of Pesticide Toxicology Volume 2. Agents
81.2 USES EBDCs are used to control about 400 fungal pathogens on more than 100 crops. The major EBDC uses around the world include grapes (fresh grapes, grapes grown for juice, and grapes grown for wine), potatoes, citrus, apples, tomatoes, melons, and bananas. EBDCs are also important products for disease control in corn, cereal grains, leafy vegetables, brassica vegetables, cranberries, onions, peanuts, sugar beets, asparagus, and nuts as well as for many other critical crops that are grown on a lower amount of acreage. Diseases of turf and ornamental crops are also controlled by EBDCs. Some of the economically important diseases controlled include early and late blight, downy mildew, and bacterial diseases. EBDCs are key components of fungicide resistance management programs because they have a multisite mode of action. For example, EBDCs deactivate the sulfhydryl containing enzymes which mediate numerous biosynthesis, mechanical, and transport activities within the fungal cytoplasm. They also inactivate ATP production, the Krebs cycle, the enzymes which convert glucose to pyruvate, and enzymes which convert amino and fatty acids to acetylcoenzyme A. Thus, resistance will not develop. After over 40 years of use no resistance has developed to any of the EBDCs.
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Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
1760
CHAPTER 81
Dialkyldithiocarbamates (EBDCs)
81.3 HAZARD IDENTIFICATION The toxicology database supporting the assessment of the potential health risks of the EBDCs and their common metabolite ethylenethiourea (ETU) has been upgraded in recent years with a complete set of modern studies of mancozeb, maneb, and metiram conducted in full compliance with OECD and other applicable national and international guidelines and internationally recognized good laboratory practices. These newer studies have superseded the older studies in the published literature and now form the core of the toxicology database relevant for the hazard identification and dose-response assessment of this family of fungicides. 81.3.1 PHARMACOKINETICS AND METABOLISM
Studies of the pharmacokinetics and metabolism of mancozeb, maneb, and metiram in laboratory animals have indicated that the EBDCs are only partially absorbed, then rapidly metabolized and excreted with no evidence of long-term bioaccumulation. Absorption of oral doses is rapid. Most of the administered dose is excreted within 24 hours, with about half eliminated in the urine and half in the feces. Biliary excretion is minimal, indicating that only approximately 50% of oral doses are absorbed. Only low level residues are found in tissues, principally in the thyroid. ETU is the major metabolite. On average 7.5% of an EBDC dose administered to rats is metabolized to ETU on a weight basis. The bioconversion factor in mice is slightly smaller at 5 to 6% (Cameron et aI., 1990; DiDonato and Longacre, 1986; Emmerling, 1978b; EPA, 1992; Hawkins et aI., 1985; Kocialski, 1989; Longacre, 1986; Nelson, 1986, 1987; Piccirillo et aI., 1992; Puhl, 1985). The spectrum of metabolites produced in laboratory animals points to two common metabolic pathways (Fig. 81.1), which both lead ultimately to the formation of glycine and incorporation into natural products. In the predominant pathway quantitatively, the dithiocarbamate linkages are hydrolyzed to produce ethylenediamine (EDA) directly, and EDA is oxidized to glycine,joining the intermediary metabolic pool at this point. The other pathway is responsible for the toxic effects of the EBDCs and involves oxidation to ethylenebisisothiocyanatesulfide and then to ETU, various derivatives of ETU, and ethyleneurea (EU) before rejoining the main pathway with conversion to EDA, glycine, and other natural products. ETU metabolism has also been extensively studied in multiple species. As with the EBDCs, oral doses are rapidly absorbed and rapidly excreted, although in this case primarily in the urine and more quickly in mice than in rats. In most species the greater majority (70% or more) of an oral dose is eliminated via the urine within 48 hours. Concentrations in blood and tissues are generally at comparably low levels with the exception of somewhat higher levels in the thyroid; levels in maternal and fetal tissues were similar 3 hours after dosing. Half-lives for elimination from maternal blood were 5.5 and 9.4 hours in
mice and rats, respectively. Unchanged ETU was the principal metabolite in rats and guinea pigs, with small amounts of EU. In mice the principal identified metabolites were ETU and imidazolinylsulfenate, and in cats, S-methyl ETU was the principal metabolite (DiDonato and Longacre, 1987; Emmerling, 1978b; Iverson et aI., 1977, 1980; Jordan and Neal, 1979; Kato et aI., 1976; Peters et aI., 1982; Ruddick et aI., 1976a, 1977; Teshima et aI., 1981). Studies of dermal absorption of the EBDCs have been challenging due to the difficulty of small scale preparation of radiolabelled samples representative of their complex polymeric structures. Thus, the reported values of 0.2-6.5% are considered to be overestimates of their actual dermal absorption potential under conditions of use (Craine, 1991; Haines, 1980; Hawkins et aI., 1984; Tomlinson and Longacre, 1988). Dermal absorption of ETU increased from 5 to 22% with decreasing applied skin concentrations (DiDonato and Longacre, 1987). 81.3.2 ACUTE TOXICITY
The EBDC's have very low acute toxicity by the oral, dermal, and respiratory routes (Table 81.1). The World Health Organization (WHO) has classified mancozeb, maneb, and metiram as unlikely to present an acute exposure hazard under conditions of normal use (WHO, 1994). Although not irritating to skin on initial contact and only slightly irritating to eyes and mucous membranes, prolonged or repeated skin contact may result in dermatitis due to their weak sensitization potential. ETU is only slightly toxic after oral administration, but it is a moderate to weak sensitizer in the guinea pig maximization test (Matsushita et aI., 1976, 1977). 81.3.3 SHORT· AND LONG·TERM TOXICITY AND ONCOGENIC POTENTIAL
The EBDCs share a comparable toxicological profile, primarily based on the toxic effects of their common ETU metabolite. A summary of the critical doses and effects in subchronic and longer term studies is presented for each of the EBDCs in Table 81.2, and a similar summary for ETU is presented in Table 81.3. Results are collected by study type. As is illustrated repeatedly in these tables, the principal target organ upon repeated exposure to all of the EBDCs is the thyroid, which is also the principal target organ of ETU. For example, all three EBDCs (mancozeb, maneb, and metiram) and ETU altered thyroid hormone levels and/or weights at the lowest affected dose after three months of dietary feeding in rats. Most of the other organs affected, such as the liver at generally higher doses or red blood cells usually in dogs, are also common to ETU. Prolonged dietary feeding of ETU produces thyroid and pituitary tumors in rats and mice, and liver tumors in mice. As normally occurs with toxicological effects due to formation of a metabolite, the effects of ETU are not as strong in the EBDCs. When EBDCs are administered, much higher doses are
81.3 Hazard Identification
1761
Metabolic Pathway of EBDC
( ~J(s.
5.----N~ H
5
EBDC
ETU
1
CH 3 CONH
o
~
NH2
U
;t:~
N-acetyl EDA
HOCHN
N,
N'~I
HOCHN
7.:EthYI~mine
NHCOH
HN
NH
EDA
NH2
U
1
N-formyl EDA
Interconversion to other
amino acids Proteins Nucleic Acids
Figure 81.1
Pyruvate Acetyl Co-A Citrate
Sugars Saccharides Fatty Acids Lipids
Metabolic pathway of EBDC.
required to produce adverse effects, and the effects themselves are generally not as pronounced or may be precluded altogether by high dose limitations.
81.3.3.1 Thyroid Effects The effects of the EBDCs on the thyroid after either short- or long-term dietary administration are consistent on both a quantitative and a qualitative basis with those of ETU. As described below, it is well accepted that these effects are the result of a secondary mechanism (hormonal imbalance), and that there is a threshold for the resulting tumors Similarly to the structurally related thionamide drugs (propylthiouracil, methimazole, and carbimazole) which are used clinically for treatment of hyperthyroidism in humans, the primary toxicological finding with
ETU in laboratory animals is inhibition of the synthesis of thyroid hormones, thyroxine (T4) and triiodothyronine (T3), leading to elevated serum levels of thyroid stimulating hormone (TSH) via feedback stimulation of the hypothalamus and pituitary (Atterwill and Aylard, 1995; Engler and Burger, 1984; O'Neil and Marshall, 1984) (see Fig. 81.2). Prolonged and continuous elevation of serum TSH levels results in hypertrophy and hyperplasia of the thyroid follicular cells in rats, mice, hamsters, monkeys, and dogs (Briffaux, 1991, 1992; Gak et aI., 1976; Chhabra et aI., 1992; Freudenthal et aI., 1977a; Graham and Hansen, 1972; Graham et aI., 1973, 1975; Leber et aI., 1978; Q'Haraand DiDonato, 1985; Schmid et aI., 1992; Ulland et aI., 1972), and ultimately in the development of follicular nodular hyperplasia, adenoma, and/or carcinoma in rats and
1762
CHAPTER 81
Dialkyldithiocarbamates (EBDCs)
Table 81.1 Acute Toxicity of Mancozeb, Maueb, Metiram, and ETU Type of study
Active
acute
ingredient
Acute oral, rat
LD(LC)50 Strain-sex
mg/kg bw (mgIL)
Reference
mancozeb
F344,M
>5000
Watts and Chan, 1984a, b
mancozeb
CRCD,M
>5000
DeCrescente and Parsons, 1980
maneb
Cr1:CD BR, MIF
>5000
Naas, 1989a
metiram
SD,MIF
6500-10,000
Jackh, 1981; Leuschner, 1979a;
Acute oral, mouse
mancozeb
>5000
Watts and Chan, 1984b
Tntraperitoneal, rat
mancozeb
B6C3F\, M Wistar, MIF
380
DeGroot,1974
metiram
SD,MIF
318
Hofmann and Munk, 1975
Intraperitoneal, mouse
metiram
NMRI,MIF
80-215
Hofmann, 1974; Leuschner, 1979b
Hofmann, 1985; Hofmann, 1975
Acute dermal, rat
metiram
SD, MIF
>2000
Grundler, 1979
Acute dermal, rabbit
mancozeb
NZW,M
>5000
DeCrescente and Parsons, 1980
maneb
NZW,MIF
>2000
Naas, 1989b
mancozeb
COBS-CR (SD) BR, MIF
5.14 mg/l
Hagan and Baldwin, 1982
Acute inhalation (4 hr)
Acute oral, rat
maneb
Cr1:CD BR, MIF
7.38 mg/l
Terrill, 1990
metiram
SD,MIF
>5.7 mg/l
Klimisch and Zeller, 1980
ETU
MIF
545-ca.2400
Peters et aI., 1980a; Graham and Hansen, 1972; Lewerenz and Plass, 1984; Teramoto et aI., 1978
Acute oral, mouse
ETU
F (13 days pregnant)
600
ETU
MIF
ca. 2400-4000
Khera,1987 Lewerenz and Plass, 1984; Teramoto et al., 1978; Peters et aI., 1980b
Acute oral, hamster
ETU
F (9 days pregnant)
>3000
Khera,1987
ETU
F
>3000
Teramoto et aI., 1978
ETU
F (11 days pregnant)
>2400
Khera,1987
Thyroid-Pituitary Feedback Mechanism
Iodine .....
TRH
r-------Ar4 Hypothalamus
=Triiodothyronine (Thyroid Hormone) =Tetraiodothyronine (Thyroid Hormone) TRH =Thyrotropin Releasing Hormone T3 T4
TSH = Thyroid Stimulating Hormone
Figure 81.2
Thyroid-pituitary feedback mechanism.
81.3 Hazard Identification
1763
Table 81.2 EBDCs: Critical Findings in the Most Relevant Studies
Study
Active
NOAEL
LOAEL
Effects observed at
(mg/kg/d)
(mg/kg/d)
LOAELlcritical results
Reference
18
180
Decreased body weight and liver
O'Hara and DiDonato, 1985
Mouse 3 month oral mancozeb
MFO activity; thyroid follicular hypertrophy and hyperplasia metiram
84
302
Decreased T4
Gelbke et aI., 1992a
Rat 3 month oral mancozcb
7.4
15
Decreased T4, increased TSH
Goldman et aI., 1986
maneb
5
25
Increased thyroid weight,
Trutter, 1988a
metiram
6
20
Decreased T4 and iodine uptake,
follicular hyperplasia Hunter et aI., 1977
increased thyroid weight; microscopic changes in muscle fibers 5.8
23.2
Slight anemia; decreased T4,
Gelbke et aI., 1992b
increased thyroid weight, general muscle weakness/ataxia and reduced grip strength without histopathological correlate at 70 mg/kg body weightld Rat I month dermal mancozeb
1000
none
No systemic toxicity
Trutter, 1988c
Rabbit 1 month dermal maneb
100
300
Increased follicular colloid
Trutter, 1988b
metiram
250
none
No systemic toxicity
Ullman et aI., 1987
79 mg/m 3
326 mg/m 3
Decreased weight gain, T4levels;
Hagan and Baldwin, 1986;
x6 hr/d
x6 hr
thyroid hyperplasia. All effects
(= 8.3)
(= 34)
reversible after 13 weeks recovery
10mg/m3
30mg/m3
x6 hr/d
x6 hr/d
2mg/m 3
20mg/m 3
x6 hr/d
x6 hr/d
Rat 3 month inhalation mancozeb
maneb metiram
Decreased weight gain; all effects
Hagan et aI., 1986 Ulrich, 1986a, 1987
reversible after 13 weeks recovery Decreased weight gain; alveolar
Ulrich, 1986b
macrophage accumulation due to nonspecific dust reaction
Dog 3 month mancozeb
3
30
Decreased weight gain, RBC
Cox, 1986
parameters maneb
3.7
15
Thyroid follicular hyperplasia
AlIen et aI., 1989
1.7
10.2
Decreased weight gain; no effect on
Muller, 1992
Rat two generation mancozeb
reproduction below adult toxic levels 7
70
Decreased weight gain, feed
Solomon et aI., 1988
consumption; increased thyroid, liver, kidney weights, microscopic changes in thyroid, kidney, and pituitary; no reproductive effect. maneb
5.6
22.4
Increased liver, kidney weight ratios,
Ryle et aI., 1991
thyroid follicular hyperplasia; no reproductive effect below toxic levels metiram
1.8
14.4
Decreased body weight, feed
Cozens et aI., 1981
consumption; no reproductive effect
(continues)
1764
CHAPTER 81
Dialkyldithiocarbamates (EBDCs)
Table 81.2 (continued)
Study
Active
NOAEL
LOAEL
Effects observed at
(mglkgld)
(mg/kgld)
LOAELlcritical results
Reference
2000
NA
No adverse effect
Nemec, 1993
8.2
48
Decreased feed consumption;
Stadler, 1991
Rat acute neurotoxicity maneb Rat 90-day neuropathology mancozeb
neurohistopatho1ogical changes Rat developmental mancozeb
32
128
Decreased maternal weight gain, feed
Gal10 et al., 1980
consumption; teratogenic NOAEL; teratogenicity at 512 mg/kg body weightld 60
360
Decreased maternal weight gain, feed
Tesh et al., 1988
consumption; maternal "reeling gait" and hindlimb paralysis, embryofetotoxicity maneb
20
100
Decreased maternal weight gain, feed
Nemec, 1992
consumption; embryofetotoxicity 100
500
Decreased maternal weight gain, feed
Kapp et al., 1991
consumption; hind1imb paresis; embryofetotoxicity and teratogenicity metiram
80
160
Decreased maternal weight gain, slight
Palmer and Simons, 1979
decreases in litter size and weight Rabbit developmental mancozeb
55
100
Maternal weight loss, decreased feed
Muller, 1991
consumption, increased abortions; no adverse embryofetal effects 30
80
Decreased maternal weight gain, feed consumption, litters; increased
Solomon and Holz, 1987; Solomon and Lutz, 1987
abortions, clinical signs, and deaths; no adverse embryofetal effects metiram
10
40
Decreased maternal body weight, feed
Gelbke et al., 1988
consumption: increased abortions; no adverse embryofetal effects Dog 12 month mancozeb
7
28
Decreased weight gain, RBC
Shaw, 1990
parameters, inc, cholesterol 2.3
23
Decreased weight gain, feed
Broadmeadow, 1991a, b
consumption, T4 maneb
6.4
32
Thyroid thickening and
Corney et al., 1992
enlargement, follicular hyperplasia metiram
2.5
31
Decreased T4, increased thyroid
Comey et aI., 1991
size and follicu1ar hyperplasia, focal hepatic lipofuscin deposition, slight anemia, diarrhea, and blood biochemical changes Monkey 6 month maneb
7.3
22
Increased thyroid weight
Leuschner et aI., 1977
metiram
5
15
Decreased T3, T4, increased
Sortwell et aI., 1979
thyroid weight and follicular hyperplasia (continues)
81.3 Hazard Identification
1765
Table 81.2 (continued)
Study
Active
NOAEL
LOAEL
Effects observed at
(mg/kg/d)
(mglkg/d)
LOAELlcritical results
Reference
17
170
Decreased weight gain
Shellenberger, 1991
13
130
Decreased weight gain, T3, T4
Everett et aI., 1992
11
44
Decreased body weight, T4;
Tompkins, 1992
Mouse oncogenic mancozeb maneb
hepatocellular adenomas at 440 mg/kg body weight/day metiram
24
79
Decreased body weight
Hunter et aI., 1979
4.8
29
Decreased weight gain, T3, T4;
Stadler, 1990
Rat chronic-oncogenic mancozeb
increased TSH, thyroid weight, follicu1ar cell hypertrophy, hyperplasia, nodular hyperplasia, adenoma and carcinoma 4
16
Decreased weight gain, T4;
Hooks et aI., 1992
increased height of thyroid follicular epithelium, prominent microfollicles maneb
20
67
Decreased body weight, T4; increased 131 I half-life, thyroid
Leuschner et aI., 1979, 1986a, b; Leuschner, 1991
weight metiram
3.1
12
Muscular atrophy
Hunter et aI., 1981
NOAEL = no observed adverse effect level. LOAEL = lowest observed adverse effect level.
mice (Chhabra et aI., 1992; Graham et aI., 1973,1975; Schmid et aI., 1992; Ulland et aI., 1972), but not in hamsters (Gak et aI., 1976). There is evidence for reversibility of the thyroid effects (Arnold et aI., 1983). The mechanism of the crucial early steps of thyroid tumor formation by ETU is well understood. ETU reversibly inhibits thyroid peroxidase-catalyzed iodination and coupling of tyrosine residues into the thyroid hormone precursor thyroglobulin in vivo (Hill et aI., 1989). Direct evidence for inhibition of thyroid hormone synthesis by ETU has been obtained in rats in vivo (Arnoldet aI., 1983; O'Neil and Marshall, 1984). ETU also reversibly inhibited thyroid peroxidase-catalyzed iodination reactions in vitro (Doerge and Takazawa, 1990). The correlation of the hormonal changes with hyperplasia and neoplasia has been clearly demonstrated in specific studies of ETU (Chhabra et aI., 1992; Freudenthal et aI., 1977a). Similarly, the long-term stimulation of the pituitary via hypothalamic thyrotropin-releasing hormone also results in morphologic changes in the pituitary of rats, mice, monkeys, and dogs (Briffaux, 1991; Chhabra et aI., 1992; Leber et aI., 1978; Schmid et aI., 1992), culminating in adenomas of the pars distalis after two years exposure in mice and rats (Chhabra et aI., 1992; Schmid et aI., 1992). A regenerative, nonhemolytic anemia that was observed in dogs is considered to be a secondary manifestation of the primary thyroid condition, since anemia is a known manifes-
tation of hypothyroidism in dogs (Duncan et aI., 1994; lain, 1986). The mechanistic linkage between the prolonged disruption of the hypothalarnic-pituitary-thyroid (HPT) axis and thyroid neoplasia has been confirmed in studies of sulfamethazine (Hard, 1998). Indeed, thyroid gland neoplasia can be induced experimentally in laboratory animals simply by a a low iodine diet (Schaller and Stevenson, 1966), i.e., without exogenous agents, indicating that the neoplasia was induced by an internal factor. Supplementing the diet with thyroid hormone abolishes the neoplastic response indicating that the internal factor is TSH (Doniach, 1970). Thus, the sequence of events relating thyroid hormone inhibition via hormonal imbalance to the onset of pituitary and thyroid follicular neoplasia in rodents is well characterized, resulting in the inference that the threshold for the early steps in the sequence, particularly the key elevation of TSH levels, is necessarily a threshold for the remaining steps in the process including carcinogenesis. For purposes of human oncogenic risk assessment, the principle of the existence of a threshold for thyroid and pituitary neoplasia resulting from thyroid inhibition has been accepted (EPA, 1998; Hard, 1998; Hill et aI., 1989, 1998; IPCS, 1990), and the specific relevance of this threshold mechanism to ETU is also accepted (EPA, 1992, 1998; Hurley et aI., 1998).
1766
CHAPTER 81
Dialkyldithiocarbamates (EBDCs)
Table 81.3 ETU: Critical Findings in the Most Relevant Studies Type of study/
NOAEL
LOAEL
Effects observed at
species
(mglkgld)
(mg/kgld)
LOAELlcritical results
Reference
1.7
17
Thyroid follicular hyperplasia and
O'Hara and DiDonato, 1985
Subchronic dietary Mouse 90 day
decreased colloid density in both sexes, and increased liver weights in females Rat 90 day
1.7
8.5
Altered thyroid function and
Freudenthal et aI., 1977a
follicular hyperplasia Dog 90 day
0.39
6.0
Decreases in rbc parameters and
Briffaux, 1991
increased cholesterol; thyroid and other effects at 80 mglkg/d Monkey 6 month
0.1-0.5
2.5
Increased iodine uptake, thyroid
Leber et aI., 1978
follicular hyperplasia and pituitary hypertrophy Chronic dietary 1 yr dog
0.18
1.8
Slightly reduced body weight gain,
B riffaux, 1992
increased thyroid weights and hypertrophy with colloid retention, pigment accumulation in the liver 2 yr rat
0.25
1.25
Thyroid vacuolarity and hyperplasia
Grahametal., 1973, 1975
2 yr rat
0.37
9.25
Thyroid, pituitary and liver effects,
Schmid et aI., 1992
thyroid and pituitary tumors, decreased body weights in males Oncogenicity Mouse 2 year
<17
17
Decreased T4, increased TSH and diffuse thyroid foIIicular
Chhabra et aI., 1992; NTP,1992
cytoplasmic vacuolation at 17 mglkglday; tumors of the thyroid, liver and/or pituitary at 56 mg/kglday and higher Rat 2 year
<1.1
1.1
Decreased T4, increased TSH and thyroid foIIicular hyperplasia at
Chhabra et aI., 1992; NTP,1992
1.1 mg/kglday; thyroid tumors at 3.7 mglkg/day and higher Reproductive Rat 2 generation
0.11-0.43
1.1-4.3
Thyroid follicular hypertrophy and
Dotti, 1992
hyperplasia; no reproductive effect at 4.3-21 mg/kg/day, HDT
(continues)
In addition, in comparison to laboratory animals, humans are expected to exhibit a lesser degree of sensitivity to thyroid inhibitors (Costigan, 1998; Hard, 1998; Hill et aI., 1998). The reasons for this are threefold: First, humans possess a substantial reserve supply of thyroid hormone, much of which is carried in the serum bound to thyroxine-binding globulin, a serum protein that is missing in laboratory rodents (Odell et aI., 1967). Therefore, release of stored thyroid hormones maintains normal serum levels for weeks in euthyroid humans (Martindale, 1972) and for weeks to several months in hyperthyroid individ-
uals (Odell et aI., 1967), despite daily doses of antithyroid drugs sufficient to completely block synthesis. This protein is missing in rodents, resulting in comparatively rapid hormone turnover, normally higher levels of TSH, and increased sensitivity to the effects of hormone depletion. Second, the molar concentrations of thiourea compounds required to inhibit thyroid peroxidase activity are far smaller in rats than in monkeys or humans (Takayama et aI., 1986), indicating that humans must be exposed at much higher levels to achieve the same degree of enzyme inhibition. Third, under conditions of prolonged thyroid
81.3 Hazard Identification
1767
Table 81.3 (continued) Type of study/
NOAEL
LOAEL
Effects observed at
species
(mg/kgld)
(mg/kg/d)
LOAELlcritical results
Reference
5
10
Anomalies of the brain, neural tube
Khera,1973
Developmental toxicity Rat
and tail at 10 mglkglday; a higher frequency of delayed ossification of the parietal bone at 5 mg/kg/day judged consistent with or close to a NOAEL at this level; maternal NOAEL = 40 mg/kg/day based on maternal lethality at 80 mg/kg/day Rat
10
20
Dilation of the lateral ventricle
Teramoto et aI., 1978
Rat
5
10
Decreased fetal weights;
Chernoff et aI., 1979
hydrocephalus at 20 mg/kg/day. Maternal NOAEL = 40 mg/kg/day based on maternal deaths and reduced weight gain at 80 mg/kg/day Rat perinatal
25
30
Hydrocephalus, failure to nurse,
(exposure from day 7
increased open field activity in
gestation to day IS
males
Chernoff et aI., 1979
postpartum) Rat
IS
25
Dilation of brain ventricles;
SaiIIenfait et aI., 1991
maternal NOAEL = 35 mglkglday, HDT Rabbit
40
80
Increased resorptions, decreased
Khera, 1973
brain weights, degeneration of fetal kidney proximal tubules; maternal NOAEL = 80 mglkg/day, HDT Hamster
90
270
Cleft palate, tail and skeletal
Teramoto et aI., 1978
anomalies, decreased fetal weight; maternal NOAEL = 810 mglkglday, HDT Hamsters
100
NA
NOAEL
Chernoff et aI., 1979
Mice
100
200
Increased supernumerary ribs;
Chernoff et aI., 1979
maternal NOAEL < 100 mglkg/day based on increased relative liver weight Guinea pigs
100
NA
NOAEL
Chernoff et aI., 1979
Cat
120
NA
Fetal NOAEL; reported maternal
Khera and Iverson, 1978
toxicity inconsistent with any other reported studies of ETU
insufficiency, caused for example by nutritional iodine deficiency, the primary human response is goiter rather than neoplasia (Hill et aI., 1989, 1998). Despite extensive epidemiological studies, no convincing evidence has yet emerged to link iodine deficiency with human thyroid cancer (Hard, 1998). Thus, not only is a threshold model appropriate for hazard assessment of the effects of ETU, but also a large uncertainty factor is not needed to insure adequate protection of the human populations.
81.3.3.2 Liver Effects
Although the level of ETU exposure resulting from bioconversion of the EBDCs at maximum tolerated doses is generally insufficient to produce tumors of the liver, ETU given directly does produce tumors of the liver in mice (Chhabra et aI., 1992; Innes et aI., 1969). Metabolism of the EBDCs to ETU is less extensive in mice (Cameron et aI., 1990; Piccirillo et aI., 1992), and induction of liver tumors with ETU has been observed only
1768
CHAPTER 81
Dialkyldithiocarbamates (EBDCs)
in mice and only at higher dietary levels that were also associated with centrilobular hepatocellular cytomegaly and increased functional demand (i.e., work-related stress to the liver), in addition to thyroid inhibition and sequelae. Hepatocellular tumors have not been seen in rats or hamsters. Although thyroid effects were present, no increase in the incidence of hepatocellular tumors was noted after two years of dietary feeding at 17-18 mg/kg bw/day (Chhabra et ai., 1992). The precise mechanism of liver tumor formation with ETU in mice has not been fully elucidated. One hypothesis is that the liver tumors are related to stress on the HPT axis. The liver is subject to metabolic regulation by thyroid and pituitary hormones, and liver neoplasms have also been produced in mice by two other thionamides, 2-thiouracil and 6-methyluracil, which also produce thyroid neoplasms in rats and mice (IARC, 1974). A second threshold hypothesis notes that ETU exhibits the hallmarks of phenobarbital-type liver promotion, including mixed function oxidase induction, sustained hepatomegaly, eosinophilic foci, and cellular proliferation (McClain, 1995; Whysner et al., 1996). Yet a third notes that redox cycling of ETU, with consumption of glutathione, has been observed with rat liver microsomes in vitro (Decker and Doerge, 1991), and that oxidative damage due to glutathione depletion is a commonly recognized threshold mechanism of tumor formation in animals. A number of additional factors also lead to the inference that these liver tumors are nonrelevant to human risk. First, there have been no reports associating the use of the structurally related propylthiouracil or other clinically administered thionamide drugs with an excess incidence of primary liver cancer in humans. Second, there is a wide variability in the incidence of liver tumors among various strains of mice, which is partly dependent on hormonal and/or nutritional factors, in addition to genetic factors. Genetic factors are particularly operative in the case of B6C3FI and other C3H-derived strains whose high and variable background tumor incidences indicate the presence of a significant population of "initiated" or latent tumor cells whose potential is readily expressed under stressed conditions of various origins. It has been acknowledged by numerous authorities that the induction of these tumors is of questionable or no relevance for assessment of oncogenic potential in human populations, where the background incidence of liver cancer is extremely low (e.g., IPCS, 1990; EU Directive 93121IEEC; IARC, 1987). In conclusion, the available information on the pathobiology and mechanism of the ETU-induced liver tumors indicates they are nonrelevant for the assessment of human risk at doses below the threshold for conventional toxic responses in these organs. 81.3.3.3 Genotoxicity
Few materials have been tested for mutagenic potential as exhaustively as the EBDCs and ETU. Among the more than 200 reported studies are more than sufficient numbers of qualified assays to ensure that each of the EBDCs and ETU individually
and collectively have been adequately tested in a wide variety of in vitro and in vivo mutagenicity tests. Care must be taken in the qualification step to exclude flawed or otherwise deficient results. For example, because of the rapid degradation of EBDCs in dimethylsulfoxide (DMSO), and the accompanying rapid liberation of metal ions, EBDC studies in which DMSO has been used as a solvent are invalid. A weight of the evidence evaluation of the scientifically valid studies shows that, when properly tested in higher organism test systems, the EBDCs and ETU are not mutagenic in the two major endpoints used to assess genotoxicity, gene mutations and chromosomal damage, and further, they do not cause adverse effects in ancillary tests of genotoxic damage. Thus, the weight of the evidence indicates that the EBDCs and ETU are not mutagenic in mammalian systems (EBDCIETU Task Force, 1992; Elia et ai., 1995). This conclusion is shared by other organizations, for example, by the NTP (1992) which stated that "ethylenethiourea has been tested extensively for genotoxicity in a variety of in vitro and in vivo test systems, and the results, with few exceptions, are negative," and the WHO (1994) which concluded that ETU was not genotoxic. 81.3.3.4 Bioequivalent Doses of the EBDCs
and ETU The bioconversion factors in rats and mice provide a quantitative basis for comparing the dose-responses of the EBDCs and ETU, and these comparisons provide quantitative support for the concept that the EBDCs' toxic effects are directly related to their conversion to ETU. Example calculations for mancozeb, maneb, and metiram using the 7.5% bioconversion factor in rats and the 6.0% factor in mice are shown in Table 81.4. It is clear that the ETU-equivalent dose levels arising from the feeding of the EBDCs are similar enough to the corresponding LOAELs and NOAELs for ETU to justify the presumption of a cause and effect relationship. 81.3.4 REPRODUCTIVE AND
DEVELOPMENTAL TOXICITY Reproductive outcome is generally unaffected by exposure to EBDCs or ETU. There were no effects on reproductive parameters, the microscopic appearance of the reproductive organs, or neonatal survival or growth resulting from exposure to any of the EBDCs or ETU at levels below those producing frank systemic toxicity in the adults (Tables 81.2 and 81.3). Further, unlike most pesticides, the potential for long-term effects of in utero exposure has been investigated. With the exception of a slight increase in thyroid tumor incidence in rats, in utero and perinatal exposure to ETU did not alter the incidences of tumors produced by postweaning lifetime exposures in either rats or mice (Chhabra et ai., 1992). Developmental toxicity is observed as malformations and embryofetotoxic effects at maternally toxic dose levels with all three EBDCs in the rat (Table 81.2). The effects seen are qualitatively consistent with those produced by ETU, and
81.3 Hazard Identification
1769
Table 81.4 Bioequivalent Doses of the EBDCs and ETU Critical doses (mglkg/day) ETU-equivalent Critical effect/active
EBDC
toEBDca
ETU
Rats Thyroid tumor LOAEL mancozeb
29
2.2
maneb
90
6.8
metiram
NA
3.7
NA
Thyroid effect LOAEL mancozeb
15
1.1
maneb
22.4
1.7
metiram
20
1.5
1.1
NOAEL mancozeb
4.8
0.36
maneb
20
1.5
metiram
12
0.9
0.37
Mice Liver, thyroid tumor LOAEL mancozeb
NA
maneb
440
metiram
NA
NA b
56c
26 NA
Thyroid LOAEL mancozeb maneb metiram
130
7.8
44
2.6
302
17
18
NOAEL mancozeb
17
1.0
maneb
11
0.66
metiram
84
5.0
1.7
a EBDC mg/kg bw/day X ETU bioconversion factor of 7.5% in rats, 6% in mice.
bLiver adenomas only. cLiver, thyroid, and pituitary tumors. NA = not applicable.
the dose-response is consistent with their causation by bioconverted systemic ETU doses. Sensitivity varies with the species. Exposure to sufficient doses of ETU at the critical stages of pregnancy produces malformations in rats, predominantly those of the central nervous system and head. Related malformations are produced in hamsters although only at very high doses approaching maternally toxic levels. The mouse and rabbit are far less sensitive. Developmental effects, even at relatively high doses, are limited to findings indicative of embryofetotoxicity. The guinea pig and cat have not shown evidence of teratogenic or other developmental effects (Table 81.3; see also Ruddick and Khera, 1975; Khera, 1987). A relationship of the developmental effects to thyroid inhibition is indicated by several lines of evidence. The malformations produced by ETU exposure in vivo are those expected as the result of thyroid insufficiency. They occur
only at doses in excess of those producing significant thyroid inhibition in adults, and they have been prevented, at least in part, by coadministration of thyroxine (Emmerling, 1978a, b). A key concern with thyroid inhibitors is that impaired thyroid function may alter hormone-mediated events during development, leading to permanent alterations in brain morphology and function (Cooper and Kavlock, 1997). The potential for this type of effect with ETU has been examined in a postnatal behavioral study in rats, resulting in a NOAEL of 25 mg/kg/day, fivefold higher than the 5 mg/kg/day NOAEL for other types of developmental toxicity demonstrated in a companion study (Chernoff et aI., 1979). Consistent with the findings with ETU, no developmental effects were observed with the EBDCs in studies in rabbits. Feed refusal, weight loss, and late-developing abortions observed in
1770
CHAPTER 81
Dialkyldithiocarbamates (EBDCs)
these studies are not relevant to human health risk. Being ruminants, rabbits are very sensitive to disruption of their gut microflora by antimicrobials and fungicides (ICH, 1994), and this response is well known to precipitate late term abortions (A. Hoberman, private communication). 81.3.5 NEUROTOXICITY
The one feature of the EBDCs' toxicological profile which is not explainable by the relationship to ETU is hindlimb paralysis and associated effects, including muscular atrophy with mancozeb and metiram and an effect on the retina with mancozeb. Hindlimb paralysis occurs at high doses with all of the EBDCs (LOAELs 60-340 mg/kg bw/day) and is a property of their common primary metabolite, ethylenebisisothiocyanatesulfide (EBIS) (Freudenthal et aI., 1977b), presumably related to the ability to release the carbon disulfide moiety (e.g., Johnson et aI., 1998). Multiple exposures are generally required to produce the effect. Acute neurotoxicity testing of maneb produced no indication of an adverse effect at doses up to 2000 mg/kg (Nemec, 1993). No evidence of neurotoxicity has been observed with ETU, and a firm chronic NOAEL for neurotoxicity of 8.2 mg/kg/day has been demonstrated in perfusion neuropathology studies of mancozeb (Stadler, 1991). 81.3.6 METABOLITES OTHER THAN ETU
In studies of metabolites other than ETU, EBIS further metabolized in rats and mice to ETU and ETU metabolites (Iverson et al., 1977; Jordan and Neal, 1979), and the thyroid functional changes which would be expected as a result of this metabolism were observed in addition to neurotoxicity (Freudenthal et aI., 1977b). No evidence of tumorigenicity was observed with EU (Innes et al., 1969), and neither EU nor any of the other mancozeb and ETU metabolites tested exhibited any teratogenic potential (Ruddick et aI., 1976b). 81.3.7 HAZARD CHARACTERIZATION
In summary, based on extensive data, the EBDCs do not pose a hazard of acute intoxication, genetic damage, or reproductive or developmental toxicity below levels that produce other kinds of toxicity in adults, or of significant systemic toxicity by the dermal route. There is no evidence of bioaccumulation. Repeated exposures to high doses of the EBDCs affect the thyroid, liver, and nervous systems in laboratory animals. The thyroid and liver effects are due to their metabolism in small amounts to ETU, which interferes with the synthesis of thyroid hormone and induces stress-related liver growth. These effects are reversible when exposures are brief or intermittent, but prolonged exposures can produce secondary changes, including anemia and thyroid, pituitary, and liver tumors in rodents. Available mechanistic information establishes a threshold for the thyroid and pituitary tumors and indicates that none of the tumor types
are relevant for human risk assessment at likely exposure levels. Thus, neither the EBDCs nor ETU pose an oncogenic risk for humans. The EBDCs' neurotoxic effects are shared with their common primary metabolite EBIS, although not with ETU, and a reliable NOAEL well above anticipated human exposure levels has been confirmed in specific neuropathology studies of mancozeb.
81.4 DOSE-RESPONSE 81.4.1 NOAEL AND ACCEPTABLE DAILY INTAKE-ETU
The acceptable daily intake (ADI) or reference dose is defined as the approximate exposure which, if incurred daily over an entire lifetime, appears to be without appreciable risk for the general popUlation, including all subgroups; i.e., it is the exposure level which provides a reasonable certainty of no harm and is therefore safe within the meaning of the U.S. Food Quality Protection Act and the WHO. When clinical studies in humans are inappropriate or unavailable, the ADI is estimated from reliably conducted toxicity studies in laboratory animals by taking the NOAEL associated with the most sensitive endpoint in the most sensitive species, and applying uncertainty factors to account for inter and intraspecies variability, and the need to protect the most sensitive individuals and subgroups, such as infants and children, among the general population. For ETU, inspection of the critical findings in the comprehensive laboratory animal studies summarized in Table 81.3 readily reveals that thyroid and related parameters are the most sensitive effects. The rat has generally been the most sensitive species to ETU followed closely by the dog and monkey, with the mouse relatively insensitive. An overall NOAEL of 0.4 mglkg/day for the effects of ETU on the HPT axis, and therefore for the effects of ETU in laboratory animals, is supported by the 0.39 mg/kg/day NOAEL for hypothyroid changes including anemia in the 90 day and one year study of ETU in dogs (Briffaux, 1991, 1992), the 0.37 mg/kg/day NOAEL for thyroid, pituitary, and liver effects in the two year chronic study in rats (Schmid et aI., 1992), and the 0.11-0.43 mg/kg/day NOAEL for thyroid effects among the parents in the rat two-generation reproduction study (Dotti, 1992). Further, it is consistent with the weight of evidence of a comprehensive database, including the 1.1 mg/kg/day NOAEL for induction of thyroid tumors in the two year carcinogen bioassay in rats. The 1.1 mg/kg/day level was a LOAEL for thyroid hormonal changes and follicular hyperplasia (Chhabra et aI., 1992). Taking into account a standard lO-fold safety or uncertainty factor for interspecies variability and a second 10-fold factor for intraspecies variation leads to the ADI of 0.004 mg/kg bw/day for ETU, as recommended by FAOIWHO (1994). Given the known lesser sensitivity of humans to thyroid effects, this is a very conservative assessment, affording extra protection of
81.4 Dose-Response the public health. A number of factors, among them the unusually complete and reliable database, including chronic effects of in utero exposure, the lesser sensitivity of humans to thyroid effects, and the fact that the NOAELs for developmental and reproductive effects are 1O-fold or more higher than those for thyroid effects, confirms that no additional uncertainty factor is needed to ensure adequate protection of infants and children.
81.4.2 NOAEL AND ACCEPTABLE DAILY INTAKE-EBDCS
The key requirements to support the identification of an aggregate NOAEL for a group of compounds are a common toxicological profile and a similar dose-response. As noted above, the EBDCs share a comparable toxicological profile, primarily based on the toxic effects of their common metabolite ETU. An issue which deserves some comment in this context is muscular atrophy, which was the critical effect in the two year study of metiram in rats. Muscular atrophy has been seen at high doses with both mancozeb and metiram and is associated in both of these materials with clinically diagnosed hindlimb paralysis at the same or higher doses. Thus, this effect is not an exception to the common toxicological profile in the qualitative 1.0 mg/kg bw/day
sense, although its appearance at the LOAEL for metiram was not typical of the dose-response seen with the other EBDCs. With this one possible exception, the three EBDCs exhibit similar dose-effect and dose-response curves, as is illustrated in Fig. 81.3, which presents the NOAELs and LOAELs for the critical studies from Table 81.2 in a graphical format. Inspection of the graph readily reveals that the dose-responses of the three EBDCs are consistent. With only minor exceptions, the NOAELs and LOAELs of the respective materials do not overlap for any particular type of study. Second, no one of the EBDCs is consistently more sensitive than the others. Of the 10 studies conducted with at least two of the three EBDCs, the LOAELs for each study type are evenly divided among the three (4 mancozeb, 3 maneb, 3 metiram), indicating that the apparent differences among the compounds in individual studies are due to dose selection, and not to any intrinsic differences in potency. Similarly, the critical LOAELs for all three compounds are nearly identical, at between 10 and 16 mg/kg bw/day. Since the toxicological profiles and dose-response are similar for the three EBDCs, it is appropriate to base the determination of the overall NOAEL for the EBDCs on the aggregated dataset for the group. An overall NOAEL of 5.0 mg/kg/day for the EBDCs as an aggregate group is supported by all of the available studies, as presented in Table 81.2 and Fig. 81.2. It
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1772
CHAPTER 81
Dialkyldithiocarbamates (EBDCs)
is further supported by consideration of the dose response and NOAELs for ETU. Taking into account the 7.5% bioconversi on factor (EPA, 1992; Kocia1ski, 1989), a 5.0 mg/kg/day dose of EBDC would result in a systemic dose of 0.4 mg/kg/day of ETU (5.0 mg/kg/day x 0.075 = 0.375 mg/kg/day), precisely equal to the overall NOAEL for ETU determined in independent studies. This comparison lends the support of the full ETU database to the 5.0 mg/kg/day aggregate NOEL for the EBDCs. In summary, an aggregate evaluation of the individual toxicology databases for the respective EBDCs, mancozeb, maneb, and metiram, and their ETU metabolite indicates an overall NOAEL of 5.0 mg/kg bw/day for the group. Application of a standard lOO-fold overall safety factor leads to a recommended acceptable daily intake of 0.05 mg/kg/day. The factors discussed above for ETU, and the unusual reliability of the database, reflecting the combined results of four individually comprehensive databases, confirm that no additional uncertainty factors need be applied. This aggregate assessment produces a comparable, if slightly higher, estimate of the ADI than the FAOIWHO recommendation. The WHO panel reviewed the data for each of the actives individually and established ADIs of 0.05 mg/kg/day for mancozeb and maneb, and 0.03 mg/kg/day for metiram, resulting in the allocation of a group ADI of 0.03 mg/kg/day for the EBDCs collectively, including mancozeb, maneb, metiram, and zineb. The basis for the establishment of a group ADI was the similarity of the chemical structures of the EBDCs, the comparable toxicological profile of the EBDCs based on the toxic effects of ETU, and the fact that the parent EBDC residues cannot be differentiated using presently available regulatory analytical procedures (FAOIWHO, 1994). 81.4.3 ACUTE REFERENCE DOSE
The acute reference dose (aRID) is the maximum single day oral exposure which is anticipated to be without appreciable risk for the general population. Toxic effects which might occur as a result of exposures occurring within the period of a single day, or after at most a very few doses, are relevant to the assessment. For ETU the only relevant acute toxicological end point is developmental toxicity. A NOAEL of 5.0 mg/kg/day is supported by the weight of the evidence of multiple studies in the rat, as the most sensitive species (Chernoff et aI., 1979; Khera, 1973; Saillenfait et aI., 1991; Teramoto et aI., 1978). The mechanistic relationship to thyroid inhibition suggests that mUltiple exposures, producing hormone depletion, would be required for full expression of ETU's developmental toxicity potential in the lower dose range. This and the unusually thorough nature of the database argue that the standard lOO-fold uncertainty factor is more than adequate to assure protection of women, infants, and children, indicating an aRID for ETU of 0.05 mg/kg/day. For the EBDCs, since expression of their neurotoxic potential requires repeated doses (e.g., Nemec, 1993) and their relevant developmental effects are due to bioconversion to
ETU, the most relevant endpoint for assessment of acute risks is the aRID for ETU of 0.05 mg/kg bw/day, applied to the combined direct and indirect (7.5% bioconverted EBDC dose) ETU exposure. 81.4.4 ENDPOINTS FOR ASSESSMENT OF DERMAL AND RESPIRATORY EXPOSURE
For assessment of the potential risks of pesticide users and bystanders, and those who encounter exposure after application, the dermal and respiratory routes of exposure are most relevant. Dermatitis due to repeated exposures and irritation of the mucus membranes are prevented by appropriate personal protective equipment. Dermal absorption of the EBDCs is low, 6.5% or less, and consistently with this, the EBDCs are of very low toxicity by the dermal route, with NOAELs in 21-day dermal toxicity studies ranging from 100 mg/kg/day for maneb to 1000 mg/kg/day (limit dose) for mancozeb. Alternatively, the aRID and ADIs derived from the oral exposure studies may be applied to the estimated systemic exposures from single or multiple doses, respectively, after adjustment for dermal absorption. Subchronic inhalation studies produced the same kinds of effects as in oral studies of the same length and a generally similar dose-response, with NOAELs of 79, 10, and 2 mg/m 3 for mancozeb, maneb, and metiram, respectively. After correction for respiration rates and respirable fraction, the estimated systemic exposure of 8.3 mg/kg/day at the NOAEL in the mancozeb study was equal to the 7.4-8.2 mg/kg/d NOAELs in the 90-day dietary studies, indicating the relevance of the aRID and ADIs for systemic exposures. In enclosed spaces, a Workplace Environmental Exposure Limit Guide of 1 mg/m 3 is recommended (AIHA, 1992). 81.4.5 CARCINOGENICITY CLASSIFICATION AND LOW DOSE RISK ASSESSMENT
The criteria for classification of substances as carcinogens differ with the classifying authority. Mancozeb and ETU were evaluated by the EU Commission in 1994, in the context of European classification criteria which assign an important role to the mode of action, and were not classified as carcinogens. In contrast, the International Agency for Research on Cancer (IARC) classified ETU as category 2B "possibly carcinogenic to humans" based on sufficient evidence in laboratory animals but inadequate evidence in humans (IARC, 1987), and the EPA (1992) cited evidence of carcinogenicity in two species and a weak genotoxic potential in classifying ETU, as a B2 "probable human carcinogen." The Agency's classification of the EBDCs as similar B2 carcinogens was largely due to the metabolic conversion to ETU. Because EPA prefers a probabilistic approach to risk assessment for carcinogens in this class, the linear multi stage model was used to calculate an upper
81.6 Risk Characterization 95% confidence limit of 0.06 (mg/kg bw/day)-l on the lifetime risk of cancer from ETU at low exposure levels (q*) (EPA, 1997).
81.5 TOXICOLOGY IN HUMANS With the exception of sporadic reports of allergic contact hypersensitivity (Bruze and Fregert, 1983; Crippa et aI., 1990; Kleibl and Rackova, 1980; Lee et al., 1981), studies of manufacturing workers and users have not discerned adverse effects of exposurc to eithcr ETU or the EBDCs. In the most comprehensive study of thyroid effects, 153 men currently or previously exposed to EBDC for many years at a manufacturing site were compared for thyroid function to 153 men not exposed to EBDC, its products, or ETU who also worked at the same plant. Workers and controls were carefully matched with respect to age, race, length of employment, and type of employment. Informed consent was obtained from all participants. No significant differences were observed on thyroid palpitation and in thyroid function tests between the EBDC workers and the control men. Urinary excretion of ETU in a subgroup of 42 workers currently exposed to EBDC was 0.002 ppm compared to 0.001 ppm in the control group (41 men), most of whom had undetectable values. The authors concluded that exposure to EBDC manufacture was not associated with an increased prevalence of thyroid abnormalities (Charkes et aI., 1985). Clinical examinations and thyroid function tests were also carried out over a period of 3 years in the United Kingdom on 8 male workers engaged for between 5 and 20 years in the manufacture of ETU and 5 male workers involved for 3 years in the mixing of ETU with rubber, and matched controls. Levels of ETU recorded on personal samplers of manufacturing workers reached 330 J.tg/m3, and levels for mixers ranged from 120 to 160 J.tg/m 3. Mixers but not process workers had significantly lower levels of T4 in their blood compared to controls. With the exception of one mixer with elevated TSH levels, who was evaluated as hypothyroid on further testing, no effects were found on TSH or thyroid binding globulin. The authors concluded that there was no evidence that thyroid function is severely affected by exposure to ETU at the levels experienced by these workers (Smith, 1984). Similarly, no hazard of clinical thyroid depression existed based on medical evidence collected on 51 workers exposed to ETU at a U.S. rubber-processing company (Salisbury and Lybarger, 1977). No difference was found in the total death rate or deaths due to cancers between 992 male workers involved in the production of EBDCs from 1948 to 1975 and control males from the city of Philadelphia. No thyroid cancer was found in this study (DeFonso, 1976). Epidemiological studies were conducted on workers in the rubber industry by Parkes (1974) and Smith (1976). Based on examination of national and regional thyroid cancer incidences in the UK, Parkes concluded that, under the conditions in which ETU had been used in the rubber industry, there is no risk of
1773
man contracting thyroid cancer as a result of industrial exposure to ethylenethiourea (Parkes, 1974). Similarly, a total of 1929 workers engaged in the production or manufacture of ETU were surveyed retrospectively for thyroid cancer and were compared with the thyroid cancer list of the Birmingham (England) Cancer Registry from 1957 to 1971. No thyroid cancers occurred in these workers. A retrospective study of 699 women who were employed at a rubber manufacturer using ETU assessed the incidence of fetal abnormalities occurring in children to women who had worked with ETU during early pregnancy. No excess incidence was found, and the study did not demonstrate any risk oftcratogcncsis (Smith, 1976).
81.6 RISK CHARACTERIZATION 81.6.1 DIETARY EXPOSURE AND RISKS
Risk is a function of hazard potential and exposure. The potential risk of consumers of foods derived from chemically protected crops is assessed by comparing the ADI for the particular crop protection chemical, or other indices of its toxicity potential when relevant, to estimated or measured dietary exposure values. Dietary exposure is in turn determined from the level of residues of the crop protection chemical and its toxicologically significant metabolites in foods and food consumption patterns for various population groups. Dietary intake may be predicted with varying degrees of accuracy. The WHO and the European Union use the International Estimated Dietary Intake (lED!) for five global diets. The exposure is calculated using average regional food consumption data and the median residue from supervised trials that have been used to assess the maximum residue limit or tolerance consistent with the worldwide labels. Residues of EBDCs and ETU in processed commodities were also calculated. The IEDI was calculated for each of the five global diets using the EBDC and ETU level with the highest dietary contribution. On this basis, the aggregate dietary exposure to the EBDCs ranged from a low of 3% of the ADI in the African diet to 36% of the ADI in the European diet. These levels are well within the criteria of WHO for fully acceptable dietary risk (less than 100% of the AD!) (Ollinger, 1999). These data overestimate the dietary risks of the EBDCs because the IEDI calculation assumes that 100% of the crop is treated and the residues are measured from crops as they are harvested. More reliable assessments of dietary exposure may be obtained at the national level (National Estimate of Daily Intake, NEDI) . Such a more accurate determination of EBDC and ETU exposure is available from a market basket survey conducted in the United States. In that study, samples of fresh and processed foods were collected every two weeks for one year from grocery stores throughout the United States in a statistically designed survey. Using the most sensitive residue analytical methods achievable, the results from the analysis of almost 6000 samples showed that about 80% of the samples had no measurable residues of EBDC or ETU. The residues that
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CHAPTER 81
Dialkyldithiocarbamates (EBDCs)
were found were very close to the limit of quantitation in the method. Thus, there is virtually no exposure to the consumer to residues of either EBDCs or ETU. A reliable NEDI was derived for the United States, a member of the European diet category, from the average consumer level residue values determined in the market basket and related studies, and detailed national data on the consumption of raw and processed food commodities as summarized in the US EPA's Dietary Risk Estimation System. The NEDI for U.S. consumer exposure to EBDCs and ETU was 0.000027 mg/kg bw/day, calculated as ETU, or less than 1% (0.68%) of the 0.004 mg/kg bw/day ADI for ETU. These NEDI values clearly show that actual consumer exposure to residues of the EBDCs is negligible (EBDCIETU Task Force, 1997b). Utilization of these market basket data in a Monte Carlo assessment of acute dietary exposure resulted in an acute margin of exposure for women of child-bearing age (13 + years) of 2652 at the 99.9 percentile of exposure, greatly beneath any levels of concern (EBDCIETU Task Force, 1997b). EPA calculated the risks to consumers using the market basket data and their standard method of estimating the probability of human cancer risks. They determined that there was only a negligible theoretical dietary risk of 1.2 in one million from the use of EBDCs on 48 crops (EPA, 1992, 1996). The negligible character of dietary exposure to the EBDCs is further underlined by the nutritional benefits which they confer. The principal use of the EBDCs is the economical protection of fruit and vegetables from disease, making these commodities more widely available in the diet. Overwhelming evidence from epidemiological studies indicates that diets high in fruits and vegetables are associated with a lower risk of numerous cancers, and for this reason, dietary recommendations to increase the intake of citrus fruits, cruciferous vegetables, green and yellow vegetables, and fruits and vegetables high in vitamins A and C have been made by numerous organizations (AICR, 1997; American Cancer Society, 1984, 1996; Block et at., 1992; Giovannucci, 1999; NASINRC, 1989; NCI, 1987; Steinmetz and Potter, 1991). 81.6.2 WORKER EXPOSURE AND RISKS
Exposure estimates for mixing, loading, and applying EBDC fungicides during outdoor agricultural applications have been calculated using the predictive operator exposure model of the UK MAFF and also the model of the German BBA as sources of surrogate exposure data. Each of these models represents a synthesis and integration of exposure data obtained in a large number of field trials conducted using various kinds of equipment in a variety of national agricultural settings. The final models were calibrated to insure that the resulting estimates would, if anything, overpredict actual exposures under field conditions. For mode ling of EBDC uses, an application rate of 2.4 kg ailhectare was chosen to represent a typical maximum use rate of EBDC products when used on a standalone basis. Use rates for these products when used as mixtures
are lower. As gloves are recommended as personal protective equipment when handling undiluted product, the use of gloves (during mixing and loading only) is presumed in the models. The exposure estimates obtained using these very conservative models indicate that operator exposures are always below, and in most cases substantially below, acceptable dermal and inhalation exposure limits for workers, even when personal protective equipment is limited to the bare minimum of gloves during mixing and loading. Therefore a significant risk for the operator during the use of EBDC fungicide products appears unlikely. As in the case of dietary exposure, alternative risk assessment approaches have been used at times by various authorities. At the conclusion of the U.S. Special Review, the EPA concluded there was adequate safety to mixers, loaders, and applicators when proper personal protective equipment is used (EPA, 1992). Estimates of exposure for other activities, including bystanders, workers reentering treated fields, homeowners, and others reentering treated areas are even lower. 81.6.3 CONCLUSION
Because of their importance to worldwide agriculture, the EBDCs and ETU have been thoroughly tested over many years. Collectively, the data demonstrate that the use of the EBDCs results in essentially negligible exposure to consumers, coupled with a significant contribution to improved nutrition, and low risk to farm workers, production workers, and people who are exposed through recreational activities.
REFERENCES Allen. T. R., Frei, Th., Biedermann, K., Luetkemeier, H., Terrier, Ch., Vogel, 0., and Wilson, J. (1989). "l3-Week Oral Toxicity (Feeding) Study with Maneb Technical in the Dog." Rep. 206605 from Research and Consulting Company, Ltd., Itingen, Switzerland. Unpublished report of Elf Atochem North America, Inc. American Cancer Society (1984). Nutrition and cancer: causation and prevention. An American Cancer Society special report. CA Cancer J. Clin. 34,
5-10. American Cancer Society (1996). Guidelines on diet, nutrition, and cancer prevention: Reducing the risk of cancer with healthy food choices and physical activity. The American Cancer Society 1996 advisory committee on diet, nutrition and cancer prevention. CA Cancer J. Clin. 46, 325-341. American Industrial Hygiene Association (AIHA) (1992). "Workplace Environmental Exposure Level Guide: Mancozeb." AIHA, Akron, OH. American Institute for Cancer Research (AICR) (1997). "Food, Nutrition, and the Prevention of Cancer: A Global Perspective." World Cancer Research Fund and the American Institute for Cancer Research, Washington, DC. Amold, D. L., Krewski, D. R., Junkins, D. B., McGuire, P. E, Moodie, C. A., and Munro, I. C. (1983). Reversibility of ethylenethiourea-induced thyroid lesions. Toxico!. App!. Pharmacol. 67, 264-273. Atterwill, C. P., and Aylard, S. P. (1995). Endocrine toxicology of the thyroid for industrial compounds. In "Toxicology of Industrial Compounds" (H. Thomas, R. Hess, and Waechter, eds.), pp. 257-280. Taylor & Francis, London.
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CHAPTER 81
Dialkyldithiocarbamates (EBDCs)
National Cancer Institute (NCI) (1987). "Diet, Nutrition, and Cancer Prevention: A Guide to Food Choices." United States Government Printing Office, Washington, DC. National Toxicology Program (NTP) (1992). "Toxicology and Carcinogenesis Studies of Ethylene Thiourea in F344IN Rats and B6C3Fl Mice (Feed Studies)." United States Department of Health and Human Services, Public Health Service, National Institutes of Health Publication 92-2843, United States National Toxicology Program Technical Rep. Ser. 388. Nelson, S. S. (1986). "Metabolism of l4C Mancozeb in Rats." Unpublished Rep. 3lH-86-02 of Rohm and Haas Company. Nelson, S. S. (1987). "Bioconversion of Mancozeb to ETU in Rat." Unpublished Rep. 3IC-87-24 of Rohm and Haas Company. Nemec, M. D. (1992). "A Developmental Toxicity Study of Maneb Technical in Rats." Rep. WIL-134011 from WIL Research Laboratories, Inc., Ashland, OH. Unpublished report of Elf Atochem North America, Inc. Nemec, M. D. (1993). "An Acute Neurotoxicity Study of Maneb Technical in Rats." Rep. WIL-134015 from WIL Research Laboratories, Ashland, OH. Unpublished report of Elf Atochem North America, Inc. OdeIl, W. D., et al. (1967). Studies of thyrotropin physiology by means of radioassays. In "Proceedings of the 1986 Laurentian Hormone Conference" (G. Pincus, ed.), Vol. 23, pp. 47-85. Academic Press, San Diego. 0' Hara, G. P., and DiDonato, L. J. (1985). "Dithane M-45 and Ethylenethiourea: 3-Month Dietary Toxicity Study in Mice." Unpublished Rep. 80R-124 of Rohm and Haas Company. OIIinger, J. (1999). "Dithiocarbamates (CCPR Code 105). EBDC/ETU, STMR-P for Apple Juice, EBDC lED! Calculation." Letter to W. H. van Eck, Codex Committee on Pesticide Residues, EBDC/ETU Task Force. O'Neil, W., and Marshall, W. (1984). Goitrogenic effects of ethylenethiourea on rat thyroid. Pesticide Biochem. Physiol. 21, 92-10l. Palmer, A., and Simons, R. (1979). "Effect of Metiram Technical on Pregnancy of the Rat." Rep. RZ 79/065, BSF 302179616 from Huntingdon Research Centre, Huntingdon England. Unpublished report of BASF Corporation. Parkes, H. G. (1974). Living with carcinogens. 1. Inst. Rubber Industry 8, 2123. Peters, A. C., Kurtz, P. J., Donorrio, D. J., Thake, D. C., and Cottrill, D. L. (1980a). "Prechronic Studies of Ethylenethiourea: Acute, Repeated Dose and Subchronic in Rats." Unpublished Project G-7l86 of Battelle Laboratories, Columbus, OH, for United States National Institute of Environmental Health Sciences. Peters, A. C., Kurtz, P. J., Donorrio, D. J., Thake, D. C, and Cottrill, D. L. (l980b). "Perchronic Studies of Ethylene Thiourea: Acute, Repeated Dose and Subchronic in Mice." Unpublished Project G-7l86 of Battelle Laboratories, Columbus, OH, for United States National Institute of Environmental Health Sciences. Peters, A. C., Kurtz, P. J., Chin, A. E., CarIton, B. D., Chrisp, C E., and Dill, G. S. (1982). "Report on the Maximum Neonatal Dose Studies with Ethylenethiourea." Unpublished Contract NOl-ES-8-2l5l of Battelle Laboratories, Columbus, OH, for United States National Institute of Environmental Health Sciences. Piccirillo, V. J., Wu, D., and Speirs, G. (1992). "Metabolism of Mancozeb in the Mouse." Rep. T9l-34l3 from Inveresk Research International, Ltd. Tranent, Scotland. Unpublished report of Elf Atochem North America, Inc. Puhl, J. R. (1985). "Metabolism of Radiolabelled Maneb in Rats." Rep. 6181101 from Hazleton Laboratories, America, Inc., Madison, WI. Unpublished report of Elf Atochem North America, Inc. Rowland, J. (1997). "Toxicity Endpoint Selection Process." Health Effects Division, United States Environmental Protection Agency Office of Pesticide Programs. Ruddick, J. A., and Khera, K. S. (1975). Pattern of anomalies following single oral doses of ethylenethiourea to pregnant rats. Teratology 12, 277-281. Ruddick, J. A., WiIIiams, D. T., Hierlihy, L., and Khera, K. S. (1976a). 14Cethylenenthiourea: Distribution, excretion, and metabolism in pregnant rats. Teratology 13, 35-40. Ruddick, J. A., Newsome, W. H., and Nash, L. (1976b). Correlation ofteratogenicity and molecular structure: ethylenethiourea and related compounds. Teratology 13, 263-266.
Ruddick, J. A., Newsome, W. H., and Iverson, E (1977). A comparison of the distribution, metabolism and excretion of ethylenethiourea in the pregnant mouse and rat. Teratology 16, 159-162. Ryle, P. R., Bell, P. E, Parker, C, Farmer, H., Offer, J. M., Anderson, A.. and Dawe, I. S. (1991). "A Study of the Effect of Maneb (Technical) on Reproductive Function of Two Generations in the Rat." Rep. MNB 119072 from Huntington Research Centre Ltd., Cambridgeshire, England. Unpublished report of Elf Atochem North America, Inc. Saillenfait, A. M., Sabate, J. P., Langonne, I., and De Ceaurriz, J. (1991). Difference in the developmental toxicity of ethylenethiourea and three N,N/-substituted thiourea derivatives in rats. Fundamental Appl. Toxicol. 17,399-408. Salisbury, S. A., and Lybarger, J. (1977). "Ethylene Thiourea." Health Hazard Evaluation Determination Rep. 77-67-499 from SI. Clair Rubber Company, Marysville, MI, United States Department of Health, Education and Welfare. Schaller, R. T., and Stevenson, J. K. (1966). Development of carcinoma of the thyroid in iodine-deficient mice. Cancer 19, 1063-1080. Schmid, H., Tennekes, H., Janiak, T., Probst, D., Luetkemeier, H., Pappritz, G., Mfu'ki, U., Vogel, 0., and Heusner, W. (1992). "Ethylene Thiourea 104 Week Chronic Toxicity (Feeding) Study in Rats." Project 256803 from Research and Consulting Company, Ltd., Itingen, Switzerland. Unpublished report of ETU Task Force. Shaw, D. (1990). "Mancozeb: 52-Week Oral (Dietary) Study in the Beagle." Project HLA 5913-616/3 from Hazleton Laboratories, UK. Unpublished report of Rohm and Haas Company. Shellenberger, T. (1991). "Mancozeb: 18-Month Oncogenicity Study in Mice." Project 85051 from Tegeris Labs, Inc., Temple Hills, MD. Unpublished report of Rohm and Haas Company. Smith, D. (1976). Ethylene thiourea-A study of possible teratogenicity and thyroid carcinogenicity. 1. Soc. Occupational Med. 26, 92-94. Smith, D. (1984). Ethylene thiourea: Thyroid function in two groups of exposed workers. British 1. Industrial Med. 41, 362-366. Solomon, H., and Holz, J. (1987). "Mancozeb: Range-Finding Developmental Study in Rats." Unpublished Rep. 85R-244 from Rohm and Haas Company. Solomon, H. M., and Lutz, M. E (1987). "Mancozeb: Oral (Gavage) Developmental Toxicity Study in Rabbits." Unpublished Rep. 86R-021 from Rohm and Haas Company. Solomon, H. M., Lutz, M. E, and Kulwich, B. A. (1988). "Mancozeb: 2-Generation Reproduction Study in Rats." Unpublished Rep. 87R-020 from Rohm and Haas Company. Sortwell, R., Heywood, R., AlIen, D., Prentice, D., and Cherry, C. (1977). "Metiram Preliminary Oral Toxicity Study in Rhesus Monkeys. Repeated Dosage for 4 Weeks." Rep. BASF 77/0149, BSF 265177264 from Huntingdon Research Centre, Huntingdon, UK. Unpublished report of BASF Corporation. Sortwell, R., Alien, D., Heywood, R., and Street, A. (1979). "Metiram (Containing 2.2% Ethylenethiourea) Oral Toxicity Study in Rhesus Monkeys (Repeated Dosage for 26 Weeks with Recovery Period)." Rep. BASF 79/0082, BSF 267178263 from Huntingdon Research Centre, Huntingdon, UK. Unpublished report of BASF Corporation. Stadler, 1. (1990). "Combined Chronic Toxicity/Oncogenicity 2-Year Feeding Study with Mancozeb in Rats." Unpublished Rep. 259-89 of DuPont HaskelI Laboratory. Stadler, J. (1991). "Neuropathology Study in Rats with Mancozeb." Unpublished Rep. 217-89 of DuPont HaskelI Laboratory. Steinmetz, K. A., and Potter, J. D. (1991). Vegetables, fruit, and cancer. I. Epidemiology. Cancer Causes Control 2, 325-357. Takayama, S., Aihara, K., Onodera, T., and Akimoto, T. (1986). Antithyroid effects of propylthiouracil and sulfamonomethoxine in rats and monkeys. Toxicol. Appl. Pharmacol. 82, 191-199. Teramoto, S., Shingu, A., Kaneda, M., and Saito, R. (1978). Teratogenicity studies with ethylenethiourea in rats, mice and hamsters. Congenital Anomalies 18, 11-17. Terrill, J. B. (1990). "Acute Inhalation Toxicity Study with Maneb in the Rat." Rep. HLA 2567-100 from Hazleton Laboratory, Vienna, VA. Unpublished report of Elf Atochem North America, Inc.
References
Terrill, J. B. (1991). "A Four-Week Study in the Rat to Determine the Lung Tissue Residue and Effect upon the Thyroid Function with Maneb." Rep. HLA 2567-101 from Hazleton Laboratory, Vienna, VA. Unpublished report of Elf Atochem North America, Inc. Tesh,1. M., McAnulty, P. A., WilIoughby, C. R, Enticott, J., Wilby, O. K., and Tesh, S. A. (1988). "Mancozeb: Teratology Study in the Rat." Rep. 87IPTC 007/365 from Life Science Research Ltd., Suffolk, UK. Unpublished report of Elf Atochem North America, Inc. Teshima, R, Nagarnatsu, K., Kido, Y., and Terao, T. (1981). Absorption, distribution, excretion and metabolism of ethylenethiourea in guinea pigs. Eisei Kagaku 27, 85-90. Thomas, G. A., and Williarns, E. D. (1991). Evidence for and possible mechanisms of non-genotoxic carcinogenesis in the rodent thyroid. Mutation Res. 248, 357-370. Tomlinson, H. L., and Longacre, S. L. (1988). "Mancozeb Dermal Absorption Study in Male Rats." Unpublished Rep. 88R-218 of Rohm and Haas Company. Tompkins, C. E. (1992). "An IS-Month Dietary Oncogenicity Study in Mice with Maneb Technical." Rep. WIL-134008 from WIL Research Laboratories, Inc., Ashland, OH. Unpublished report of Elf Atochem North America, Inc. Trutter, J. A. (l988a). "Subchronic Toxicity Study with Maneb Technical in the Rat." Rep. HLA 153-140 from Hazleton Laboratories, Vienna, VA. Unpublished report of Elf Atochem North America, Inc. Trutter, J. A. (I 988b). "A 21-Day Dermal Toxicity Study in Rabbits with Maneb Technical." Rep. HLA 153-139 from Hazleton Laboratories, Vienna, VA. Unpublished report of Elf Atochem North America, Inc. Trutter, J. A. (l988c). "Mancozeb: 4-Week Repeat Dermal Toxicity Study in Rats." Rep. HLA 417-432 from Hazleton Laboratories, Vienna, VA. Unpublished report of Rohm and Haas Company.
1779
Ulland, B. M., Weisburger, 1. H., Weisburger, E. K., Rice, J. M., and Cypher, R (1972). Thyroid cancer in rats from ethylenethiourea intake. 1. Nat. Cancer Inst. 49, 583-584. Ullman, L., Sacher, R, Porricello, T., Luetkemeier, H., Vogel, W., Vogel, 0., Wilson, J., and Terrier, H. (1987). "Report on the Subacute 21-Day Repeated Dose Dermal Toxicity Study with Polyrarn DF in Rabbits. Rep. BASF 87/0260, ZNT 86/314 from Research and Consulting Company AG, Ltd., Itingen, Switzerland. Unpublished report of BASF Corporation. Ulrich, C. E. (1986a). "Thirteen-Week Subchronic Inhalation Toxicity Study on Maneb in Rats (Final Report)." Rep. 550-001 from International Research and Development Corporation, Mattawan, Ml. Unpublished report of Elf Atochem North America, Inc. Ulrich, C. E. (l986b). "Report on the Thirteen Week Subchronic Inhalation Toxicity Study on Metiram in Rats." Rep. RZ 86/407 and Addendum BASF 87/0414 from International Research and Development Corporation, Mattawan, Ml. Unpublished report of BASF Company. Ulrich, C. E. (1987). ''Thirteen-Week Subchronic Inhalation Toxicity Study on Maneb in Rats (Addendum to Final Report Covering Recovery Phase)." Rep. 550-001 from International Research and Development Corporation, Mattawan, Ml. Unpublished report of Elf Atochem North America, Inc. Watts, M. H., and Chan, P. K. (l984a). "Dithane M-45: Acute Oral Toxicity Study in Rats." Unpublished Rep. 83R-218 from Rohm and Haas Company. Watts, M. H., and Chan, P. K. (I 984b). "Dithane M-45: Acute Oral Toxicity Study in Rats and Mice." Unpublished Rep. 83R-213A and B from Rohm and Haas Company. Whysner, J., Ross, P. M., and Williarns, G. M. (1996). Phenobarbital mechanistic data and risk assessment: enzyme induction, enhanced cellular proliferation, and tumor promotion. 1. Pharmacol. Therapeutics 71, 153191. World Health Organization (1994). "The WHO Recommended Classification of Pesticides by Hazard and Guidelines to Classification, 1994-1995." International Program on Chemical Safety, WHOIPCS/94.2.
CHAPTER
82 A Toxicological Assessment of Sulfur as a Pesticide* Derek w. Gammon, Thomas B. Moore, and Michael A. O'Malley Department of Pesticide Regulation, California EPA
82.1 INTRODUCTION 82.1.1 USAGE Elemental sulfur is the most heavily used crop protection chemical in California (Table 82.1) as well as in the United States. In 1993-1995, for example, annual usage was about 70 million pounds active ingredient (a.i.) in California, which is about onethird of the total weight of pesticides used in agriculture. It is generally applied to crops as a dust to combat fungal disease, at rates of approximately 10 to 30 lbs. per application per acre, as well as being used for postharvest disease control. The range of fungal diseases controlled by sulfur includes brown rot, scab, mildew, powdery mildew, leafspot, and rusts (Farm Chemicals Handbook, 1998). It is also used, to a lesser extent, for the control of mites and insects (fleahoppers) which may be secondary to fungal damage of the plant. Multiple applications are often needed for crops which are particularly susceptible to fungal attack. The main crops on which sulfur is used in California (1995) are grapes (71 %), tomatoes (12%), and sugarbeet (8%). It has become an important component of IPM systems since it can be used in "organic" farming. 82.1.2 ENVIRONMENTAL FATE As a natural substance, environmental fate requirements for sulfur in the United States have been waived by U.S. EPA. In a variety of literature reports, sulfur has been shown to be oxidized, in the presence of water and soil, to the sulfite and then the sulfate, i.e., sulfuric acid. The supplementation of fertilizers with sulfur has been used intentionally to acidify soils which are too alkaline for a particular crop. Conversely, the use of sul*The opinions expressed in this chapter represent the views of the authors and do not necessarily reflect the views and policies of the Department of Pesticide Regulation. Handbook of Pesticide Toxicology Volume 2. Agents
fur as a fungicide can make the soil too acidic for the continued optimal growth of a particular crop. For example, in Southern Tanzania, sulfur dust was used to control powdery mildew on cashew nut trees. After 4 years, the topsoil pH was reduced by 0.7 units, to below pH 5.5, the ideal pH for cashew nut tree yield (Majule et al., 1997). However, in the case of a 3-year field study on highbush blueberry bushes in mineral soil, sulfur amendment increased both early growth and blueberry yield. It was concluded that the effects of sulfur were probably mediated by a decrease in soil pH with corresponding increases in Mn and Fe levels (Haynes and Swift, 1986). It was also found, in a lysimeter study, that increasing the sulfur content of soil led to a rise in sulfur content of plants, such as corn, wheat, barley, sunflower, and mustard (Gador and Motowicka-Terelak, 1986). Sulfur is, of course, a natural constituent of plants, as it is of all organisms. In addition to elemental sulfur which may be present from pesticidal exposure, sulfur is commonly found as sulfates and in the amino acids cysteine, methionine, and glutathione. Agricultural practice may introduce elevated sulfur levels into arable land inadvertently by, for example, the use of (animal-derived) manure or other fertilizers as well as from the use of pesticides. By far the greatest contribution in the latter case is likely to be from elemental sulfur used as a fungicide. As described below, there are an increasing number of instances of dermatitis in farm workers and, in many ways more serious, an elevated number of cases of disease in ruminants caused by exposure to high levels of sulfur, and several case reports are given. Another source of soil acidity caused by sulfur is from industrial pollution. Typically from the burning of coal or oil, sulfur dioxide or hydrogen sulfide can be liberated into the atmosphere. This gaseous sulfur can return to earth following rainfall, producing the so-called "acid rain" phenomenon since sulfuric acid can readily be produced from soil oxidation of sulfur.
1781
Copyright © 200 1 by Academic Press. All rights of reproduction in any fonn reserved
1782
CHAPTER 82 Sulfur
Table 82.1 Usage of Elemental Sulfur (a.i.) in California, 1993-1995 a.b
CROP
1995
I 994c
1993
in thousands
in thousands
in thousands
acresd
Ibs.
acres d
Ibs.
acresd
Ibs.
Alfalfa
294
13.2
227
10.7
Almonds
184
42.4
126
26
Bermudagrass
162
5.9
Cantaloupe
634
31.3
357
17.9
363
20.9
Carrots
177
10.8
347
15.2
454
17.9
Cotton
460
17.3
240
10.9
319
13.3
Date
741
11.3
655
10.9
838
14.1
31.2
1.26
6.4
139 48.4
11.0
62.6
2.2
Grapes, table
26,200
2970
24,900
2660
25,600
2630
Grapes, wine
23,200
2460
23,500
2340
25,700
2450
Lemons
193
4.8
190
5.8
168
5.3
Melons
133
5.8
282
11.8
349
17.0
Nectarines
277
35.3
240
31.0
198
25.5
Peaches
873
83.3
644
69.8
588
66.1
Pears
170
13.6
169
12.9
197
16.2
Peas
204
10.1
305
14.5
281
14.6
Pistachio
500
39.4
592
30.7
846
33.9
Plums
259
25.0
262
23.7
267
26.9
Prunes
224
23.0
135
13.8
157
17.4
64.7
262
61.9
256
63.9
Strawberry
238
Sugarbeet
5300
Tomatoes, fresh
1270
Tomatoes, processing
7250
TOTAL (lbs.)
171 62.8 292 69.8 million
210
6760 1620
76.9
7810
290
218
6870
70.7
1380
219
7700
70.5 million
73.5 million
aDPR, 1995, 1996a, 1996b. b All crops which received more than 100,000 Ibs. a.i. in 1995 are included. c 132,000 Ibs. of sulfur were applied to 8600 acres of oranges in 1994. d Acres treated include multiple applications to the same land.
82.2 TOXICOLOGY PROFILE OF ELEMENTAL SULFUR Acute toxicity categories for products containing elemental sulfur were assigned according to FIFRA Pesticide Assessment Guidelines, Subdivision F, Hazard Evaluation: Human and Domestic Animals, Revised Ed., Nov., 1984. The results are summarized in Table 82.2.
82.2.2 ACUTE EXPOSURE DERMAL TOXICITY: 81-2 A single-dose limit test was conducted using the rabbit (five/ sex), dosed on the skin at 2000 or 5000 mg/kg. Lower doses are only required if there is > 20% mortality per sex. There were no compound-related, acute mortalities for any of these formulations at the tested doses of 2000 or 5000 mg/kg. This suggests that it may be more appropriate to consider them all in Category IV rather than Ill.
82.2.1 ACUTE EXPOSURE ORAL TOXICITY: 81-1 A single-dose limit test was conducted using the rat (five/sex), dosed by gavage at 5000 mg/kg. Lower doses are only required if there is > 20% mortality per sex. Seventeen of 20 formulations had a LD50 > 5000 mg/kg (Category IV); 3 were between 500 and 5000 mg/kg (Category III).
82.2.3 ACUTE EXPOSURE INHALATION TOXICITY: 81-3 A single-dose limit test was conducted using the rat (five/sex), dosed by inhalation at 2 mg/l for 4 hours. Lower doses are only required if there is > 20% mortality per sex. Fourteen of 17 for-
82.2 Toxicology Profile of Elemental Sulfur
1783
Table 82.2 Acute Toxicity of Sulfur Fonnulations used in California
Sulfur product Special Electric Refined Super Adhesive Dusting (98% A.I.)
Oral"
Dennalb
IV
III
InhaI.C
Eye
Dennal
Dennal
irrit. d
irrit.'
sensit. f
III
IV IV
Manufacturing Use (98%)
IV
III
IV
III
Spray Sulfur (98%)
IV
III
IV
III
IV
Valor Brands products dusting Sulfur (98%)
IV
III
IV
III
IV
0
90% Sulfur WP (90%)
IV
III
IV
III
IV
0
Bensul 85 (85%)
IV
III
IV
III
IV
0
Clean Crop Apple & peach Kolofonn Fungicide (84%)
IV
III
IV
III
IV
0
LX 112-2 (53%)
IV
III
IV
III
IV
Sulfur 6L (51 %)
IV
III
IV
III
IV
Happy Jack Sarcoptic Mange Medicine (28%)
IV
III
III
III
Safer Garden Fungicide Concentrate (12%)
IV
III
IV
III
IV
Fonnula 242 (0.4%)
IV
IV
IV
XF-97097 (96.75%), with myclobutanil, 0.5%
IV
IV
IV
III
IV
BT 320 Sulfur 50 (50%), with BT,0.064%
IV
III
IV
III
IV
Britz BT50 & Sulfur Dust (50%)
IV
III
IV
III
IV
Cook/Sevin Plus Multi-purpose Garden Dust (30%) with carbaryl 5%, PBO 0.45%, pennethrin 0.03%
III
III
III
IV
IV
Britz Botran 6-25 Dust (25%) with dichloran, 6%
IV
III
IV
III
IV
Britz Copper Sulfur Dust (25%) 15-25 with Cu, 15%
IV
III
IV
II
IV
CopperlSulfur Flowable (15.5%) with Cu Sulfate, 27.5%
III
III
III
II
IV
Kocide 404S (15%) with Cu hydroxide, 26%
III
III
III
0
0
0
BT,0.064%
IV
aCategory IV: LDso > 5000 mg/kg; Category Ill: LDso = 500-5000 mglkg. bCategory IV: LDso > 5000 mg/kg; Category Ill: LDso = 2000-5000 mg/kg. CCategory IV: LCSO > 2 mg/l; Category Ill: LCSO > 0.5-2 mg/l. d Category IV: minimal effects, clearing in <24 hours; Category Ill: corneal involvement or irritation, clearing in ~7 days; Category II: corneal involvement or irritation, clearing in 8-21 days; Category I: corrosive (irreversible ocular damage) or corneal involvement or irritation, clearing in >21 days. eCategory IV: mild or slight irritation (no irritation or slight erythema); Category Ill: moderate irritation at 72 hours (moderate erythema). fBuehler test: score of 0 (no erythema) to 3 (severe erythema, with or without edema).
mulations had a LCso > 2 mg/l (Category V); 3 were between 0.5 and 2 mg/l (Category Ill). 82.2.4 PRIMARY EYE IRRITATION: 81-4
A single-dose limit test was conducted using the rabbit (six of either sex), dosed in one eye at 0.1 mllanimal (or 100 mg for a solid). The other eye served as the untreated con-
trol and responses were graded (Table 82.2). Two of 20 formulations showed minimal effects, clearing within 24 hours (Category IV); 15 showed corneal involvement or irritation, clearing in ~7 days (Category Ill); 2 showed corneal involvement or irritation, clearing in 8-21 days (Category 11); one showed corrosive (irreversible ocular damage) or corneal involvement or irritation, clearing in >21 days (Category I), probably as a result of the copper hydroxide in this formulation.
1784
CHAPTER 82 Sulfur
82.2.5 PRIMARY DERMAL IRRITATION: 81-5 A single-dose limit test was conducted using the rabbit (six of either sex), dosed at 0.5 mVinch 2 (or 0.5 g/inch2) for 4 h. The responses were graded. Nineteen of 20 formulations caused mild or slight irritation (Category IV); 1 showed moderate irritation at 72 h (Category Ill).
82.2.6 PRIMARY DERMAL SENSITIZATION: 81-6 Using the Buehler test, induction doses were applied to clipped skin at 0.4 mL or 500 mg/guinea pig (ca. 1 to 2 g/kg) three times, on a weekly basis, followed two weeks later by a challenge dose, to a naive site. Dermal sensitization was measured, as erythema with or without edema, in response to the challenge dose, at 24 and 48 h, on a scale of 0 to 3. All of the (seven) formulations tested scored zero, i.e., were negative.
82.3 TOXICOLOGY OF SULFUR DIOXIDE Sulfur dioxide (S02) is used as a fumigant because of its antimicrobial properties. It is a colorless gas with a high water solubility. In solution, it hydrates to sulfurous acid (H2S03) which dissociates in turn to form bisulfite (HS0 3) and sulfite (SO~-) ions. The bisulfite ion is quite reactive by means of ionic and free radical mechanisms (Shapiro, 1977). Sulfur dioxide is used in California for the treatment of grapes held in cold storage to control the fungus Botrytis cinerea. The recommended treatment rate is up to a 1% gas concentration for up to 20 treatments with 7 to 10 day intervals between treatments depending on the variety of grape. The main crop uses of S02 in California are summarized in Table 82.3. Sulfur dioxide is used in the United States as a food additive under the authority of the Food and Drug Administration in beer, wine, fruits and vegetables, fruit juices, syrups, meats, and fish. It acts as a preservative in these foods by being both an antimicrobial and an inhibitor of the enzymes which contribute to the discoloration process. Sulfur dioxide has been used in wine
Table 82.3 Usage of Sulfur Dioxide (a.i.) in California. 1993-1995 a CROP
1995 (Ibs.)
1994 (Ibs.)
1993 (Ibs.) 194,000
Grapes, table
144,000
267,000
Grapes, wine
24,000
1100
12,000
Commodity
14,000
11,700
48,100
4000
5500
fumigation Other fumigation Structural pests TOTAL (Ibs.)
200,000
aDPR, 1995, 1996a, 1996b.
6000 15,100
13,400 285,000
276,000
making to selectively inhibit the growth of acetic acid and lactic acid producing bacteria. Product registrations for sulfur dioxide in California are for the 100% compressed gas. Precautionary labeling for these products require the signal word "Danger" with the wording "inhalation may be fatal or cause serious illness. Prolonged or repeated exposure may cause impaired lung function .... Liquid or excessive vapor exposure can cause serious skin and eye injury. Harmful if swallowed." When sulfur dioxide is used as a fumigant, respiratory protection is required unless the ambient concentration is less than 2 ppm, which is the threshold limit value for occupational exposure. Case studies of cats and dogs fed fresh pet food preserved with sulfur dioxide resulted in examples of animals suffering from thiamine deficiency (Studdert and Labuc, 1991). These animals demonstrated a syndrome of depression, pupillary dilation, and ataxia which occasionally progressed to seizures and sudden death caused by acute cardiac failure. In the preserved food samples in which the S02 content was greater than 800 mg/kg, the thiamine levels were decidedly reduced. In the presence of sulfiting agents such as sulfur dioxide, thiamine is cleaved into its constituent pyrimidine and thiazole moieties, rendering it inactive. It should be noted that the principal toxic effect of elemental sulfur on the CNS of ruminants is a direct effect of sulfur and not a secondary effect arising from thiamine deficiency. Other investigators examined pigs fed barley with high moisture content which had been treated with sulfur dioxide (Gibson et al., 1987). Treatment of the barley (l % sulfur dioxide (wtlwt» demonstrated an enhanced preservation of the barley with a significant time delay before mold growth became evident. However, the thiamine content in the barley was greatly reduced, resulting in a thiamine content in the meat of the treated pigs which was 7.6% that of the control animals. These animals gave evidence of cardiac hypertrophy along with reduced feed intake and body weight gain. Once again, it is possible that direct effects of S02 contributed to the toxicity of the barley to the pigs, rather than these being purely secondary consequences of thiamine deficiency. The World Health Organization specifically recommends that foods which are significant sources of thiamine in the human diet should not be treated with sulfur dioxide or other sulfiting agents. Pollution of the environment has been a major health concern as a consequence of excessive exposures to sulfur dioxide and smoke, for example in the Meuse Valley of Belgium in 1936, in Donora, Pennsylvania in 1948, and in London in 1952. In London where 4000 deaths and numerous incidences of illness were attributed to the exposure, atmospheric sulfur dioxide levels achieved a daily average as high as 1.34 ppm. Pulmonary effects manifested by exposure to sulfur dioxide are attributable to its irritancy. Exposure to sulfur dioxide alone results in direct effects on the nasopharynges aDd trachea with reduced transport of the mucous layer either due to cessation of ciliary movement in an acute exposure or to an excessive thickening of the mucous as a consequence of chronic exposure. The acute pulmonary response is typical of a irritant effect with bronchial
82.4 Veterinary Effects of Sulfur
restriction resulting in increased flow resistance. The chronic effect is similar to that of chronic bronchitis without the involvement of a bacterial infection. These effects have been well reviewed by Costa and Amdur (1996).
82.4 VETERINARY EFFECTS OF SULFUR Probably the major health concern of sulfur for ruminants is the association between excessive sulfur ingestion and polioencephalomalacia (PEM), also known as cerebrocortical necrosis. This was first recognized as a disease in sheep and cattle over 40 years ago (Jensen et aI., 1956; Terlecki and Markson, 1961). It involves a softening of the gray matter of the brain and is a major disease worldwide (Olkowski, 1997). Clinical signs can occur from a few hours to several weeks after exposure to excessive sulfur. Signs usually include, initially, mild excitation and restlessness accompanied by loss of appetite. Affected animals avoid light and signs may progress to headpressing, rigidity, blindness, violent convulsions, coma, and death. Young animals are particularly badly affected. Removal of affected animals from the source of sulfur generally reduces the severity of clinical signs. In the past, the causes of PEM have been ascribed to a variety of agents, including a lack of vitamin Bl (thiamine) and, more recently, to an excess of dietary sulfur. It has been suggested that the increase in reported cases of PEM over recent years is a result of industrial pollution. However, with the movement away from coal to oil and natural gas, this seems unlikely. For example, according to Beauchamp et al. (1984), the ratio of H2 S content of coal, oil, and natural gas, per unit weight of fuel burned, is approximately 35:8: 1. Assuming that H2S liberation is a reasonable marker for possible industrial sulfur exposure, this suggests a reduction, rather than an increase, in environmental exposure to sulfur from industrial sources over the years. Several reports have described PEM arising from feeding sheep and/or cattle on elevated levels of sulfur in the diet (Hill and Ebbett, 1997; Jeffrey et aI., 1994; Low et aI., 1996), in drinking water (Hamlen et aI., 1993), as well as a case of sheep being allowed to forage on a field of alfalfa which had been treated with elemental sulfur (Bulgin et aI., 1996). Reports of field cases have been duplicated in laboratory studies implicating excessive sulfur ingestion as the cause of PE M (e.g., Gould et aI., 1991; McAllister et aI., 1992; Sager et aI., 1990). PEM was identified in several sheep and cattle farms in England (Jeffrey et aI., 1994). It was associated with the use of ammonium sulfate as a feed additive, in place of ammonium bicarbonate, as the usual urinary acidifier. After this was discontinued, there were no further cases of PEM. Necropsy of six calves and two lambs from five of these farms showed lesions in the thalamus and striatum, of great severity, unlike in cases of thiamine deficiency. No lesions were found in PEM-affected animals in the cerebellum, hippocampus, or superior colliculi. Symptoms were similar to those already described and there were deaths on three out of five farms.
1'/8:'
On a cattle farm in New Zealand, PEM was diagnosed following gross and histopathological examination of the brain of deceased animals (Hill and Ebbett, 1997). The cattle had been feeding for two months on hay plus a rationed amount of kale (Brassica oleacea) when they were transferred to a field of kale. Within two days, neurological signs were observed which were typical of PEM. Twenty-six of 99 heifers (26%) were symptomatic of which 12 died (12%) and 14 recovered (14%) after removal of the cattle from this field. Chemical analysis of the kale revealed that it contained 8500 mg/kg of suI fur (0.85%, DM, dry matter), which is double the maximum range of recommended dietary needs of cattle for sulfur (0.4%, DM). Examples of sulfur-induced PEM in sheep include an outbreak on a sheep farm in Scotland, after changing from grazing to a ration of pellets containing 0.43% DM sulfur (Low et aI., 1996). Clinical signs appeared 15 to 32 days after changing to the artificial diet, and the incidences were 16 of 46 (35%) for Swaledale lambs and 5 of 25 (20%) for Scottish blackface lambs. Clinical signs, which were quite unlike those of vitamin B 1 deficiency, included depression, blindness, head-pressing, nystagmus, and dorsiflexion of the neck or opisthotonus. In some animals, the severity was such that the sheep either died or were killed in extremis (4/16 Swaledale; 4/5 Scottish blackface). Histopathological examination of the lambs revealed evidence of PEM in the majority of the animals. The mean intake of pellets during the study period was 880 g/head/day (Swaledale) and 760 g/head/day (Scottish blackface). Because the lambs had an initial average body weight of 20 kg, this intake converts to a food intake of approximately 40 g/kg/day and a sulfur ingestion of approximately 170 mg/kg/day. Administration of vitamin Bl (by injection) did not reverse the clinical signs, but there were no new cases evident after vitamin Bl was given combined with removal of the lambs from the high sulfur diet. An outbreak of PEM was also reported in cattle which had been drinking water containing a high concentration of sodium sulfate, 7200 ppm vs a recommended optimal level of 1000 ppm, in Canada (Hamlen et aI., 1993). The incidence was 111110 (10%) and mortality of affected cattle was 4111 (36%). Clinical signs, which first appeared three days after exposure to the well, and histopathology (n = 3), were the same as those reported above, with the additional findings of extensive thrombosis and vascular necrosis in midbrain and thalamus. The blood clinical chemistry for affected animals appeared to be normal from the standpoint of copper and vitamin Bl levels and transketolase activity (a thiamine-dependent enzyme). The PEM dissipated and no new cases arose after the cattle were moved to a water supply with acceptable levels of sulfur. Unusually, old rather than young animals were affected, but this could have resulted from low exposure to sulfur in calves which were nursing. The level of magnesium in the affected well was also high, 1050 vs 200 ppm, recommended. Another example of sulfur toxicity to livestock is a report (Bulgin et aI., 1996) of a flock of sheep grazing on a field of alfalfa stubble which had been sprayed 14 to 16 hours previously with an aqueous suspension of 35% elemental sulfur at
1786
CHAPTER 82
Sulfur
53 lbs./acre, active ingredient.! Within two to four hours the sheep became uncoordinated with 91 % prostration, despite the sheep being moved to uncontaminated pasture after 2 h, when the problem became apparent. There was 10% (220 out of 2200) mortality after a week, the majority of these sheep (206) dying of acute effects, within 24 hours. Surviving ewes were considered fully recovered at 90 days. Necropsy of sheep which died between 2 and 48 h after the onset of clinical signs revealed a rumen pH of 6.0 to 6.5, a strong smell of rotten eggs (H2S), but no digestive tract lesions. Necropsy of sheep dying between 5 and 30 days showed PEM, consisting of yellow/tan areas of the cerebral cortex caused by neuronal degeneration and cavitation of cortical grey matter. It should be noted that alfalfa normally is moderately rich in sulfur, having a content of ca. 0.4% DM (Olkowski, 1997), without added extraneous sulfur. Attempts have been made to study the toxicity of sulfur in laboratory experiments. The appearance of clinical signs of PEM in calves fed on a high sulfate diet coincided with or immediately followed the first odor of H2S in rumen gas (Sager et aI., 1990). PEM was not correlated with copper or thiamine deficiency. These findings were extended by Gould et al. (1991), who fed a high sodium sulfate diet to calves and noted that H2S accumulation in the stomach was significantly higher in animals with signs of PEM than in asymptomatic calves. Microbes in the rumen readily reduce sulfate to sulfide. The findings were extended to sheep by McAllister et al. (1992). Ten lambs were dosed with sodium hydrogen sulfide (0.94 M) every 20 min, administered directly into the esophagus. Clinical signs of PEM developed within 45 min of first dosing, in all lambs, and PEM was identified in 4/9 brains examined histologic ally. All (4) animals had visual impairment including blindness, dying at 20 to 96 h. Two lambs had visual impairment without PEM but both died within 90 min of pulmonary congestion and edema (as seen with acute H2S toxicity in the rat), probably before brain lesions had time to develop.
82.5 HUMAN HEALTH EFFECTS OFSULFUR There are many cases of dermatitis associated with the agricultural use of elemental sulfur. For example, in California between 1974 and 1985, there were 677 cases reported, more than for any other pesticide (Table 82.4). This would seem to suggest that sulfur is a potent skin irritant in humans. Isolated cases have also been reported in applicators and field workers in Washington State. However, standard (epicutaneous) skin irritation tests in laboratory animals for most agricultural formulations have not shown irritation (Matsushita et aI., 1977; Table 82.2). In nonstandard tests though, using subcutaneous injection in the Wistar rat, a 25% aqueous solution of wettable powder or a 22% solution of lime sulfur caused a four I The restricted entry interval for field workers following sulfur use is 24 hours, i.e., appropriate personal protective equipment must be worn to enter a treated field within 24 h of a sulfur application.
Table 82.4 Summary of Cases of Possible, Probable and Definite Illness Reported to the California Pesticide Illness Registry Involving Exposure to Sulfur as a Primary Pesticide Between 1982 and 1995 Type of illness Eye Activity Clean/fixing
Eye
+
other
Skin Skin
2
2
10
2
other
+
Respiratory/ systemic
Total 3
pesticide equipment Handling
2
!4
concentrate Drift Emergency
50
37
0
7
58
10
8
11
7
87
148
4
5
9
55
15
34
155
6
6
\0
41
response Flagger Handlers Manufacturing/
2
0
formulating Other Packing/
19 0
2
processing Field residue Total
81
52
425
51
61
603
221
115
503
85
201
978
or five irritant reaction. 2 Using a similar maximization test with the guinea pig, a 1% or 5% aqueous solution of elemental or lime sulfur was a moderately strong allergen (Matsushita et al., 1977). Limited case reports also implicate elemental sulfur as a human contact allergen. Schneider (1978) reported two cases of contact allergy in patients who used medications containing elemental sulfur to treat superficial fungal dermatoses. Both patients had positive patch test reactions to 5% elemental sulphur in various vehicles. A control series was not reported. Wilkinson (1975) reported the case of a professional gardener with a previous history of atopic eczema who developed an eczematous eruption involving the elbow flexures and the right hand. He had a positive patch test reaction to 5% sulfur in petrolatum, but a control series was not reported. Gregorczyk and Swieboda (1968) described 15 cases of desquamative dermatitis among 425 Polish sulfur miners in which irritant dermatitis due to elemental sulfur may have played a part. Several instances of apparent allergic reaction to 1% elemental sulfur were also observed in a recent study of California nursery workers (O'Malley and Rodriguez, 1998). Allergic contact dermatitis was identified in a hospital investigation of patients suffering from eczematous dermatitis (Vena et al., 1994). Patients were subjected to patch tests with (sodium or potassium) metabisulfite (S20;-), bisulfite 2 I = no reaction; 2 = slight hyperemia; 3 emia and edema; 5 = necrosis.
= hyperemia; 4 = marked hyper-
82.5 Human Health Effects of Sulfur
(HS0 3), or sulfite (SO~-). Fifty cases of allergic reaction out of 2894 patients of either sex (1.7%) were reported after exposure to metabisulfite, with 100% cross-reactivity between the sodium and potassium salts and with the bisulfite. Only 2 (4%) of these gave a positive reaction to sulfite. Because metabisulfite is readily converted to bisulfite under aqueous conditions, it is not surprising that they showed cross-reactivity. However, because of the low cross-reactivity toward sulfite, it appears unlikely that sulfite is the ultimate allergen, in vivo, although it is readily formed from the metabisulfite or bisulfite under acidic conditions. It remains possible that some cases of dermatitis resulting from elemental sulfur are due to the subsequent conversion to one of these derivatives. A case history of human Sulfur Spring dermatitis from dermal exposure was reported from Taiwan (Sun and Sue, 1995). Over a 10-year period, 44 cases of dermatitis were recorded in visitors to a particular hot springs resort. Two springs were considered, a green sulfur spring (GSS) and a white sulfur one (WSS), and it transpired that all the dermatitis cases had visited the GSS. Of these 44 cases, 32 (70%) had visited the GSS only once, for 10 to 20 min; 25 (57%) had also visited WSS, without signs of dermatitis; and 24 (55%) had a history of skin diseases prior to visiting the GSS. The chemical and physical properties of the GSS and WSS springs were compared with tap water and microbial infection was ruled out as a cause, since no cultures could be grown from water or affected skin. The principal causes of the dermatitis were considered, by the authors, to be soluble sulfur, which was present at 600, 100, and 80 ppm, in the three water sources, respectively; acid irritation, since pH was 1-2,4, and 7, respectively, was considered a probable contributory factor. It was also noted that there was a large variation in chloride levels, 3000, 20, and 20 ppm. Other factors which could have contributed to the dermatitis in this hot spring were high temperature (100, 50, and 20°C, respectively) and ammonia nitrogen (200, 0.2, and 0.0, respectively). These disparate pieces of information suggest that active irritants (e.g., sulfuric acid or hydrogen sulfide) or allergens (e.g., sulfites or hydrogen sulfide) may be produced by oxidation or reduction of sulfur. Thus, sulfur may be the precursor of dermal irritants and allergens rather than being one per se.
82.5.1 OCCUPATIONAL EXPOSURE For the years between 1982 and 1995, the California illness registry contained 1698 reports of definite, probable, and possible illnesses involving sulfur, including 978 cases (58%) for which sulfur was identified as the primary cause of the reported illness. To evaluate the typical effects of direct exposure to sulfur, the 155 cases involving handlers (mixers, loaders, and applicators cases) are discussed in more detail below. These cases constitute 16% of the cases for which sulfur was identified as the primary cause of the reported illness. Reactions to elemental sui fur can be broken down by illness type.
1787
82.5.1.1 Eye Ocular symptoms were present in 68 (44%) of the handler cases, including 58 cases of isolated ocular symptoms, 8 cases of eye and skin symptoms, and 2 cases of respiratory and eye symptoms. The nature of the ocular reaction to sulfur is illustrated by the cases below: Eye complaints following contact with suI fur 87-310
04/0111987
Sutter
Peach orchard sprayer had exposure to sulfur in eyes and was diagnosed as having chemical keratitis (superficial corneal injury).
87-508
05/0111987
Fresno
While dusting with sulfur in a vineyard, the material got into a worker's eye and it became irritated. He was wearing safety glasses but not goggles.
88-944
06/0111988
Kern
While loading sulfur for helicopter application, this employee experienced eye irritation.
As indicated by the above cases, the relationship between the sulfur exposure and the subsequent ocular irritation is usually simple to evaluate, because the irritation corresponds directly to the site of contact.
82.5.1.2 Skin Dermatitis was present in 70 (45%) of the handler cases; in 55 it was the only illness reported and in 15 the dermatitis occurred in conjunction with other illness symptoms. As illustrated on next page, some cases appear to have been related to chemical irritation (85-655), while others conceivably could have been related to sulfur allergy (87-174). However, provocation tests to confirm the suspected allergy were not documented in any of the cases reported to the California registry. Figures 82.1 and 82.2 show examples of dermatitis of the arm and torso of an exposed worker. These skin lesions appeared within a few minutes of spending 45 min dosing a rose bed with a mixture of elemental sulfur and malathion; they are typical of sui fur. The worker was wearing a short-sleeved shirt without gloves and was sweating profusely in the 95°F heat. This individual is unlikely to be allergic to sulfur since he has applied elemental sulfur on many occasions since this incident (wearing appropriate protective clothing) without experiencing any ill effects.
82.5.1.3 Respiratory Tract and Systemic Illness Classification of respiratory symptoms as systemic or purely topical is sometimes difficult. Thirty-four (22%) of the handler cases involved either respiratory or systemic illness. Complaints ranged from principally respiratory (84-726) to nonspecific systemic symptoms such as "vomiting" or "feeling shaky" (82-1436). Some cases had rhinitis (cases 83-595 and 87 -1 042) or asthma symptoms (84-726, 86-916) that suggested a possible immediate allergy to sulfur (e.g., 85-1134), but no cases had
1788
CHAPTER 82 Sulfur
Dermatitis complaints following exposure to elemental sulfur 85-655
05/14/1985
Sonoma
82-1211
06/29/1982
Madera
Developed skin rash after applying sulfur with wet coveralls.
86-2145
08/29/1986
Riverside
Man was dusting dates with sulfur, wearing a short-sleeve shirt. He experienced contact dermatitis on his upper arms, neck, and back.
87-174
02/28/1987
Tulare
A worker complained of a rash after mixing, loading, and applying Kolospray (81 % sulfur powder). He had a 2 year history of sensitivity to the material and reported that the rash occurred despite wearing complete safety gear. The treating physician suspected that the dermatitis was due to an allergic reaction and recommended avoiding sulfur powder in the future.
88-650
04/11/1988
Kern
An employee had been applying flowable sulfur on 04-11-88 and sulfur dust on 04-02-88. He developed a rash on his arms and neck, which he feels was due to the sulfur dust application.
Employee had applied sulfur with a hand held duster, after which he developed contact dermatitis or a chemical bum.
documented provocation tests to demonstrate that the reactions were allergic, rather than irritant, in nature. 82.5.1.4 Trauma or Illness Due to Combustion
of SuIfur The tendency of sulfur to oxidize makes it prone to combustion. Sulfur combustion was involved in several significant illness episodes reported to the California illness registry. These included both spontaneous combustion and combustion following aircraft accidents.
82.6 DISCUSSION Environmental exposure to sulfur arises from two main anthropogenic sources, industrial automobile emissions and from the use of sulfur in agriculture. This chapter automobile has concentrated on the latter. The extensive use of elemental sulfur as a fungicide in agriculture has, on occasion, led to veterinary
problems in animals ingesting toxic levels of sulfur. In addition, sulfur dioxide is used as a fumigant, generally as a preservative for food and drink. In ruminants, sui fur is converted by microorganisms in the rumen to hydrogen sulfide, which is readily absorbed. Sulfide can then inhibit a variety of enzymes involved in oxidative metabolism. It also inhibits respiration by blocking the carotid body and by combining with hemoglobin to produce sulfhemoglobin, thus reducing the oxygen-carrying capacity of the blood. High concentrations of sulfur may lead to secondary thiamine deficits. Sulfite ion is a strong nucleophile and readily binds to thiamine (to the positively charged nitrogen in the thiazole ring), leading to secondary deficits of vitamin B I. Veterinary problems associated with the ingestion of excessive sulfur include polioencephalomalacia (PEM), a severe brain disease of ruminants. A combination of anecdotal reports of field incidents and laboratory studies has clearly shown that dietary sulfur in these animals needs to be carefully regulated.
Respiratory and systemic symptoms following sulfur exposure Upper respiratory symptoms 83-595
03/29/1983
Merced
Diagnosed as allergic rhinitis; worker did not see doctor until 2 weeks after the reported exposure.
87-1042
06/08/1987
Kern
While dusting grapes with suI fur, he began having an apparent allergic reaction to the sulfur (watery eyes and sneezing).
Systemic and lower respiratory symptoms Glenn
A nursery worker was exposed to sulfur drifting from an application to an adjacent field. She began to vomit and "feel shaky."
04/25/1984
Kern
Experienced breathing difficulties after working with sulfur, triggered asthma attack.
06/03/1985
Tulare
Loading plane, experienced breathlessness and burning in eyes. Allergic reaction to sulfur.
86-916
06/05/1986
Stanislaus
Asthmatic response when exposed to sulfur dust. He reported wearing a face mask respirator, gloves, an apron, and hat while applying. M.D. advised him to not spray sulfur anymore.
89-1250
OS/25/1989
Napa
Worker was applying sulfur with a power sprayer attached to the rear of a tractor. He was wearing work clothes and Tyvek, boots, (prefitted) respirator with pesticide cartridge. Developed tightness in chest and cough. Diagnosis chemical bronchitis.
90-1214
OS/26/1990
Kern
Applicator dusting grapes experienced dizziness, shortness of breath, and fainting spell. Diagnosis exposure to sulfur dust. Off work for injury sustained when fell off tractor due to fainting spell. Wearing coveralls, goggles, and dust mask.
82-1436
07/13/1982
84-726 85-1134
82.6 Discussion
1789
Figure 82.1 Apparent irritant reaction after a sweaty forearm was contaminated with a mixture of suI fur and malathion. Reprinted with permission from M. A. O'Malley (1997), State of the Art Reviews in Occupational Medicin e 12, 327-345.
Elemental sulfur appears relatively inert in both the Buehler and Draize skin test models (see above) but is reported to show moderate capacity to cause sensitization in the guinea pig maximization test (Gregorczyk and Swieboda, 1968). This ambiguous response to sulfur in animal tests is contradicted by use experience, where dermatotoxicity in humans appears to be common. A possible explanation may lie in the transformation of sulfur through oxidation and reduction (Matsushita et aI., 1977) which may not readily occur in epicutaneous tests in rodents, which thermoregulate by increased respiration rather than by perspiration. The tendency of sulfur to spontaneously transform under field conditions is underscored by the cases of
combustion associated with its use, principally during the summer months in California, where daytime temperatures in the hot, dry inland valleys commonly exceed lOO°F (38°C). Current labeling and California regulations prohibit the aerial application of sui fur dust when ambient temperatures exceed 90°F (32°C), in order to reduce combustion incidents due to sulfur. Data from the California Illness Registry do not clearly indicate how many symptoms of sui fur exposure are due to irritant reactions, how many are possibly due to allergic mechanisms, and how many are due to unknown physiologic mechanisms. No cases were recorded to have provocation testing, the only means of clearly distinguishing between irritant and allergic
1790
CHAPTER 82
Sulfur
Figure 82.2 A 1+ reaction to sulfur - Subject 43 in a California nursery study. A total of 5 positive reactions to sulfur among 43 subjects. Reprinted with permission from M. A. O'Malley (1997), State of the Art Reviews in Occupational Medicine 12, 327-345.
Post-crash combustion 90-1220
06/0811990
Kern
A pilot applying sulfur to sugarbeets crashed when his plane suffered engine failure. The sulfur ignited on impact causing him to inhale the fumes as well as causing bums. He suffered chest tightness, breathing difficulties, and second degree bums over 20% of his body.
88-3029
09/30/1988
Solano
Pilot had just finished applying one load of sulfur to sugar beets when he crashed and died.
85-1461
06124/1985
San Joaquin
Pilot exposed to burning sulfur when his plane crashed.
Spontaneous combustion 94-818
06/ll1l994
Kern
Pilot was applying suI fur dust to sugar beets when spontaneous combustion of sulfur caused the cockpit to fill with smoke. While trying to land the plane, it flipped over. He crawled out of the cockpit and was taken to a hospital.
94-1278
0612511994
Sutter
While standing on the wing of a crop dusting plane, a worker was loading dusting sulfur into the plane's hopper when the sulfur dust ignited. The resulting explosion knocked him off the plane. He suffered respiratory system and skin bums.
93-1220
0712211993
Solano
Worker was loading sulfur dust from a hopper into an aircraft. The sulfur dust in the hopper shifted, causing sulfur dust to be dumped onto the airplane fuselage. The sulfur caught fire from the heat of the engine which burned the worker's exposed skin.
91-1252
07122/1991
Sonoma
An applicator inhaled smoke from a sulfur/grease fire when a bearing on the application equipment burned. He put out the fire with an extinguisher. The fire did not spread to the hopper full of sui fur.
87-772
04/2411987
Imperial
Aircraft pilot's eyes were burned when the plane caught on fire. Sulfur smoke and fumes were released into the cockpit. The cornea of both eyes were burned.
87-2122
051l4/1987
San Mateo
Worker became ill from burning suI fur fumes while dusting sulfur on ornamentals. He apparently spilled dusting sulfur on the muffler of the motorized hand duster during mixing and loading. Symptoms: nausea, dizziness, vomiting, and wheezing.
reactions. Five instances of apparent allergic reaction to 1% elemental sulfur, ranging from 1+ to 2+/3+,3 were also observed in a study of California nursery workers. Reactions to sulfur 3Reactions were scored as 1+ (weak reaction, macular erythema), 2+ ( strong reaction, edematous or vesicular), or 3+ (extreme reaction, spreading, bulbous, ulcerative). Equivocal reactions are designated as +1- (Adams, 1990).
were correlated with a history of working as a pesticide applicator in the nursery business, but none of the participants had a detailed memory of pesticides they had handled or specifically remembered spraying elemental sulfur (O'Malley and Rodriguez, 1998). Two case reports implicate elemental sulfur as a human contact allergen. Schneider (1978) reported two
References
cases of contact allergy in patients who used medications containing elemental sulfur to treat superficial fungal dermatoses. Both patients had positive patch test reactions to 5% elemental sulfur in series of vehicles, but a control series was not reported. Wilkinson (1975) reported that a professional gardener with a previous history of atopic eczema developed an eczematous eruption involving the elbow ftexures and the right hand. He had a positive patch test reaction to 5% sulfur in petrolatum, but a control series was not reported. Gregorczyk and Swieboda (1968) described 15 cases of desquamative dermatitis among 425 Polish sulfur miners in which irritant dermatitis due to elemental sulfur may have played a part.
REFERENCES Adams, R. M. (1990). "Occupational Skin Disease." Saunders, Philadelphia. Beauchamp, R. 0., Jr., Bus, J. S., Popp, J. A., Boreiko, c., and Andjelkovich, D. A. (1984). A critical review of the literature on hydrogen sulfide toxicity. CRC Crit. Rev. Toxicol. 13,25-97. Bulgin, M. S., Lincoln, S. D., and Mather, G. (1996). Elemental sulfur toxicosis in a flock of sheep. J. Am. Vet. Med. Assoc. 7, 1063-1065. Costa, D. L., and Amdur, M. O. (1996). In Casarett and Doull's "Toxicology, the Basic Science of Poisons" (c. D. Klaassen, ed.), 5th ed. McGraw-Hill, New York. DPR (1995). "Summary of Pesticide Use Report Data Annual 1993." Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA. DPR (1996a). "Summary of Pesticide Use Report Data Annual 1994." Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA. DPR (1 996b ). "Summary of Pesticide Use Report Data Annual 1995." Department of Pesticide Regulation, California Environmental Protection Agency, Sacramento, CA. Farm Chemicals Handbook (1998). Volume 84 (R. T. Meister, ed.). Meister Publishing Company, USA. Gador, J., and Motowicka-Terelak, T. (1986). Effect of contamination with sulfur on soil properties and crop yields in a Iysimeter experiment: n Effect of elemental sulphur application to the soil on the yields and chemical compositions of some crops. Pamietnik Pulawski 88, 25-38. Gibson, D. M., Kennelly, J. J., and Arherra, F. X. (1987). The performance and thiamine status of pigs fed sulfur dioxide treated high-moisture barley. Can. J. Anim. Sci. 67, 841-854. Gould, D., McAllister, M. M., Savage, J. c., et al. (1991). High sulfide concentration in rumen fluid associated with nutritionally induced polioencephalomalacia in calves. Am. 1. Vet. Res. 52, 1164-1169. Gregorczyk, L., and Swieboda, K. (1968). Uber den einfluB von schwefelverbindungen auf die haut und auf die schleimhaute [On the influence of sulfur binding on the skin and mucous membranes]. Polskie Tygognik Lekarski 23, 463-466.
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Hamlen, H., Clark, E., and Janzen, E. (1993). Polioencephalomalacia in cattle consuming water with elevated sodium sulfate levels: A herd investigation. Can. Vet. J. 34, 153-158. Haynes, R. J., and Swift, R. S. (1986). Effects of soil amendments and sawdust mulching on growth, yield and leaf nutrient content of highbush blueberry [Vaccinium corymbosum cultivar B1uecrop] plants. Scientia Hortic. 29(3), 229-238. Hill, F. 1., and Ebbett, P. C. (1997). PoIioencephalomalacia in cattle in New Zealand fed chou moellier (Brassica oleracea). New Zealand Vet. J. 45, 37-39. Jeffrey, M., Duff, J. P., Higgins, R. J., Simpson, V. R., Jackman, R., Jones, T. 0., Mechie, S. C., and Livesey, C. T. (1994). Polioencephalomalacia associated with the ingestion of ammonium sulphate by sheep and cattle. Vet. Rec. 134, 343-348. Jensen, R., Griner, L. A., and Adams, O. R. (1956). Polioencephalomalacia of cattle and sheep. J. Am. Vet. Med. Assoc. 129,311-321. Low, J. C, Scott, P. R., Howie, F., Lewis, M., Fitzsimons, J., and Spence, J. A. (1996). Sulfur-induced polioencephalomalacia in lambs. Vet. Rec. 138, 327-329. Majule, A. E., Topper, C. P., and NortcIiff, S. (1997). The environmental effects of dusting cashew (Anacardium occidentale L.) trees with sulphur in southern Tanzania. Trap. Agric. 74(1), 25-33. Matsushita, T., Yoshioka, M., Aoyama, K., Arimatsu, Y., and Nomura, S. (1977). Experimental study on contact dermatitis caused by fungicides benomyl and thiophanate-methyl. Ind. Hlth. 15(3-4), 141-148. McAIlister, M. M., Gould, D. H., and Hamar, D. W. (1992). Sulfide-induced polioencephalomalacia in lambs. J. Camp. Pathol. 106, 267-278. Olkowski, A. A. (1997). Neurotoxicity and secondary metabolic problems associated with low to moderate levels of exposure to excess dietary sulfur in ruminants: A review. Vet. Human Toxicol. 39,355-360. O'Malley, M. A., and Rodriguez, H.-P. (1998). "Contact Dermatitis in California Nursery Workers: Part n. Pilot Field Study." California EPA, DPR, Worker Health and Safety Branch HS-1767. Sager, R. L., Hamar, D. w., and Gould, D. (1990). Clinical and biochemical alterations in calves with nutritionally induced polioencephalomalacia. Am. J. Vet. Res. 51,1969-1974. Schneider, H. G. (1978). Schwefelallergie [Sulfur allergy]. Hautarzt 29,340342. Shapiro, R. (1977). Genetic effects of bisulfite (sulfur dioxide). Mutation Res. 39, 149-176. Studdert, V. P., and Labuc, R. H. (1991). Thiamine deficiency in cats and dogs associated with feeding meat preserved with sulphur dioxide. Aust. Vet. J. 68,54-57. Sun, C. C, and Sue, M. S. (1995). Sulfur spring dermatitis. Contact Dermat. 32,31-34. Terlecki, S., and Markson, L. M. (1961). Cerebrocortical necrosis in cattle and sheep. Vet. Rec. 73, 2327. Vena, G. A., Foti, C., and Angelini, G. (1994). Sulfite contact allergy. Contact Dermat. 31, 172-175. Wilkinson, D. S. (1975). Sulfur sensitivity. Contact Dermat. 1,58.
CHAPTER
83 Rodenticides Alain F. Pelfrene Alain Pelfrene & Associated Consultants
83.1 INTRODUCTION Rats and mice compete with humans for food. This loss to rodents causes economic loss everywhere. In some developing countries it can cause starvation. Rodents are also hosts for human diseases, including plague, endemic rickettsiosis, leishmaniasis, spirochetosis, tularemia, leptospirosis, tick-borne encephalitis, and listeriosis. Rats occasionally bite people. Finally, rodents do a variety of other damage, mainly by gnawing. Insofar as possible, rodent populations should be controlled by limiting their access to food and harborage. Individual animals or small groups may be removed conveniently by trapping. However, there will always be a need for poisons in rodent control. Unfortunately, effective permanent control through poisoning is not simple. The animals must be enticed to ingest a toxicant in sufficient dosage if the effort is to succeed. But rodents rarely or ever constitute an important problem unless they have a supply of food and water. This means that, in spite of containing a foreign substance, the solid or liquid bait used should be at least as attractive to the rodents as their usual supply of food or water. The first problem may be that the intended poison makes the bait unacceptable to animals that have never encountered the poison. This is called primary bait refusal. Because of this common difficulty, many efforts to find better rodenticides have emphasized highly toxic substances of such bland taste and odor that animals always will take a lethal dose the first time. However, this is an impractical objective. At least a few animals will get only a sublethal dose on first encounter and will be conditioned thereby to avoid the poison even though it seems tasteless and odorless at first. This reaction is called secondary bait refusal or bait shyness. There is even some indication that rodents learn from the behavior of their companions in such a way that the manner of death of some of them conditions the behavior of others that have consumed no poison. A third problem with many rodenticides is that they are very nearly as dangerous to humans and useful animals as to rodents. This problem can be minimized by selecting a poison with a wide margin of safety, by coloring the bait, by combining the poison with an emetic, or by restricting the placement of baits. However, all these solutions have their Handbook of Pesticide Toxicology Volume 2. Agents
limitations. There is no rat poison that cannot harm humans if sufficiently misused. Considering all these difficulties, four requirements for an ideal rodenticide may be stated as follows: (a) The poison must be surely effective when incorporated into baits in such small quantity that its presence is not detected to an interfering degree. (b) Finished baits containing the poison must not excite bait shyness in any way and the necessity of prebaiting must thereby be avoided. (c) The manner of death must be such that surviving individuals will not become suspicious of its cause but will remain on the premises and eat freely of the bait until the themselves die. (d) The poison, in the concentration used for control, must be specific for the species to be destroyed unless its use can be made safe for humans and domestic animals by some other means. Part of the safety of the anticoagulant rodenticides is made possible by their cumulative properties and depends on the fact that they are offered to rodents in such a way that a single dose is harmless even to the rodents themselves. Quite aside from the important species differences in susceptibility, which favor human safety, people are protected further by the fact that except in suicide or murder, substantial continuing exposure is far less likely than a single accidental exposure. Although this chapter is devoted to synthetic organic rodenticides, it is necessary to recall that inorganic and botanical compounds may still be important for rodent control in some areas. Furthermore, some of them, notably arsenic, phosphorus, and strychnine, are very important as sources of human poisoning. On the contrary, some synthetic rodenticides have other uses. The most important examples are the use of vitamin D and of certain anticoagulants as rodenticides and as drugs in human medicine. In addition, several of the organic fluorine compounds have been used experimentally or in practice as systemic insecticides and/or raticides. Compared to toxic substances in general, biochemical actions of the synthetic rodenticides that have been studied in humans are unusually well known. This makes it possible to assign them to groups (see the following sections) that are meaningful not only chemically but in terms of biochemical lesions. The same is not true of numerous miscellaneous com-
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Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
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pounds including crimidine (BSI, ISO) that apparently have not yet produced poisoning in humans and certainly have not been used as drugs or studied experimentally in human subjects.
11 FCH,-C-ONa Sodium monofluoroacetate
83.2 FLUOROACETIC ACID AND ITS DERIVATIVES
2·Fluoroacetamide
2·F!uoroethanol MNFA
Sodium fiuoroacetate came to prominence in the United States as a result of a search for rodenticides that would not be subject to shortages imposed by World War 11 (Ward, 1945). This and related compounds had been considered earlier as systemic insecticides. At about the same time, it became known that fiuoroacetate is the toxic material in the South African plant "gifblaar" (Dichapetalum cymosum). Later it was shown that the same compound is present intermittently in Acacia georgiana. The main toxic ant in D. toxicarium is fiuorooleic acid, but fiuoropalmitic acid is present also. The ground seeds of D. toxicarium have been used by natives as a rat poison. It gave problems with secondary poisoning and human toxicity similar to that later associated with synthetic sodium fiuoroacetate. Under these circumstances, there were practical as well as academic reasons to study the mode of action of organic monofiuoro compounds. It appears that the toxicity of all of the compounds depends on the same mechanism. Highly toxic compounds either have two carbon atoms or are metabolized to this form (Chenoweth, 1949; Peters, 1963a; Raasch, 1958; Saunders, 1947). 83.2.1 SODIUM FLUOROACETATE 83.2.1.1 Identity, Properties, and Uses Chemical Name name. Structure
Sodium monofiuoroacetate is the chemical
See Fig. 83.1.
Synonyms Sodium monofiuoroacetate is also known as Compound-I 080 or ten-eighty. The CAS registry number is 62-74-8. Physical and Chemical Properties The empirical formula for sodium monofiuoroacetate is C2H2FNa02 and the molecular weight is 100.3. Its forms an odorless, white, nonvolatile powder that decomposes at about 200°C. Although the compound is often said to be tasteless, dilute solutions actually tasted like weak vinegar. Sodium fiuoroacetate is very water soluble and hygroscopic but is of low solubility in ethanol, acetone, and petroleum oils. Formulations and Uses Sodium fiuoroacetate is formulated as an aqueous solution containing a warning color. Sodium monofiuoroacetate is used to kill rats, mice, other rodents, and predators. It is an intense mammalian poison, and it is used in many countries but only by trained personnel.
OH
011
)
Figure 83.1 pesticides.
)
FCH,-CII,-CH,F
FCH,-C-CH,C1
Glycerol difluorohydrln
Glycerol chlorofluorohydrin
Some organic fluorine rodenticides and other organic fluorine
83.2.1.2 Toxicity to Laboratory Animals Basic Findings The first paper on sodium monofiuoroacetate as a rodenticide (Kalmbach, 1945) drew attention to its very high acute toxicity. LD 50 values for ordinary laboratory rats and for wild animals of the same species were reported as 2.5 and 5.0 mg/kg, respectively. The wild black rat (Rattus rattus), another commensal species, was much more susceptible (LD 50:0.1 mg/kg). A LD 50 of 0.22 mg/kg has been reported for Rattus norvegicus (Dieke and Richter, 1946). The likelihood of danger to people, domestic animals, pets, and nontarget wildlife was pointed out. The acute toxicity of the compound to an extremely wide range of wildlife was reported by Ward and Spencer (1947). See Table 83.1. Fluoroacetate acts mainly on the central nervous system and the heart. It seems that there are species in which fiuoroacetate affects chiefiy the heart, such as the rabbit, the goat, and the horse, and others in which only the central nervous system is affected, such as the dog, the guinea pig, and the frog. In the cat, the rhesus monkey, the domestic pig, and birds, both systems are involved. The above results were obtained by Chenoweth and Gilman (1946) using methyl fiuoroacetate instead of sodium fiuoroacetate. However, since both compounds yield the fiuorocitrate ion in the body, where it is converted to fiuoroacetate, which is responsible for the induction of pharmacologic and toxic signs (see below), it seems that this experiment is nevertheless interesting in showing a large degree of species variability in the site of action. In all species, there was a delay of 0.5-2 hr or more between administration, either oral or intravenous, and the onset of the symptoms, and the route of administration did not significantly affect the toxicity of fiuoroacetate. Laboratory rats acquire a tolerance to sodium fiuoroacetate by ingesting sublethal doses over a period of 5-14 days. However, this tolerance is lost if intake of the compound is interrupted for as little as 7 days (Kalmbach, 1945). Tolerance of some but not all species was confirmed by several investigators, including Kandel and Chenoweth (1952). These authors found that, whereas small doses of fiuoroacetate increased tolerance to challenge doses of fiuoroacetate or
83.2 Fluoroacetic Acid and Its Derivatives Table 83.1 Single-Dose LD 50 for Sodium F1uoroacetate LD50 (mg/kg)
Reference
Species
Route
Rat
oral
0.22
Rat
oral
2.5
Ward and Spencer (1947)
Rat
oral
1-2
Phillips and Worden (1957)
Rat
intraperitoneal
3-5
Ward and Spencer (1947)
Mouse
subcutaneous
19.3
Hutchens et al. (1949)
Mouse
subcutaneous
17.0
Tourtellotte and Coon (1951)
Mouse
intraperitoneal
10.0
Ward and Spencer (1947)
Mouse
intraperitoneal
16.5
Tourtellotte and Coon (1951)
Mouse
intraperitoneal
14.7
Raasch (1958)
Guinea pig
oral
0.4
Guinea pig
intraperitoneal
0.37
Hutchens et al. (1949)
Rabbit
subcutaneous
0.28
Hutchens et al. (1949)
Dog
oral
0.06
Tourtellotte and Coon (1951)
Cow
oral
0.39
Robinson (1970)
Calf
oral
0.22
Robinson (1970)
Opossum
oral
0.79
Bell (1972)
Mallard duck
oral
4.8
Hudson et al. (1972)
South African
oral
500
Dieke and Richter (1946)
Ward and Spencer (1947)
Chenoweth (1949)
clawed toad (Xenopus laevis)
4-ftuorobutyrate, tolerance to neither could be evoked by small doses of 4-ftuorobutyrate. The citrate content of the rat brain appeared to have no relation to tolerance, and the citrate that accumulated after a small dose did not prevent the further accumulation of citrate after a larger dose. The sensitivity of mice to sodium ftuoroacetate depends on temperature. Under otherwise identical conditions, the LD 50 values were 12.1 and 5.16 mg/kg at 23 and l7°C, respectively (Misustova et aI., 1969). The survival of individual rats in a particular dosage group may be predicted by following their body temperature. There is a critical level that varies somewhat according to the interval after dosing. Animals that regained their initial temperature within 96 hr usually lived, but those that failed to regain normal temperature within this time usually died (Filip et aI., 1970). For groups of animals, the course of the temperature can be described by a computer-generated curve (Hosek and Love, 1952). The temperature change is correlated with citrate metabolism (Kirzon et aI., 1970). Primates and birds are more resistant; rodents and carnivores are most susceptible. In general, cold-blooded vertebrates are less sensitive than warm-blooded ones (Egekeze and Oehme, 1979). Sodium monoftuoroacetate in carcasses creates a secondary poisoning hazard to which carnivorous predators are extremely susceptible (Bell, 1972). Plasma elimination half-life in rabbits was shown to be 1.1 hour and the retention time in tissue greater with larger
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doses. Tissues residues were substantially lower than in plasma (Gooneratne et aI., 1995). When orally administered to sheep and goats at dose levels of 0.1 mg/kg body weight, the plasma elimination half-life was found to be 10.8 hours in sheep and 5.4 hours in goats. Concentrations of sodium ftuoroacetate in muscle, kidney, and liver (0.042,0.057, and 0.021 J.Lg/g) were clearly lower than those in the plasma (0.098 J.Lg/g) 2.5 hours after administration. After 96 hours, only traces of the compound were detectable in sheep tissues «0.002 to 0.008 J.Lg/g). The authors concluded that even with accidental exposure to sublethal doses, sodium ftuoroacetate would not persist in tissues for more than a few days because of its rapid clearance and because occurrence of residues in meat intended for human consumption would be highly unlikely (Eason et aI., 1994). Absorption, Distribution, Metabolism, and Excretion Using ether as a solvent, it was possible to recover 60-70% of the total dose from the body (including gastrointestinal contents) of rabbits killed by 10 times the LD 50 level. The concentration in the brain was twice that in other organs (Tomiya et aI., 1976). Sodium monoftuoroacetate is rapidly absorbed by the gastrointestinal tract. It is not well absorbed by the intact skin, but absorption may be greater in the presence of dermatitis or other skin injury. Biochemical Effects It was in connection with the mode of action of ftuoroacetic acid that the term "lethal synthesis" was coined (Peters, 1952). Peters (1963b) later reviewed the research and extended the concept. Very briefty, no mammalian enzyme was found that was inhibited by ftuoroacetate in vitro. However, in vivo, the ion undergoes synthesis to form ftuorocitrate and this inhibits mitochondrial aconitase either in vivo or in vitro. The result is that the Krebs cycle is blocked, which leads to lowered energy production, reduced oxygen consumption, and reduced cellular concentration of ATP; furthermore, since the citrate synthetase continues to work, citrate accumulates in the tissues (Buffa and Peters, 1950). It is thought that toxicity is due not to the accumulation of citrate per se but to the blockage of energy metabolism. However, increased tissue and plasma concentration of citrate is probably responsible for some of the symptoms seen during acute poisoning. Citrate is a potent chelator of calcium ion, and it has been demonstrated that in cats intravenously injected with ftuoroacetate at 0.03 mmol/kg the ionized calcium level in blood fell by an average of 27.2%,40 min after the injection. There was a corresponding prolongation of the QT interval of the electrocardiogram (ECG), and treatment with CaCh significantly prolonged the life of the treated animals as compared with unmedicated positive controls (Roy et aI., 1980). The characteristic delay at the onset of poisoning by sodium fturoacetate is accounted for by the time necessary for its metabolism and biochemical mode of action. The toxicity of ftuoroacetate is entirely different from that of inorganic ftuorides. It depends on the firmness of the F-C bond such that ftuoroacetate is an antimetabolite.
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Consistent with the theory that fluorocitrate is the active toxicant, it was found to be at least 100 times more toxic than fluoroacetate when injected directly into the brain under various experimental conditions. An intracerebral dose of 0.115 I-lg failed to kill rats weighing about 250 gm, and it did not cause convulsions; doses of 0.287 I-lg (about 0.001 mg/kg) or greater caused convulsions and killed almost all rats (Morselli et aI., 1968). On the other hand, a dosage of 40-60 mg/kg is necessary to kill by the intraperitoneal route, and an oral dosage of 40 mg/kg constitutes only an LD 50. The great difference was attributed to failure of fluorocitrate to reach aconitase within critical cells of the brain and heart (Peters and Shorthousc, 1971). Species differ in the degree to which the concentration of citrate increases in different organs and also in the timing of these increases (Kirzon et aI., 1973). These biochemical differences presumably underlie the clinical differences between species, especially the relative importance in neurological and cardiac effects. Accumulation of citrate was evident in mice within 2 hr after intraperitoneal injection of sodium fluoroacetate at a rate of 30 mg/kg, which is about 1.7 times the LD 50 in that species. The concentration of citrate increased from 48 ppm in controls to 74, 101, and 166 ppm within 2, 5, and 24 hr, respectively, after injection. The mice were dead at 24 hr (Matsumura and O'Brien, 1963). Whereas Williamson et al. (1964) agreed that the initial effect of fluoroacetate is to produce fluorocitrate, they considered that the secondary inhibition of phosphofructokinase by the accumulated citrate was actually lethal because it deprived the cell of pyruvate, which would eventually overcome the inhibition of aconitase. Effects on Organs and Tissues Loracher and Lux (1974) concluded on the basis of studies of neuromembrane depolarization that diminished inhibitory conductance is apparently important as a causative factor in convulsions induced by sodium fluoroacetate. The decreased level of ionized calcium in blood induced by the chelating effect of citrate certainly plays a role in the depolarization of the neuromembrane, as it does on the cardiac cell membranes. The effect of sodium fluoroacetate on the heart rhythm is due, as demonstrated by Noguchi et al. (1966), primarily to action on the cells themselves and not on the vagus nerve. Irregularity of rhythm and a condition analogous to fibrillation were produced in cultures of heart cells that had grown until cell-to-cell contact was prevalent and beating was synchronized. The average times necessary to produce irregularity and fibrillation were 9 and 48 hr, respectively, at a concentration of 10 ppm in the medium, but only 2 and 9 hr, respectively, at a concentration of 100 ppm. At a concentration of 1000 ppm, fibrillation was immediate and cytoplasmic vacuoles appeared rapidly. Effects on Reproduction A dosage of sodium fluoroacetate just below the maternal LD 50 reduced oxygen consumption of the embryos as well as the mother but was not teratogenic (Spielmann et aI., 1973).
Treatment of Poisoning in Animals Hutchens et al. (1949) demonstrated a significant reduction of mortality in mice, guinea pigs, and rabbits (but not dogs) treated with ethanol at a rate of 800 mg/kg administered subcutaneously as a 10% solution in normal saline. The response occurred when the alcohol was given before signs of poisoning appeared and was best when given within 10 min of poisoning. In mice, sodium acetate and ethanol acted synergistically to antagonize poisoning (Tourtellotte and Coon, 1951). The beneficial effect of ethanol in rodents was confirmed by Chenoweth et al. (1951), but these authors found ethanol less effective in the dog and utterly useless in the monkey. In a study of a wide range of chemical substances in mice, rats, rabbits, dogs, and rhesus monkeys, they concluded that commercially available monacetin containing about 60% glycerol monoacetate was superior to any other substance tested as an antidote for poisoning by fluoroacetate. Not only did it reduce mortality, but it was able to normalize heart and brain rhythms as indicated by ECG and electroencephalogram (EEG) tracings. Light pentobarbital anesthesia for 18-24 hr significantly reduced mortality among dogs poisoned by sodium monofluoroacetate at a rate of 0.10 mg/kg (Hutchens et aI., 1949; Tourtellotte and Coon, 1951). 83.2.1.3 Toxicity to Humans Accidental and Intentional Poisoning Sodium fluoroacetate was introduced in 1946 in the United States for use by pest control operators, including persons hired for the purpose by government agencies. The poison was mixed with a dye. Solutions were supposed to be placed in shallow paper cups made in such a way that they would not tip over. These water baits were supposed to be used only in places that would be unoccupied and locked during exposure of the poison, and all cups and dead rodents were supposed to be collected and incinerated by authorized persons at the end of the exposure period. However, the regulations were not always followed. By the end of the year, at least one child who found an "empty" paper cup had died, and her 3-year-old brother had been severely poisoned. By the end of 1949, there had been at least 12 deaths and 6 cases of nonfatal poisoning. In addition, there had been 4 deaths, all in children, that probably were caused by sodium monofluoroacetate, but other sources of poisoning could not be ruled out. Of the 12 deaths clearly caused by sodium monofluoroacetate, 5 involved small children who had found and often chewed on a poison cup, 3 involved juveniles who had found the poison in a soft drink bottle, and 4 were suicides of adults. Except one, each of the survivors was a child who had found a poison cup. These accidents made such an impression on the few people who had legal access to sodium fluoroacetate that they became far stricter in carrying out the recommended precautions and in selecting situations in which the compound was used at all. As a result, the safety record of the compound in the United States improved greatly. A typical fatal case involved a 40-year-old man who was found unconscious in his bedroom. He had an 8-year history
83.2 Fluoroacetic Acid and Its Derivatives
of severe depression, and his family had been warned of the possibility of suicide. When admitted to hospital, he had slight muscular spasms and nystagmus of both eyes; the heart rate was 92 beats per minute and rhythm was irregular. Following gastric lavage and a soft soap enema, the nystagmus became worse, and the patient had an epileptiform convulsion. The blood pressure fell to 90/40 mm Hg. Treatment consisted of plasma, oxygen, and procaine hydrochloride in the hope of desensitizing the heart. The blood pressure improved to 118175 mg Hg, but there was no decisive change until the heart and later the respiration stopped about 17 hr after admission (Harrisson et aI., 1952a). Another fatal case was remarkable for its combination of prolonged survival following the ingestion of an almost certainly very large dose. Briefly, about 113,000 mg of sodium fluoroacetate was missing from a professional rat exterminator's supplies after his 17-year-old son made a solution and drank it. The boy vomited promptly and then within an hour walked into an hospital emergency room. He gradually became comatose during gastric lavage, and consciousness was never regained. Within less than 3 hr of ingestion, he had a grand mal convulsion associated with fecal incontinence. The clinical course, which lasted slightly over 5 days, was characterized by cardiac irregularity, which responded to a considerable degree to procainamide hydrochloride; dilation and failure of the heart with acute pulmonary edema, which responded surprisingly well to digitalis (lanatoside C); bouts of severe hypotension, which responded only questionably to levarterenol (norepinephrine) but somewhat better to mephentermine; cortical irritability, which responded to barbiturates and later responded more effectively to ethanol; frequent severe carpopedal spasm, controlled somewhat by calcium gluconate; and finally growing evidence of infection including a temperature reaching 42.3°C in spite of efforts to reduce it. The diagnosis based on autopsy was poisoning, bronchopneumonia with septicemia, focal infarction of the right kidney, and mediastinal emphysema (Brockmann et al., 1955). Serious illness followed by full recovery occurred in a 2-year-old boy who was found licking crystals from the screw cap of a bottle of sodium fluoroacetate solution. The parents did not know whether he drank any of the solution. Almost immediately after he was found, the boy began to vomit. He was brought to hospital about 6 hr later because he began to have generalized convulsive movements and became stuporous. On admission, the boy was comatose and exhibiting carpopedal spasms, tetanic convulsive movements, irregular respiration, and great cardiac irregularity. While a solution of calcium gluconate was being injected, there were a few seconds of cardiac asystole. Thereafter, the irregular cardiac rhythm resumed but at a much slower rate. Tetanic convulsions stopped immediately, and the child became completely flaccid. A few hours after admission, the child became responsive. Very soon the boy suffered a generalized tonic clonic convulsion lasting several minutes and followed by deep coma. Briefly, the boy remained unresponsive for 4 days. Cardiac rhythm continued to change frequently during the first 3 days. Tonic convulsions lasting several minutes occurred many times every hour, sometimes about
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every 10 min for many successive hours. During spasm, the pupils dilated and remained inactive to light; between seizures the pupils were miotic but responsive to light. On two occasions respiration stopped and artificial respiration was required briefly. On the evening of the fourth day, 100 hr after ingestion, the boy began to open his eyes and look about. He tried to talk but was unable to articulate. He could neither sit up nor reach for objects but appeared alert. On the fifth day and sixth days, he rapidly regained all his motor ability, slowly lost his drowsiness, and became articulate. On the evening of the sixth day he was clinically well. He was discharged on the eleventh day. Reexamination 1 year later showed that the boy had had no further neurological trouble, and this mental and physical development has proceeded normally (Gajdusek and Luther, 1950). In another case in which the initial dosage undoubtedly was smaller, there were no important clinical changes until 20 hr after ingestion, when the 8-month-old girl had a generalized seizure lasting about 1 min. In spite of treatment with phenobarbital, three additional seizures occurred during the next 12 hr. There was no further illness, and the patient was discharged 4 days later. Follow-ups revealed no change in behavior, intellect, or motor performance (Reigart et aI., 1975). Any serious but reversible interference with respiration or general circulation is liable to produce some cases in which the patient survives but with severe brain damage. The cardiac arrhythmias characteristic of poisoning by sodium fluoroacetate are likely to produce such interference. An example involved an 8-year-old boy who was in status epilepticus when he entered hospital. The convulsions were controlled to some degree. There was no striking change until 14 hr after admission, when ventricular asystole occurred. Heart action was renewed but only after sufficient delay that the child suffered brain damage and was clearly mentally defective after a very long and stormy hospital course (McTaggart, 1970). During the decade 1971-1981, 111 cases of accidental or unintentional poisoning with sodium fluoroacetate were collected by the National Poison Center of Israel. These cases included three cases of death and one case of mass accidental poisoning affecting 30 children, although the great majority of them only consumed a very small number of wheat grain baits impregnated with the compound. These latter cases did not result in clinical symptoms of poisoning (Roy et aI., 1982). These authors also described the clinical features of two cases of acute poisoning in which gastrointestinal disorders were rapidly followed by central nervous system manifestations (disorders of consciousness, convulsions, coma) and cardiac disorders, the most frequent cause of death. Ventricular ectopic beats preceded the ventricular arrhythmia, which was then followed by ventricular tachycardia and fibrillation. The electrocardiogram was characterized by a prolonged QT interval. A metabolic acidosis was commonly observed. Chung (1984) reported on five cases collected between 1975 and 1981 in Taiwan. The amount ingested ranged from 8 to 40 ml of a 1% formulation of sodium monofluoroacetate. All five patients survived. All cases had signs of transient cardiac dysfunction, but in addition acute re-
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CHAPTER 83
Rodenticides
nal failure was seen in three of the five patients, two of them with frank uremia. The acute renal failure was reversible. In a retrospective study of 38 cases of poisoning collected between 1988 and 1993 in Taiwan by Chi et al. (1996), 18% of the patients died. Laboratory symptoms included nonspecific ST-T and T waves on the ECG (72%), hypocalcemia (42%), and hypoka1emia (65%). Hypotension, respiratory rate, pulse rate, increased serum creatinine, and decreased pH were considered as the most important predictors of mortality. Use Experience In spite of the great toxicity of sodium fluoroacetate, there apparently has been only one case of illness among those who used it without suicidal intent. Even in this case, the kind of illness was so atypical of poisoning and so complicated by the unrelated factor of hypertrophy of the prostate that evaluation of the case is difficult. The patient entered hospital with renal failure and other serious illness. Although he was only 59 years old, he had a 5-year history of symptoms of prostatism but no history of urinary tract infection, renal calculi, or hematuria. He had had gout for 10 years, and he had been digitalized for 12 months. For 6 months he had experienced increasing lassitude, vomiting, and pruritis. Inspection revealed rapid breathing and muscle wasting. More detailed physical examination revealed mild left ventricular failure and evidence of liver disease, hypothyroidism, extrapyramidal disease, and gout, as well as distended bladder, caused by prostatic hypertrophy. These findings were substantiated by laboratory examinations. Following catheter drainage of the bladder, blood urea declined and renal function improved further following prostatectomy 10 days after admission. Recovery was very slow. Neurological and thyroid findings cleared within 6 months. Renal function continued to improve for about 2 years, after which the patient remained well. Involvement of fluoroacetate was suspected because of the history of exposure, the finding of organic fluorine in the urine, and histological changes found in kidney biopsies. There was no doubt of exposure; the patient had been employed for 10 years as an exterminator of rabbits, and for about 4 weeks each year he had applied sodium fluoroacetate to pieces of carrots that served as bait for the animals. During this work, he had worn rubber gloves and he had never knowingly ingested any of the poison. The report of concentrations of 15.4 and 14.8 ppm of sodium fluoroacetate (analyzed as organic fluoride) in two samples of urine collected 2 weeks after admission and the absence of such organic fluoride in samples collected 5 weeks and 6 months later was accepted as consistent with the history of exposure. A kidney biopsy performed 4 days after admission revealed periglomerular fibrosis, some capsular adhesion and other glomerular changes, plus swelling and vacuolation of tubular cells, increased interstitial fibrous tissue, a few small foci of inflammation, and mild thickening of the arterial walls. A second biopsy 4 weeks later showed little change in the glomerular lesion, but the tubules were no longer vacuolated. However, many tubules had been lost, and many of those remaining were atrophic. Interstitial fibrosis was prominent. The kidney lesions were considered similar to those described in
rats in association with acute poisoning. It was acknowledged that lower urinary tract obstruction may have been a predisposing factor (Parkin et aI., 1977). Even if the patient had been exposed to sodium fluoroacetate a short time before he was admitted to hospital, it is difficult to understand why excretion of organic fluoride from this source would continue 2 weeks later. Although no urinary levels of organic fluoride have been reported for other workers, one must note that 15 ppm would indicate a daily output of about 22.5 mg/person/day in a person with average urinary volume. This in turn would indicate a minimal absorption rate of about 0.32 mg/kg/day, an astonishingly high level. The renal changes previously described in rats (Cater and Peters, 1961) followed one or a few very large dosages of fluorocitrate, and the fat droplets were tiny compared to those seen in the human patient. Dosage Response In a fatal case, 465 mg (equivalent to a dosage of over 6 mg/kg) was recovered from the stomach contents, urine, brain, liver, and kidneys (Harrisson et aI., 1952b). No account was taken of sodium fluoroacetate in other organs and tissues or of that removed by vomiting, lavage, and enema; therefore, the ingested dosage must have been considerably larger. Several children varying in age from 0.66 to 8.0 years were poisoned seriously or even fatally by chewing on only one paper cup placed earlier for rat control. The cups were made to receive 15 ml of 0.33% solution, that is, 50 mg of sodium fluoroacetate. The average age of the children was 2.37 years, and the weight of such a child is about 13 kg. Thus, the maximal dosage must have been approximately 3.8 mg/kg, but the true dosage must have been considerably smaller because part of the material originally added to the cup may have been lost and not all that dried in the cup would have been ingested. A dosage of 0.5-2.0 mg/kg must be considered highly dangerous. The estimated mean lethal dose in humans ranged from 2 to 10 mg/kg (Gajdusek and Luther, 1950; Harrisson et aI., 1952a). Laboratory Findings The following concentrations expressed as sodium fluoroacetate were found in samples taken at autopsy from a man who survived about 17 hr after being found unconscious: urine, 368 ppm; liver, 58 ppm; brain, 76 ppm; and kidney, 65 ppm (Harrisson et aI., 1952b). Pathology In a fatal case, autopsy revealed petechial hemorhages and congestion of the organs consistent with recent fits. All the findings were nonspecific, but it is interesting that they included diffuse tubular degeneration of the kidneys, which is consistent with the findings in the only case of alleged chronic human poisoning by sodium fluoroacetate. Treatment of Poisoning Apparently, most patients who survived poisoning by sodium fluoroacetate as well as those who died of it received no medication that offered any possibility of specific antidotal action. In at least one case (unpublished), a poisoned child was treated with whiskey and survived. Unfortunately, no details are available, and there can be no assurance
83.2 Fluoroacetic Acid and Its Derivatives
that the child would not have progressed equally well without treatment. Although monacetin apparently has not been administered to a human patient, the work of Chenoweth et at. (1951) in various animals, especially monkeys, offered good reason to think it would be valuable for treating human poisoning. They recommended that it be injected intramuscularly at least every hour for several hours at the rate of 0.1-0.5 ml/kg per injection. There is no clinical evidence for or against the use of acetate in humans. On the contrary, acetamide has been administered to patients, and it seemed to be the reason for their survival. It is available at Accident and Emergency Departments throughout New Zealand. Acetamide is administered intravenously as a 10% solution in 5% glucose. In severe cases, 500 ml is given in 30 min every 4 hr; in milder cases, 200 ml is given on the same schedule. There can be no doubt that removal of the poison and supportive care are indicated. A number of patients have shown clear-cut poisoning but survived without sequelae following such treatment. Supportive care should include continuous cardiac monitoring. There is strong clinical evidence that the danger of cardiac arrhythmia can be reduced significantly by judicious and continuing use of procainamide hydrochloride. Even so, equipment for defibrillation should be ready. There is reason to hope it would be successful if required because at least one patient was revived with only external massage of the heart. There is also clinical evidence that cortical irritability can be lessened by barbiturates. There is no basis for speculating on the value of diazepam in this connection. Contrary to the evidence in monkeys, clinical evidence in humans has indicated that ethanol is beneficial and perhaps superior to barbiturates. Whereas the effect seemed to involve needed sedation, the possibility of a more fundamental effect in the biochemical lesion was not excluded. On the basis of laboratory studies, Chenoweth et at. (1951) recommended against administration of calcium, potassium, sodium chloride, bicarbonate, or acetate. They considered that any necessary replacement of fluid should be done cautiously with plasma, and they considered digitalization as definitely contraindicated. However, clinical experience argues strongly against two of this prohibitions, and there is no clinical evidence to support some of the others. Calcium gluconate has proved useful in controlling carpopedal spasm, including such spasm in a patient who survived whithout sequelae. Digitalis (lanatoside C) not only improved the function of a poisoned heart that had failed to the point of acute pulmonary edema but also produced no detectable side effects. Finally, there is clinical evidence that mephentermine is more effective than levarterenol in raising blood pressure if that becomes necessary in the course of poisoning by sodium fluoroacetate.
Structure
1799
See Fig. 83.1.
Synonyms Fluoroacetamide is also known as Compound 1081. Trade names for fluoroacetamide include Fuorakil®, Fussol®, Megarox®, and Yancock®. The CAS registry number is 640.19.7. Physical and Chemical Properties Fluoroacetamide has the empirical formula C2H4FNO and a molecular weight of 77 .06. It is a crystalline solid that sublimes on heating but melts at 107-109°C. It is very soluble in water, moderately soluble in acetone, and sparingly soluble in aliphatic and aromatic hydrocarbons. History, Formulations, and Uses At one times fluoroacetamide was used as a systemic inseCticide for scale insects, aphids, and mites on fruits; however, it has been considered too toxic to mammals for commercial use as an insecticide. Its use as a rodenticide was suggested by Chapman and Phillips in 1995. It is used as a bait (20 gm active ingredientlkg) in areas to which the public have no access, such as sewers and locked warehouses. It is formulated as dyed cereal-based bait which is mixed with water for use. 83.2.2.2 Toxicity in Laboratory Animals Basic Findings Fluoroacetamide is a compound of moderate to high acute toxicity depending on the species (see Table 83.2). In the WHO Recommended Classification of Pesticides by Hazards (World Health Organization, 1986), the technical material is listed in class IB, "Highly hazardous." The compound is absorbed by the skin (Phillips and Worden, 1957). Animals acutely poisoned by this compound show listlessness, irritability, chronic convulsions, abasia, piloerection, and irregular respiration (Araki, 1972). One characteristic usually observed in animals dying from acute poisoning with fluoroacetamide as well as with sodium fluoroacetate is postmortem rigidity (Bentley and Greaves, 1960). Death generally Table 83.2 Single-Dose LD 50 for Fluoroacetamide LD50 Species
Route
(mg/kg)
Reference
Rat
oral
15
Phillips and Worden (1957)
Rat
oral
I3
Bentley and Greaves (1960)
Rat
dermal
20a
Phillips and Worden (1957)
Mouse
oral
30.62
Araki (1972)
Mouse
subcutaneous
34.20
Araki (1972)
Mouse
intraperitoneal
85
Matsumura and O'Brien (1963)
Rabbit
oral
1.5-2.0
Rabbit
intravenous
0.25
Buckle et al. (1949)
83.2.2.1 Identity, Properties, and Uses
Chicken
oral
4.25
Egyed and Shlosberg (1977)
Chemical Name
aLowest lethal dose.
83.2.2 FLUOROACETAMIDE
2-Fluoroacetamide is the chemical name.
Phillips and Worden (1957)
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Rodenticides
occurs in coma after convulsions have stopped (Phillips and Worden, 1957). There is no obvious difference in susceptibility between the sexes (Bentley and Greaves, 1960). The time that elapses between dosing and the onset of convulsions appears to be related to the dosage level, and fiuoroacetamide affects behavior (Bentley and Greaves, 1960). Subacutely poisoned animals show anorexia, emaciation, and alopecia (Araki, 1972). Perhaps because of strain differences, investigations have reported slightly different thresholds for the largest repeated dosage tolerated by rats without clinical signs. As discussed later, the threshold for testicular injury is much lower. Phillips and Worden (1957) found that 3 mg/kg/day for 20 days was without effect on appetite or general health. Mazzanti et al. (1964) found similar results in rats on a dietary level of 50 ppm (about 2.5 mg/kg/day for 90 days). However, Steinberger and Sud (1970) reported that this same dietary level caused a reduction of food intake and of growth. The poisoning of farm animals by effiuent from a factory that manufactured fiuoroacetamide caused the Ministry of Agriculture, Fisheries, and Food to recommend that the compound should not be used as an insecticide in agriculture, for home gardens, or food storage in Great Britain, and it was withdrawn from the market (Allcroft and Jones, 1969; Allcroft et aI., 1969; Anonymous, 1964a, b). Absorption, Distribution, Metabolism, and Excretion Investigators agree that fiuoroacetamide is less toxic than fiuoroacetate. This has been attributed to the fact that metabolism of the former compared to the latter is slower (Matsumura and O'Brien, 1963). In fact, Phillips and Worden (1957) reported that they recovered, from the urine or rats receiving fiuoroacetamide at a rate of 3 mg/kg/day, 62% of the total intake unmetabolized, and they confirmed the identity of the compound by melting point and mixed melting point. This finding raises the possibility that the toxicity of fiuoroacetamide (albeit lower than that of fiuoroacetate) is in part inherent and does not depend entirely on metabolism to fiuoroacetamide on the rat testis, an effect apparently not reported for fiuoroacetate. Biochemical Effects Evidence that fiuoroacetamide has essentially the same mode of action as sodium fiuoroacetate is the finding that mammals poisoned by the amide contain greatly elevated levels of citrate (Allcroft et aI., 1969; Egyed and Brisk, 1965; Egyed and Miller, 1971; Egyed and Shlosberg, 1977; Matsumura and O'Brien, 1963). Further evidence is offered by the fact that cockroaches convert fiuoroacetamide to fiuoroacetate as well as to fiuorocitrate, and mouse amidase hydrolyzes fiuoroacetamide (Matsumura and O'Brien, 1963). Effects on Reproduction Selective destruction of the germinal epithelium of the testes of male rats apparently was reported first by Mazzanti et al. (1964), who studied only a single dosage level resulting from a dietary level of 50 ppm. On this diet, the body weight of 150-160 gm rats increased by 88% in 90 days but the testes were reduced to slightly less than one-third of the weight in controls. After 64 days, the tubules were almost
completely lacking in seminal cells; only some spermatogonia, the Sertoli cells, and the interstitial cells were apparently undamaged. Peculiar giant cells were observed. It was noted that fiuoroacetamide acts first on the more mature cells of the germinal epithelium and not on the cells where mitoses are more numerous. Dividing cells in the intestinal mucosa were undamaged. In a later study, male rats that received a dietary level of 50 ppm (usually calculated as about 2.5 mg/kg/day but said to be about 3.4 mg/kg/day in these rats) showed a marked morphological change in the nucleus of tep-13 spermatids within 24 hours, and the effects became more pronounced and the entire cell became distorted in 5 days. After 10 days of treatment, earlier-step spermatids showed degenerative changes and giant cell formation. Eventually, even spermatocytes were affected. Androgen secretion by the testis apparently was not affected. Dietary levels of 20, 10, and 5 ppm produced characteristic changes in late-stage spermatids but no effect on spermatocytes. The 5 ppm level had no effect on the weight of the testis of rats fed as long as 28 days, but higher levels led to a marked decrease in weight. Subcutaneous administration of fiuoroacetamide at a rate of about 1.0 mg/kg/day produced the characteristic change in stage-13 spermatids within 4 days and a 50% reduction in the weight of the testis in 28 days. Spermatogenesis continued, and spermatocytes and young spermatids remained apparently normal, but late spermatids were distinctly abnormal. Subcutaneous doses of about 0.2, 0.04, and 0.02 mg/kg/day produced little or no change in the weight of the testis and produced progressively less histological injury so that change was barely discernible at the lowest dosages. Thus, the effect of fiuoroacetamide on spermatogenesis is specific and not secondary to general toxicity, which (in the form of reduced growth) was evident only at the highest oral dosage (Steinberger and Sud, 1970). Testicular degeneration caused by fiuoroacetamide has been confirmed in rats and reported in other species (Egyed, 1973). Fluoroacetamide at an oral dosage of 15 mg/kg also interferes with reproduction in female mice, whether administered 2 days before or 10 days after fertilization; pregnancy was prolonged, prenatal mortality was increased, and the young suffered from cyanonis, respiratory distress, reduced growth, and decreased survival (Tokavera et aI., 1971). Effects on Wildlife and Nontarget Species In some countries, fiuoroacetamide is used to control field rodents, thus exposing nontarget species to either direct toxic effects by feeding on the baits or secondary effects by feeding on carcasses of rodents killed by the compound. Theses effects have been experimentally studied by Braverman (1979) on several nontarget species such as mongoose (Herpestes ichneumon), hyena (Hyaena hyaena), snakes, birds, cats, and dogs. This experiment showed a degree of susceptibility of the animals similar to that reported for sodium fiuoroacetate; it also confirmed that the dog was the most sensitive species. Some species showed a relative tolerance to direct poisoning; this was the case for barn owls, buzzards, and the black kite.
83.2 Fluoroacetic Acid and Its Derivatives
A secondary poisoning study was done by offering the carnivore carcasses of birds (Meriones tristrami) which had fed freely on poisoned grains. The results were quite variable; the mongoose was the most susceptible, whereas the risk of secondary poisoning to birds of prey was not high. An outbreak of poisoning by fluoroacetamide in four greylag geese (Anser anser) and teal (Anas crecca) has been reported in Israel (Shlosberg et aI., 1975). Clinical signs in one goose were described as severe convulsions, incoordinated twisting of the neck, total anemia, prostrating depression, and death. Treatment of Poisoning in Animals Sodium acetate did not protect rats poisoned by fluoroacetamide. However, when administered as a mixture by mouth at a ratio of 4:1 or 9:1, acetamide raised the LD 50 of fluoroacetamide from 15 to 22 mg/kg. When acetamide was administered by mouth at a dosage of 180 mg/kg within 65 min or less after fluoroacetamide at the otherwise fatal oral dosage of 20 mg/kg, all rats survived. The same was true when the ratio (9:1) remained the same, the delay did not exceed 60 min, and the dosage of poison was as high as 35 mg/kg. However, the antidote was ineffective when the delay was 105 min or greater (Phillips and Worden, 1956). The value of acetamide was confirmed by Hashida (1971). When administered to rats as a mixture by mouth, L-cysteine hydrochloride was antidotal, raising the LD 50 of fluoroacetamide from 15 to 25 and 30 mg/kg, respectively, at dosage ratios of 4:1 and 9:1 (Phillips and Worden, 1957). Acetamide at an oral dosage of 2500 mg/kg was also effective in treating chickens when given within 20 min after fluoroacetamide at a dosage of 10 mg/kg (slightly more than twice the LD 50 level). The same dosage given 30 min after the poison or 500 mg/kg given with the poison were ineffective (Egyed and Shlosberg, 1977). In limited tests, neither acetamide nor mono acetin was effective in treating poisoned sheep (Egyed, 1971). The ineffectiveness of sodium acetate and apparently of monoacetin and the effectiveness of acetamide an L-cysteine as antidotes for poisoning by f1uoroacetamide raise the possibility that the effective compounds do not prevent biochemical lesions directly but rather competitively retard the conversion of fluoroacetamide to fluoroacetate and thus permit more time for excretion of unmetabolized fluoroacetamide. 83.2.2.3 Toxicity to Humans Accidental and Intentional Poisoning At about 11 :30 hr, an 18-month-old girl removed a 120-ml bottle of 1% fluoroacetamide from a low drawer in the family kitchen and drank some of the contents. On the advice of a pharmacist, the child was given olive oil, the white of an egg, and milk at about noon and was made slightly sick. The child remained lively and played in the garden until her usual bedtime, 18:30. A about 23:30 that evening the child vomited but was put back to bed when she appeared all right. Apparently the child was not checked until 10:30 hr next morning, when she was found
1801
in a semiconscious state. On a physician's order, she was taken to hospital, but convulsions occurred on the way and the patient arrived about 11 :30 hr in a shocked state. The child was given about 10 ml acetamide in water once, 3.7 ml of brandy in water each hour, and symptomatic treatment. She continued to have occasional convulsions and remained unconscious until she died almost 96 hr after ingesting the poison. Both the heart and kidney contained 6.3 mg of organic fluoride per gram of dry tissue; the citrate content (108 ppm in heart and 23.9 ppm in kidney) was not considered significantly high. From the evidence available, it was estimated that the baby had consumed about 300 mg of fluoroacetamide or 23 mg/kg (Great Britain Ministry of Agriculture, Fisheries, and Food, 1961; WHO, 1963). Treatment of Poisoning Treatment of poisoning by fluoroacetamide should be the same as that for fluoroacetate (see Section 83.2.1.3) with due attention to removal of the poison and general care of the patient. Based on animal studies, rapid and energetic treatment with acetamide is recommended. A dosage of 315 mg/kg was effective in rats, but a much higher dosage was required in chickens. It is of special importance that the first dose be given in the earliest possible moment. Repeated administration was not used in the animal experiments but would appear wise. A combination of intravenous monoacetin (glyceryl monoacetate, 0.55 gmJkg) , sodium acetate (0.12 gmJkg), and ethanol (0.12 gmJkg) has also been recommended (Dipalma, 1981). Dipalma also suggested as an alternative course the oral administration of 100 ml of monoacetin plus 500 ml of water every hour for about 2 hr. Hemoperfusion involving fixed-bed uncoated charcoal was used in one case, but it was not helpful and the patient died (de Torrente et aI., 1979). 83.2.3 FLUOROETHANOL 83.2.3.1 Identity, Properties, and Uses Chemical Name Structure
2-Fluoroethanol is the chemical name.
See Fig. 83.1.
Physical and Chemical Properties Fluoroethanol has the empirical formula C2HsOF and a molecular weight of 64.07. It is a solid melting at approximately room temperature (26.5°C). It has a density of 1.091, a boiling point of 103°C and a flash point of 31°C. Use
It is used as a rodenticide.
83.2.3.2 Toxicity to Laboratory Animals Fluoroethanol is a compound of high acute toxicity, as indicated by an intraperitoneal LD 50 of 5 mg/kg in the rat (Bartlett, 1952). According to Bartlett, fluoroethanol is relatively inactive and its toxicity depends on its oxidation to fluoroacetate by tissued alcohol dehydrogenase.
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CHAPTER 83 Rodenticides
o
83.2.3.3 Toxicity to Humans
11
s
NH--C--NH
Three cases of poisoning of workers by fluoroethanol occurred in a chemical plant, in at least two instances as the results of accidental rupture of a container and rapid evaporation of the fluid. A typical patient suffered onset in about 90 min and was discharged from hospital in 4 days. All patients had tremor, severe muscular weakness, nausea, headache, and a slight swelling of the liver. (Hemorrhagic gingivitis in one patient and prediabetic hyperglycemia in another were explained by their past histories and were unrelated to poisoning.) Examination of the other 40 workers in the plant failed to reveal any complaints or clinical finding that could be related to the compound (Colamussi et aI., 1970). There is no specific treatment for subacute poisoning except, of course, complete cessation of exposure. If acute poisoning should occur, it should be treated like poisoning by fluoroacetate (see Section 83.2.1.3). This compound does not seem to be marketed any longer.
69 N0 2
11
~-C--NH'
~ ANTU
Pyriminil
o 11
CH -N-C-NH 3
I ON
H
Alloxan Slreplozocin
83.3 SUBSTITUTED UREAS One of the compounds that has been promoted as a rodenticide relatively safe for other mammals is pyriminil, as substituted urea (see Fig. 83.2). It is not clear whether this group of compounds has been explored extensively with a view to selecting the one with the best combination of effectiveness for killing rodents and safety for humans and useful animals. However, it has become apparent that pyriminil and some other substituted ureas are specific poisons for the f3 cells on the pancreas and, therefore, cause diabetes mellitus. This effect may not be related to the mode of action of pyrimidil as a rodenticide, but it has great bearing on the overall safety of the material. 83.3.1 PYRIMINIL 83.3.1.1 Identity, Properties, and Uses
Chemical Name N-(3-pyridylmethyl)-N'-(4-nitrophenyl)urea is the chemical name. Structure See Fig. 83.2. Synonyms Pyriminil is also known as PNU, pyrinuron, and RH-787. It was sold under the trade names Vacor®, Rat Killer®, DLP-787 20% bait, and DLP-787 10% House Mouse Tracking Powder. The CAS registry number is 53558-25-1. Physical and Chemical Properties Pyriminil has the empiric formula C13H12N403 and a molecular weight of 272.27. It decomposes at 223°C. History, Formulations, and Uses Pyriminil was introduced in 1975 and developed as an acute rodenticide. It was used to control Norway rats, roof rats, and house mice; it was
Figure 83.2 ANTU, a thiourea rodenticide, and three substituted ureas known to cause diabetes in one or more species.
especially effective against rodents resistant to anticoagulant poisons. Pyriminil was sold for indoor use only as a prepared bait containing 2% active ingredient and a 10% tracking powder. The product was withdrawn from the market by the U.S. manufacturer in 1979 (Chappelka, 1980), but it is still manufactured on a small scale for local use-in the People's Republic of China, for example. 83.3.1.2 Toxicity in Laboratory Animals There are greater differences in the susceptibility of different species to pyriminil (technical material) as shown in Table 83.3 (Peardon, 1974). The marked susceptibility of Norway rats was of course the basis for its use as a rodenticide. Cats also are very susceptible. Apparently, a good description of the signs of acute poisoning in laboratory animals has not been published. A simple list of signs and symptoms in dogs has been given in the distribution company's technical bulletin. The onset of the symptoms may be delayed 4-48 hr. They include nausea and emesis, depression, initial constriction of pupils followed later by dilated pupils and visual impairment with slow pupillary response to light, ataxia, fine to coarse tremors, hind-limb weakness, decreases reflexes, deep breathing, and dehydration. Similar symptoms have been reported in a horse that had eaten at least 250,000 mg (about 250 mg/kg). The animal showed severe muscular fasciculations, dilated pupils, and profuse sweating within 24 hr after ingestion. Laboratory tests revealed severe hyperglycemia (418 mg/lOO ml) and indications of liver injury (elevated liver enzymes). The animal was treated with intravenous nicotinic acid (2.2 mg/kg) followed by four subsequent injections of 1 gm and recovered; it was considered clinically normal 3 months later. Three other poisoned horses showed the
83.3 Substituted Ureas Table 83.3 Single-Dose LD 50 for Pyrimidil in Various Species Oral LD 50 Species
Sex
(mg/kg)
Albino rat
M
Norway rat
M
Roof rat
M
Cotton rat
M,F
12.30 4.75 18.00 20-60 84 98 10-20 30-100 300 500 62 2000-4000 500 205 710 1780
Albino mouse
M
House mouse
M
Deer mouse
M
Guinea pig
M
Rabbit
M
Dog
M
Cat
M,F
Rhesus monkey
M,F
Pig
M
Vole
M,F
Chicken
M
Pigeon
M,F
same signs as well as intense abdominal pain, hind-limb weakness, ataxia, and persistent inappetence (Russell et aI., 1978). Peoples and Maddy (1979) have reported without details poisoning in domestic animals (two horses, three cats, and 17 dogs) in California. The case of a 22-kg dog seen eating a full 30-gm packet of Vacor (780 mg active ingredient) is mentioned. Immediately following ingestion, the dog vomited but became blind 2 days later.
Absorption, Distribution, Metabolism, and Excretion Pyriminil is rapidly absorbed by rats, mice, and dogs after oral administration. Blood levels peaked in 1-6 hr, depending on species and site of the radiolabel 14c. Gastrointestinal transit of 14C is more rapid in dogs than in rats. Urinary and fecal excretions are of similar importance in all three species. Tissue distribution of two 14C labels (nitrophenyl and pyridyl) varied, especially in dogs. The liver contained mor of the dose than any other single organ (Deckert et aI., 1978a). Rats tolerated, metabolized, and eliminated single or multiple sublethal dosages (5 mg/kg) but were less efficient than dogs in detoxifying dosages in excess of 20 mg/kg. It was concluded that the tolerance of dogs for the compound depended on their efficient hepatic extraction, metabolism, and excretion of it (Deckert et al.,1978b). Several metabolites of pyriminil have been identified Deckert et al. (1978a, 1979). These include aminopyriminil, p-aminophenyl urea, p-acetamidophenyl urea, p-nitroaniline, p-phenylenediamine, p-acetamidoaniline, nicotinic acid, nicotinuric acid, and nicotinamide. The concentrations of these metabolites varied from one species to another. The presence of the parent compound in rat and human urine suggests that they may be more sensitive to the compound than the dog because of less efficient metabolism (Deckert et aI., 1979).
1803
Biochemical Effects Repeated, sublethal doses of pyriminil increased the urinary and fecal excretion of a later dose of the compound tagged with 14C; however, the same animals showed increased hexobarbital sleeping time and other evidence of inhibition of certain liver microsomal enzymes, especially p-nitro-anisole O-demethylase. Whatever microsomal enzymes are responsible for metabolism of pyriminil are induced by pretreatment with 3-methykholanthrene, which increases the biliary excretion of the metabolites and decreases pyriminil toxicity 50-fold Deckert et al. (1977, 1978a). Mild pyriminil-induced hyperglycemia was observed in rats; it was also shown to be reversible by insulin (Deckert et aI., 1977). The diabetogenic effects of pyriminil were also confirmed in patients poisoned by the product. This effect is the result of a direct toxic action of the f3 cells of the pancreas. Wilson and Gaines (1983) have demonstrated that pyriminil at concentrations ranging from 10- 2 to 10- 5 M preferentially intoxicates rat pancreatic f3 cells in culture, whithin 1 hr of contact. It was also shown in the study that nicotinamide can reduce pyriminil-induced f3 cell injury, thus confirming previous findings by Karam et al. (1980) that nicotinamide could partially reverse pyriminil inhibition of glucose-stimulated insulin secretion by freshly isolated islets of Langershans from the rat. In addition to its diabetogenic effect, pyriminil has a direct effect on glucose metabolism. The erythrocytes of patients poisoned by the compound showed a marked depression of glucose consumption as well as decreased uptake of methylene blue in the presence of glucose. In addition, a 0.1 mM concentration in vitro caused decreased utilization of glucose and decreased uptake of methylene blue by erythrocytes from normal people and rabbits (Lee and Lee, 1977). The mechanism of action has been investigated and it was shown that pyriminil specifically inhibits the NADH:ubiquinone reductase activity of complex I in mammalian mitochondria. The activity of other respiratory enzymes of mitochonrai is unaffected at concentrations that completely inhibit the redox and energetic function of complex I. Inhibition of complex I activity quantitatively correlates with the inhibition of insulin release in insulinoma cells and pancreatic islets and is also consistent with the doses reported in cases of human poisoning. These results indicate that the toxic and diabetogenic action of pyriminil primarily derives from the inhibition of mitochondrial respiration of NAD-linked substrates in the high energy demanding pancreatic islets (Esposi et aI., 1996). Treatment of Poisoning in Animals The mechanism of action of pyriminil remains uncertain, but it is of interest that alloxan, streptozotocin, and dithizone, all which can induce diabetes mellitus in experimental animals, are substituted ureas. However, it would appear that some species such as dogs, cats, and laboratory primates are refractory to both the diabetogenic and neurotoxic effects of pyriminil (Karam et aI., 1980). Whereas 6-aminonicotinamide is not a substitute urea, it is toxic to f3 cells and it is recognized antagonist of nicotinamide (Herken, 1971). Because nicotinamide can prevent the toxic effect of streptozotocin (Ganda et aI., 1976), alloxan (Rossini et
1804
CHAPTER 83
Rodenticides
aI., 1975), and N-3-pyridylmethl-N'-4nitrophenyl urea (Deckert et aI., 1977), it seems possible that all of these compounds act as nicotinamide antagonists.
83.3.1.3 Toxicity to Humans Accidental and Intentional Poisoning There are many reports on human poisoning in the literature describing the main clinical and laboratory features of those poisonings. A 25-year-old man with a history of psychiatric disturbances attempted suicide by injecting an unknown amount of pulverized methaqualone tablets and ingesting two packets of rat poison, each containing 737 mg of pyriminil. Seven days later he was admitted to a local hospital for treatment of a staphylococal abscess of the left antecubital fossa. He received antibiotics and had rapid clinical improvement. It was recorded that since attempting suicide the patient had noticed lassitude, anorexia, abdominal bloating, constipation, and the onset of painful paresthesia with numbness of this legs and difficulty in walking. A random plasma glucose level on admission was 309 mg/l 00 ml, and check samples taken on subsequent days were slightly higher. Ketones and glucose were present in the urine. On the fourth hospital day, insulin therapy was started. The diabetes gradually was controlled, although tolerance for carbohydrate and need for insulin were erratic. An upper gastrointestinal tract series done on the tenth hospital day showed gastric and proximal small bowel hypomobility bordering on atony. The patient was discharged on hospital day 19 on a regimen of insulin and temporary thoridiazine. The patient remained well for 16 days and then returned because of nausea and vomiting. He was found to have severe autonomic and peripheral polyneuropathy characterized by orthostatic hypotension, greatly diminished response to pinprick and vibratory sensation in the lower extremities, and other changes. Although the diabetes was now better controlled, the serum sodium was low 016 mEq/liter), and the syndrome of inappropriate antidiuretic hormone was demonstrated. The hyponatremia responded to fluid restriction, and the orthostatic hypotension was improved by support stockings. Ten months after the suicide attempt, the patient experienced two episodes of weakness and lethargy that were relieved by eating. He had lost about 18 kg and appeared cachetic (45 kg, 174 cm) but alert and well oriented. His gait was ataxic and there was substantial muscle wasting. A very thorough examination showed reduced disappearance rate of intravenous glucose and depressed C-peptide response to intravenous glucose when compared with a normal control but no impairment of glucagon release after stimulation by intravenous arginine. Nerve conduction studies demonstrated severe sensory and mild motor neuropathy. Quadricep capillary basement membrane thickness was in the diabetic range. Insulin was discontinued and tolbutamide prescribed. Following discharge, the patient regained 5 kg and experienced subjective improvement of his neuropathy (Prosser and Karam, 1978). Whereas most clinical studies have place greatest emphasis on the diabetogenic action of pyriminil, its injury to the nervous system was no less remarkable, as emphasized in a paper
by LeWitt (1980a). This injury often involved autonomic impairment (postural hypotension often severe enough to cause fainting when the patient sat up, impaired pupillary responses, impotence, decreased sweating, urinary retention, dysphagia, and gastrointestinal hypomobility), peripheral neuropathy (loss of muscle-stretch reflexes, sensory loss, neurogenic myopathy), and encephalopathic and dyskinetic features (loss of cortical function ranging from confusion to coma, cerebellar ataxia, tremor, motor hyperactivity, nystagmus, and diffuse electroencepha10graphic changes). In addition, some cases involved chest or epigastric pain and some showed ischemic electrocardiographic changes. Cardiac arrhythmias were occasionnally the cause of death. Neurological disorders often appeared within hours after ingestion. Occasionally, onset was delayed or insidious. Symptoms related to different parts of the nervous system began and later improved at different times in the same patient, and the order of progression varied from case to case. Neurological improvement took many months, and full recovery was uncommon, orthostatic hypotension in particular tending to persist. Causes of delayed death included inanition, sepsis, aspiration pneumonia, and insulin-induced hypoglycemia. Accidental ingestion of pyriminil by a 25-month-old boy resulted in acute vomiting, lethargy, seizures, hypoglycemia (followed by hypreg1ycemia and glucose intolerance), and autonomic and peripheral neuropathy (John son et aI., 1980). A review of reports unpublished in 1978 indicated 7 deaths and 2 nonfatal cases in Korea and 4 fatal and 11 nonfatal cases in the United States. At least in the United States, all the cases were in adults; all but one were attempted suicide; all the survivors developed diabetes mellitus and autonomic nervous system dysfunction, chiefly dysphasia, dystonia, and bowel and bladder dysfunction. Hypothermia and paresthesias were seen. A later review revealed nearly 90 cases in the United States and over 250 in Korea (Frethold et al., 1980). A case of acute poisoning (approximately 67 mg/kg) in a 42-year-old man with all the signs already described but characterized by a severe orthostatic hypotension with full spontaneous recovery 11 months after hospitalization was reported by Osterman et al. (1981). Gallanosa et al. (981) have compared the main features of four cases reported with enough details in the literature with those of one case of their own. Dosage Response A dose as low as 780 mg was fatal within 150 days. A dose of 2340 mg was fatal within 1 day, but a patient survived 40 days after ingesting 7020 mg. One patient survived 2340 mg, and at least two survived 1560 mg but not without characteristic, persistent illness. The smallest dose known to have produced characteristic illness was 390 mg (about 5.6 mg/kg) (LeWitt, 1980b). Laboratory Findings The most important findings for guiding treatment and often for diagnosis include nearly transient hypoglycemia followed by persistent hyperglycemia, glycosuria, ketosis, and elevation of serum amylase and lipase activities. p-Nitroaniline at a concentration of 5.1 ppm has been
83.4 Thioureas
reported in the liver of a person who died after accidentally ingesting pyriminil (Osteryoung et al., 1977). In the case of a 7-year-old boy who was found dead a day after another child saw him ingest a packet of pyriminil, unchanged compound at a concentration of 1.5 ppm was found in the urine hydrolysate and two metabolites were found in the liver and some other samples. Aminopyriminil (nitro group metabolized to an amine) was found at concentrations of 5.6, 104,0.3, and 0.6 ppm in liver, kidney, spleen, and urine, respectively. Acetamidopyriminil (amino group conjugated with acetic acid) was found in traces in the blood and liver (Frethold et al., 1980). Karam et al. (1980) reported (in addition to the clinical features) autopsy findings from several cases of acute poisoning, including that of a 7-year-old boy. All three cases showed extensive islet degeneration of the pancreatic tissue with generalized destruction of fJ cells and sparing of IX and ~ cells as well as of the exocrine glandular tissue. Islet-cell surface antibodies were detected in four of the six reported cases. It may be that these antibodies are the result rather than the cause of fJ-cell destruction. Pathology Loss of fJ cells of the pancreas has been observed generally in persons killed by pyriminil (Frethold et aI., 1980; Karam et aI., 1980; LeWitt, 1980a; Prosser and Karam, 1978). Lesions of the nervous system have not been found so regularly. In one case reported by Le Witt (1980a), no lesions of the central or peripheral nervous system were found; in another case, cerebral edema and neuropathic changes restricted to the sensory spinal roots were found. Autopsy of a 39-year-old man who survived 19 days revealed (a) severe loss of ganglion cells and rare degenerating neurons in the paravertebral sympathetic ganglia, Cb) marked loss of neurons in the sensory spinal ganglia with multiple residual nodules of Nageotte, (c) marked degeneration of the sensory roots and posterior columns, (d) slight perivascular lymphatic infiltrates in both the sympathetic and sensory ganglia, (e) swelling of nerve fibers and thinning of the myelin sheaths of the sural nerve, and (f) isolated degenerated and regenerating fibers in the skeletal muscles (Papasozomenos, 1980). Treatment of Poisoning Patients who develop diabetes mellitus clearly must be treated for that condition in the usual way. There is good reason from animal experiments to believe that diabetes could be prevented if the patient were given large, repeated doses of nicotinamide beginning promptly after ingestion of the poison. However, cases have ended in diabetes and neuropathy when nicotinamide was started 9 and 14 hr, respectively, after ingestion. Nicotinamide was considered possibly beneficial in the case of an infant, even through administration was started something over 12 hr after ingestion of pyriminil (Johnson et al., 1980). However, the fact that the child received the poison "on a piece of gum" offered by another child suggests that the initial dose was small, and complete recovery may have been due to that fact alone (Pont et aI., 1979). The dose and duration of the treatment with nicotinamide are still uncertain (Anonymous, 1979). Nicotinic acid has also
1805
been tried as an antidote (Pont et aI., 1979), but its use is contraindicated because (a) it is toxic in humans, (b) it protects animals only against alloxan and not streptozotocin (Ganda et al., 1976), and (c) its vasodilatory effects may complicate the control of blood pressure. Cases that require insulin may progress so that insulin is no longer required but the patient can be maintained on sulfonylureas.
83.4 THIOUREAS The development of ANTU as a poison for adult Norway rats was described by Richter (1945). The entire development was a result of a chance observation associated with studying the taste of phenylthiourea, which is bitter to most people but tasteless to a few who inherit this specific lack of sensation as a Mendelian recessive trait. When an attempt was made to explore this taste difference in animals, it was found that if a few crystals were placed on the tongues of rats, all of them died overnight. The wide and prolonged use of phenylthiourea for taste and inheritance tests without any untoward effect indicated its safety for humans, whereas the results in rats suggested that it might serve as a rat poison. Further study revealed that rats detect and reject phenylthiourea too effectively for it to be practical as poison. This led to a systematic search of other thiourea derivatives with high toxicity but little or no taste. All monosubstituted thiourea derivatives tested produced pulmonary edema and pleural effusion in the laboratory rat (Dieke et al., 1947). The toxicity of thiourea to wild Norway rats was enhanced by a single aromatic substitution on one of the nitrogen atoms. Two or more substitutions on one or both nitrogen atoms lowered the toxicity, as was also true of substitution on the sulfur atom. ANTU was chosen as the most suitable compound. Although the dog is susceptible to ANTU, most animals, including monkeys, are resistant. This offered the hope that humans would be resistant also, and extensive field trials in areas of Baltimore led to no toxic symptoms either in workers or in the over 500,000 persons living in the treated areas (Richter, 1945). A disadvantage of ANTU as a rodenticide is that young Norway rats and roof rats of all ages are too resistant to the compound for it to be practical for their control. Another disadvantage is the prompt appearance of both tolerance and bait refusal in adult Norway rats that have received a nonfatal dose. Tolerance is completely lost within 30 days, but refusal may last longer (Richter (1945, 1946). Gaines and Hayes (1952) found that bait shyness lasted at least 4 months under field conditions. Several interesting observations were made during the survey of thioureas. All of these compounds produce hyperplasia of the thyroid gland. Whereas nonlethal doses of unsubstituted thiourea have little effect on pigmentation or hair growth, phenylthiourea destroys pigment both in the skin and in the hair but without affecting growth of the hair, and ANTU completely stops pigment production and growth of hair. Withdrawal of the substituted thioureas is followed in less than 10 days by
1806
CHAPTER 83 Rodenticides
recovery of pigment and hair growth. Finally, different strains of Norway rats on different diets showed thiourea LD 50 values as different as 4 and 1830 mg/kg. The difference was modified but not eliminated by placing the rats on the same diet as that of the most susceptible ones. Age differences in the susceptibility of Norway rats to thiourea are similar to those with ANTU (Dieke and Richter, 1945; Richter, 1945). 83.4.1 ANTU 83.4.1.1 Identity, Properties, and Uses Chemical name name.
1-(1-Naphthyl)-2-thiourea is the chemical
ANTU forms colorless crystals and the technical grade is a gray crystalline powder with a bitter taste. Its melting point is 198°C (pure). Its solubility in water at 25°C is 0.06 gm/lOO ml; in acetone, 2.43 gm/lOO ml; and in triethylene glycol, 8.6 gm/lOO ml (technical). History, Formulations, and Uses ANTU was discovered as a rodenticide in 1945. The formulations include baits (1030 gm/kg) and tracking powders (200 gm/kg). It is used specifically against the Norway rat. In some countries it has been withdrawn from use because of the carcinogenicity of ,B-naphthylamines present as impurities (Worthing and Walker, 1983). 83.4.1.2 Toxicity in Laboratory Animals
Structure
See Fig. 83.2.
Synonyms ANTU, an acronym for a-naphthylthiourea, is the approved common name (BSI, ISO) for this compound. Trade names include Anturat®, Bantu®, Kill Kantz®, Krysid®, Rattrak®, and Rat-tu®. Code designations for ANTU include Chemical-109 and U-5227. The CAS registry number is 86-88-4.
Basic Findings Different investigators have been in good agreement about the acute oral toxicity of ANTU to Norway rats. Dieke and Richter (1945) and Lehman (1951,1952) found oral LD 50 values of 6.9 and 6 mg/kg, respectively. There is a wide variation in susceptibility among different species, especially to intraperitoneal administration (see Table 83.4). The Norway rat is particularly susceptible, the young being slightly more resistant than the adults.
Physical and Chemical Properties ANTU has the empirical formula CllHION2S and a molecular weight of 202.27. Pure
Absorption, Distribution, Metabolism, and Excretion Early studies on the metabolism of phenylthiourea and of
Table 83.4 Single-Dose LD 50 for ANTUa LD 50 Species
Route
Rat
intraperitoneal
10
Rat
intraperitoneal
7
Lisella et al. (1971)
Rat
intraperitoneal
5
DuBois eta!' (1947)
(mg/kg)
Norway, domestic r
intraperitoneal
2.5
Norway, domestic II
intraperitoneal
6.25
Norway, wild, adult
intraperitoneal
6.20-8.10
Norway, wild, young
intraperitoneal
16-58
Alexandrine
intraperitoneal
250
Norway, wild
oral
6.9
Mouse
intraperitoneal
56
Rabbit
intraperitoneal
400
Guinea pig
intraperitoneal
140
Guinea pig
intraperitoneal
350
Dog
intraperitoneal
16
Dog
dermal
38
Cat
oral
Monkey
intraperitoneal
Monkey
oral
4250
Chicken
intraperitoneal
2500
Chicken
oral
4250
aFrom rARC (1983).
500 175
Reference Boyd and Neal (1976)
DuBois et al. (1947)
83.4 Thioureas diphenylthiourea suggest the basis of the toxicity of monosubstituted thioureas. It was shown by Dieke et al. (1947) that the oral LD 50 of the phenyl compound is 8.6 mg/kg, whereas a dosage of 2000 J?g/kg of the diphenyl compound did not produce illness. Both rats and rabbits excrete little phenylthiourea as compounds with the -C=S group intact (Carroll and Noble, 1949; Williams, 1959). In rabbits, the proportion of such compounds was only about 12%, but the corresponding proportion was about 70-80% for diphenylthiourea. These observations suggest that toxicity was associated with desulfuration in vivo (Williams, 1959). It has been speculated that ANTU acts on the lung by the release of hydrogen sulfide (Petit et aI., 1970), but this seems highly unlikely because rats rendered tolerant to ANTU are not tolerant to hydrogen sulfide (or carboxyl sulfide or phosgene) (Carroll and Noble, 1949).
Biochemical Effects By using a mixture of 3SS_ and 14C_ labeled ANTU, it was possible to show that some of the sulfur and a smaller proportion of the carbon were covalently bound to macromolecules of the lung and liver following in vivo administration. By contrast, practically no radioactive carbon was bound when an equal amount of the almost nontoxic, 14C_ labeled oxygen analog of ANTU (a-e 4C]naphthylurea) was administered. In the presence of NADPH, ANTU was metabolized by either lung or liver microsomes in vitro in such a way that the rates of binding of 3SS or 14C to macromolecules of the microsomes were greater than those associated with boiled microsomes or with normal microsomes without NADPH. Binding in the presence of active enzyme and NADPH was covalent and accompanied by a decrease in the level of cytochrome P-450 detectable as its carbon monoxide complex. Pretreatment of rats at the rate of 2 mg/kg/day for 5 days produced a decrease of their microsomal enzyme activity as measured by metabolism of parathion. All such pretreated rats survived a dosage of ANTU (10 mg/kg) which killed 6 to 10 controls, and binding of 3SS by proteins of the liver and especially the lungs of the pretreated animals was less than that of the controls. Pretreatment with 4-ipomeanol protected all rats from an otherwise uniformly fatal dose of eSS]ANTU and cause a slight reduction of covalent binding of 3SS to lung (but not liver) proteins. Finally, rats pretreated with dimethylmaleate, which depletes tissue stores of glutathione, were killed by ANTU at 5 mg/kg, a dosage which was harmless to controls. In every instance, rats killed by ANTU showed a hydrothorax of at least 4 ml, whereas those protected by pretreatment with ANTU or ipomeanol developed no hydrothorax. These findings were interpretated as evidence that (a) the toxicity of ANTU depends on metabolic activation and on covalent binding of the reactant(s) to lung macromolecules and (b) tolerance to ANTU is the result of inhibition of microsomal enzymes and consequent reduction in the metabolic activation of a challenge dose (Boyd and Neal, 1976). An extension of this reasoning would attribute the normal tolerance of young Norway rats to ANTU to their relative lack of microsomal enzyme activity. Further study showed that about half of the atomic sulfur released from ANTU reacted with cysteine side chains of mi-
1807
crosomal protein to form a hydrodisulfide. The other moiety released by micro soma enzymes is a-naphthylurea (Lee et aI., 1960).
Effects on Organs and Tissues ANTU induced reverse mutations in Salmonella typhimurium strain TA1538 in the presence but not in the absence of Arocolor- or phenobarbitalinduced rat liver microsomal preparations. A preparation purified by thin-layer chromatography was as active as a technical grade material, thus excluding the attribution of activity to impurities. ANTU also transformed Syrian hamster embryo cells in vitro without the addition of an activating system (Kawalec et aI., 1979). ANTU was tested for carcinogenicity (Fitzburg and Nelson, 1947) in mice (Innes et aI., 1969) by administration in the diet. No tumor was reported in either study, but the International Agency for Research of Cancer (IARC) (1983) found that both studies were inadequate to evaluate the carcinogenicity of ANTU to experimental animals. Pathology As far as the rat lung is concerned, ANTU causes marked edema of the subepithelial spaces of the alveolar walls without erosion or other damage to type I and type 11 epithelial cells. Thus, edema caused by ANTU differs morphologically and presumably in mechanism from that produced by injection of epinephrine or by an injection of a mixture of fibrinogen and thrombin into the cerebrospinal cistern (Hatakeyama and Shigei, 1971). The edema caused by intraperitoneal ANTU in rats is dosage-related in the range of 3-50 mg/kg. Although interstitial edema was the first observable change, bleeding and scalloping of endothelial cells were observed within 2 hr, and epithelial damage was apparent electron microscopically within 6 hr following 50 mg/kg. The injury was apparently similar to but more rapid than that caused by 99% oxygen at 1 atmosphere pressure (Meyrick et aI., 1972). Not only pulmonary edema but also pleural effusion shows a dosage-response relationship (Sobonya and Kleinerman, 1973). Using a different approach, Bohm (1973) demonstrated changes which he interpreted as indicating increased permeability to colloidal carbon in the pulmonary arteriodes as well as capillaries and venules of rats within 3.25 hr after an intraperitoneal injection of ANTU at a rate of 10 mg/kg. In anesthetized sheep given 20, 50, 75, or 100 mg/kg ANTU intravenously, the first phase of the response consisted of transient increases in pulmonary artery pressure and plasma and lymph thromboxane B2 concentrations. These changes were not dependent on the dose of ANTU administered. At 2-4 hr after administration, pulmonary artery pressure and thromboxane concentrations were normal or near normal. ANTU produces a two-phase response with the steady state characterized by a dose-dependent increase in lung microvascular permeability (Havill et aI., 1982). These authors, on the basis of experimental results in sheep, suggest that the severe pulmonary hypertension that follows ANTU administration may be mediated by vasoconstrictor products of arachidonic acid metabolism and
1808
CHAPTER 83
Rodenticides
that the complement or coagulation systems may be involved as well, resulting in pulmonary microemboli. O'Brien et al. (1985) have reported that isolated lungs from rats treated 4 hr earlier with ANTU had decreased conversion of angiotension I to angiotension 11 and that the extent of decrease was related to the dose of ANTU administered and to the perfusate flow rate. It may be that permeability of membranes of the kidney as well as those of the lung and pleura is increased inasmuch as urinary excretion of albumin occurs (Patil and Radhakrishnamurty, 1977). Treatment of Poisoning in Animals Mortality of rats caused by 5 mg/kg of ANTU was reduced when allylthiourea, isopropylthiourea, ethylenethiourea, or ethylidenethiourea was administered simultaneously with or a very short time after the ANTU. The first two compounds reduced the survival time of rats that died, but the last two compounds slightly prolonged it (Meyer and Saunders, 1949). Although the reduction in mortality was statistically significant, the degree of protection was small. Furthermore, these results in the rat may be more closely related to the phenomenon of tolerance than to antidotal action in the usual sense. In any event, Carroll and Noble (1949) found that tolerance to phenylthiourea and ANTU could be produced not only by small dosages of the compounds themselves but also by a number of related and some apparently unrelated compounds. The ability of the effective thiourealike substances to confer protection was unrelated to their acute toxicities ot antithyroid activities. Protected rats failed to develop pulmonary edema or pleural effusion following dosages of toxic thioureas lethal to untreated rats. Thyroidectomized rats could be made tolerant as readily as intact rats. Following a large dose, phenylthiourea was excreted in the urine of a tolerant rat in sufficient quantity to kill a normal animal. 83.4.1.3 Toxicity to Humans Accidental and Intentional Poisoning The absence of a report of uncomplicated poisoning is noteworthy in view of the extensive use of ANTU in Baltimore and some other places and the fact that an occasional bait must have been eaten by children. Several series of cases were reported from France, where chloralose was used either alone for killing crows or rats or in combination with ANTU for killing rats. In one series of 22 cases, all showed some degree of coma and motor agitation, both characteristic of chloralose poisoning; however, more intense pulmonary symptoms were present where ANTU was involved. The low toxicity of ANTU for humans is indicated by the fact that all the patients recovered, although all had ingested the poison with suicidal intent and, therefore, in relatively high dosages (Tempe and Kurtz, 1972). In another series of cases involving chloralose, 14 involved ANTU also, 1 involved chloralose only, and the presence of or absence of ANTU was not established in the remainder. In addition to the respiratory difficulty that may be present with any coma, 11 of
the 14 persons poisoned by a combination of chloralose and ANTU required intubation mainly because of tracheobronchial hypersecretion, and 9 of them required artificial respiration. All survived (Favarel-Garrigues and Boget, 1968). The authors characterized the beginning of tracheobronchial hypersecretion as a secretory storm that started early and sometimes suddenly. The secretion was a white froth that, unlike edema fluid, was not sticky or high in protein. The mildness of X-ray changes contrasted with the clinical gravity of the situation. Oxygenation of the blood was always more nearly complete than in acute pulmonary edema. The hypersecretion disappeared rapidly, often in less than an hour. Apparently, if patients poisoned by a combination of chloralose and ANTU are treated properly, their illness is no more protracted than in poisoning by chloralose alone (see Section 83.1). Use Experience Laubstein (1962) reported a case of eczema that he attributed to occupational exposure to ANTU. On the basis that ,B-naphthylamine is an impurity in ANTU, Case (1966a) raised the possibility that persons who distribute ANTU may be in danger of bladder cancer. No epidemiological evidence was offered. Later, Case (1966b) mentioned that an investigation of the occupational history of two rodent operators who were suffering from bladder tumors had revealed that different batches of ANTU differed in the degree of contamination with naphthylamine, and some of the contaminant was ,B-naphthylamine. As a result, the Ministry of Agriculture, Fisheries, and Food recommended in May 1966 that the use of ANTU be restricted to professional operators, and in November 1966 an advisory committee recommended that use of the compound stop until their investigation was complete (Anonymous, 1966). In 1982, Davis et al. reported 14 cases of urothelial tumors observed among 51 rodent operatives exposed to ANTU in the United Kingdom between 1961 and 1980. In the United States as a whole, the age-adjusted death rate for cancer of the bladder increased from 3.1 to 4.1 per 100,000 from 1931 to 1945, when ANTU was discovered, and it continued to increase more slowly until 1953, when it reached 4.4. Since 1953 the values have varied around slightly over 4.3 as a mean. The declining increase in rate from 1945 to 1953 occupied a period less than the average latent period for cancer of the bladder among men with heavy exposure to naphthylamine used in the manufacture of dyes. Thus there is no evidence for any carcinogenic action of ANTU in the general population. Because ANTU had been used so extensively in Baltimore, as described by Richter and his colleagues, the matter was investigated here. Because of the wide fluctuations in rates based on small frequencies, it was not practical to compare death rates for single years; therefore, the data were combined for 3-year periods. For 1949-1965, the rate per 100,000 population varied from the earliest value of 4.4 to 2.9 with no definite trend but certainly with no increase. Dosage Response The threshold limit value is 0.3 mg/m 3 of air over an 8-hr work shift (OSHA standard).
1809
83.5 Anti-vitamin K Compounds
Treatment of Poisoning have to be symptomatic.
If treatments were required, it would
OH
CJCl o
83.5 ANTI-VITAMIN K COMPOUNDS As reviewed by Link (1944, 1959), knowledge of the antivitamin K compounds began not with vitamin K but with hemorrhagic disease of cattle, which was first recognized in the 1920s on the prairies of North Dakota and neighboring Alberta. It was found that the condition was not caused by a microorganism or a nutritional deficiency but was associated with sweet clover that had gone bad. Hence the condition was known as "sweet clover poisoning." When cattle or sheep had improperly cured hay made from the common varieties of sweet clover (Melilotus spp.) as their only food, the clotting power of their blood decreased in about 15 days and they often died of internal hemorrhage in 30-50 days. If the disease had not progressed too far, it could be reversed by substituting good hay or by transfusion of blood freshly drawn from normal cattle. Link first learned of the problem in December 1932. During the following February, a Wisconsin farmer came to his laboratory with a dead heifer, a milk can containing blood with no power to clot, 100 pounds of spoiled sweet clover and the all too common, tragic story of cattle dying on an isolated farm. In Link's laboratory, a practical bioassay for hemorrhagic effect was developed. It was not until June 1939 that the active poison was isolated and crystallized. Using improved methods of isolation developed after the identity of the compound was known, it was shown that the compound was present in spoiled hay at a concentration of about 60 ppm. The structure was shown to be 3,3' -methylene-bis(4-hydroxycoumarin), later known as dicoumarol or by the trade name Dicumarol®, and it was synthetized in April 1940. The biological synthesis during spoilage of the hay can be rationalized as an oxidation of coumarin (the compound responsible for the characteristic sweet smell and bitter taste of sweet clover) and the subsequent condensation of two molecules of 4-hydroxycoumarin with formaldehyde (see Fig. 83.3). When synthetic dicoumarol became available in quantity, the essentials of its pharmacological action were established quickly. Between 1940 and 1942, it was rapidly adopted for treatment of thromboembolic disease in humans. About 50 clinical reports were published between 1941 and 1944. In 1942, Link himself set up field trials to test the suitability of dicoumarol as a rat poison. Tests by O'Connor (1948) using a concentration of 0.44 mg/g were reported as highly successful. However, tests carried out by the U.S. Public Health Service (Hayes and Gaines, 1950) led to the same conclusion as those of Link: that dicoumarol was impractical as a rat poison. While the medical and possible rodenticidal uses of dicoumarol were being explored, over 100 analogs of the compounds were synthesized in Link's laboratory; they were arranged according to chemical classification and assigned numbers by Overman et al. (1944). In the hope of finding a therapeutic agent other than dicoumarol, the anticoagulant activity
OH
0
Coumarin Dicoumarol
o
Coumafuryl
Warfarin
Diphacinonc
Chlorophaci none
o
[
CH3] CH 2-CH 2-CH 2-tH -CH) .I
Vitamin K J
Figure 83.3 Coumarin, dicoumarol, some synthetic rodenticides, and a natural form of vitamin K.
of some of those analogs was reappraised using not only rabbits (the specie used to detect the hemorrhagic agent of sweet clover poisoning) but also rats, mice, and dogs. Work between 1946 and 1948 identified compounds No. 42 and No. 63 as much more potent than dicoumarol in the rat and dog and as capable of producing a more uniform anticoagulant response and of maintaining a more severe hypoprothrombinernia without visible bleeding than was possible with dicoumarol. Partly on the basis of these observations and partly on the basis of lack of taste and odor, ease of manufacturing the pure compound, and convertibility to a stable water-soluble salt, compound No. 42 was selected. Early in 1948 it was proposed as a rondenticide and promoted by the Wisconsin Alumni Research Foundation. It soon became evident that compound 42 was an important rodenticide. Link (1959) recalled that, although late in 1950 he proposed warfarin for clinical trial, fear of using a highly successful rat poison as a drug prevented significant progress until April 1951, when knowledge of an unsuccessful suicide effectively treated
1810
CHAPTER 83
Rodenticides
Synonyms The name warfarin (BSI, ICPC, ISO) is in common use except in France, where the compound is called cumafene; in Russia, where it is called zoocoumarin; and in Japan, where it and coumatetralyl both are spoken of as coumarins (JMAF). During development, warfarin was known as Compound-42 or WARF-42. As a drug, the sodium salt is called Coumadin®. Trade names for the rodenticide have included Arthrombine-K®, Dethmore®, and Panwartin®. The CAS registry number is 81-81-2. Difenacoum Br
Brodifacoum
Physical and Chemical Properties Warfarin has the empirical formula C19H1604 and a molecular weight of 308.32. It forms tasteless, odorless, and colorless crystals with a melting point of 15-161°C. It is practically insoluble in water and benzene, moderately soluble in alcohols, and readily soluble in acetone and dioxane. The sodium salt is fully soluble in water. History, Formulations, and Uses The history of warfarin is outlined above. It is formulated as a dust (10 gm of active ingredient per kilogram) for use in holes and runs and as a powder (1 and 5 gm of active ingredient per kilogram) for mixing with bait to a final concentration of 50 ppm for control of the common rat or 250 ppm for control of the ship rat and mice. Warfarin also is available in many forms of prepared bait.
Figure 83.4
83.5.1.2 Toxicity to Laboratory Animals
by vitamin K and transfusion of fresh whole blood (Holmes and Love, 1952) brought reassurance. Progress was so rapid that warfarin was used in 1955 for treating then-President Eisenhower. The use of Dicumarol as a drug and the use of warfarin as a drug and as a rodenticide did not go unnoticed by those who sought a compound even more effective than warfarin-and free of patent restrictions. The result was a number of alternative compounds available either as drugs or rotenticides, or, in the case of diphacinone, used like warfarin for both purposes. The appearance of rats resistant to warfarin and to other early anticoagulant rodenticides has stimulated the search for more potent, fast-acting compounds. These are usually called "single-dose-rodenticides" or "second-generation" anticoagulants, among which difenacoum and brodifacoum are coumarin derivatives (Fig. 83.4). Coumarin compounds are relatively free of untoward effects when used therapeutically and have been given for long periods without signs of toxicity. 83.5.1 WARFARIN 83.5.1.1 Identity, Properties, and Uses Chemical Name 3(a-acetonylbenzyl)-4-hydroxycoumarin is the chemical name. Structure
See Fig. 83.3.
Animals intoxicated by warfarin exhibit increasing pallor and weakness reflecting blood loss. Appetite and body weight are not specifically affected. The blood loss may be evident in the form of bloody sputum, bloody or tarry stools, petechiae, or externally visible hematoma. Hematoma formation is more common than free hemorrhage. There is no typical location for hematoma formation, the location of bleeding being apparently a matter of chance in the absence of obvious trauma. Bleeding associated with the central nervous system may be of such location and extent as to cause paralysis of the hindquarters several days before death occurs. Pregnant rats appear slightly more susceptible than nonpregnant ones (Hayes and Gaines, 1950). This may be related to obvious morphological factors, but the decreased metabolism of warfarin in pregnant rats suggests the presence of an inhibitory factor (MacDonald and Kaminsky, 1979). Warfarin may be the only compound for which a log timelog dosage curve with all three segments has been demonstrated experimentally; the 90-day dose LD 50 in rats is only 0.077 mg/kg/day, and the chronicity index is 20.8. Rats tolerated for 300 days a daily dosage slightly greater than the extrapolated 90-day LD 0.01 dosage, specifically 0.02 mg/kg/day. In spite of its considerable cumulative effect, there is a level of intake that is safe for the rat. The same phenomenon permits the use of warfarin as an anticoagulant drug. Other investigators have reported completely different results in the same species. Pyorala (1968), who made elaborate studies of the different susceptibility of male and female rats to warfarin, reported LD 50 values of 62-102 mg/kg for males
83.5 Anti-vitamin K Compounds and 21-33 mg/kg for females. Hagan and Radomski (1953) reported values of 323 and 58 mg/kg in males and females, respectively. Why these values differ by one or two orders of magnitude from those reported by others is not clear. Warfarin is a racemic mixture whether it is used as a rodenticide or as a drug. Almost all published toxicity figures are for the mixture. However, West et al. (1961) were able to separate the isomers and to determine their absolute configuration. Based on prothrombin time measured 24 hr after a single oral dose, the (-) (S) isomer was 5.5 times as active as the (+)( R) isomer. Based on mortality within 10 days after starting daily dietary intake, the (-) (S)-warfarin was 8.5 times as active as the (+)(R) isomer (Elbe et aI., 1966).
Absorption, Distribution, Metabolism, and Excretion Absorption of warfarin from the skin of rats is slow but measurable. Three dermal doses at the rate of 50 mg/kg had about the same pharmacological effect as three oral doses at 0.6 mg/kg (Sanger and Becker, 1975). Because of either species or formulation differences, the results were very different with guinea pigs and rabbits that received a 0.5% solution of the sodium salt in water (with 8% alcohol and 0.1 % of a surface-active agent); single applications at rates of 0.7 and 0.25 mg/kg caused a marked change in prothrombin times in guinea pigs and rabbits, respectively. In fact, one dermal dose at the rate of 0.25 mg/kg was about as effective in rabbits as an oral dose of 2.0 mg/kg (Fristedt and Stemer, 1965). There is great individual variation in the binding of warfarin by the serum proteins of laboratory rats. The rate of excretion showed a strong positive correlation with the concentration of free drug in the plasma (Yacobi and Levy, 1975). Ninety-six hours after intraperitoneal injection of warfarin, the concentrations of activity in the kidney, liver, and pancreas were 3, 12, and 15 times, respectively, greater than that in the blood (Link et aI., 1965). The significance of the pancreatic accumulation remains obscure. Warfarin is readily hydroxylated in vitro and in vivo by rat liver microsomal enzymes to form 6-,8-, and especially 7- hydroxywarfarin. Formation of another metabolite is catalyzed by the soluble fraction of liver in either the presence or absence of oxygen (Ikeda et aI., 1968a, b; Ullrich and Staudinger, 1968). Formation of all these metabolites is stimulated by phenobarbital, chlordane, or DDT. The metabolism is a true detoxication. The inducers can increase the LD 50 of warfarin by more than 10-fold (Ikeda et aI., 1968a). A later study of rats that had received [I4C]warfarin revealed the following compounds in the urine: unchanged warfarin (6.6%), 4' -hydroxywarfarin (21 %), 6-hydroxywarfarin (15.4%), 7-hydroxywarfarin (8.9%), a glucuronide of 7-hydroxywarfarin (3.0%), and an intramolecular condensation product, 2,3-dihydro-2-methyl-4-phenyl-5-oxo-gamma-pyranol (3.2c)(1)benzopyran (DHG) (6.6%). These metabolites were found in the feces also but in different relative concentrations (Barker et aI., 1970). No radioactive carbon dioxide derived from warfarin has been found in exhaled air (Link et aI., 1965). Many of the same metabolites were excreted by guinea pigs,
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but the proportions were different. Salicylic acid, not found in the rat, was found in guinea pig urine. Of all metabolites recovered, only 4' -hydroxywarfarin and DHG showed anticoagulant activity. That of 4' -hydroxycoumarin was slight. That of DHG showed two peaks, of which the second was stronger. This suggests metabolism of the compound, perhaps back to warfarin (Deckert, 1973). Rats injected intraperitoneally with 4C]warfarin excreted approximately 90% of the activity in 14 days, about half in the urine and half in the feces (Link et aI., 1965). Approximately 10% of the activity from 4C]warfarin was excreted in the bile of rats within 5 hr after intraperitoneal injection, but little radioactivity appeared in the feces. Nearly all of the metabolites in the bile were conjugated; they could be released with about equal ease by incubation with ,B-glucuronidase of with gut flora (Elmer et aI., 1977; Powell et aI., 1977). The metabolites identified were the same as those found slightly later in the urine. When guinea pigs were injected with 1 or 2 mg of 4C]warfarin, about 50% of the activity was recovered from urine excreted during the first 12 hr and 87% was found in urine within 7 days. A smaller percentage of large doses was excreted promptly (Deckert, 1973). The action of warfarin (and of fumarin and coumatetralyl as well) on the smooth muscle of the isolated intestine of the rabbit, of the rat (Rattus rattus and Rattus norvegicus), and of Bandicota bengalensis was studied in vitro. There was a fairly identical reduction in peristaltic activity by all three compounds in all four species. The effect was reversible, thus indicating no permanent damage to the tissue (Renapurkar and Deoras, 1982). Warfarin is bound to albumin but can be displaced from albumin by several compounds, including metals (Brodie, 1964; Chakrabarti, 1978).
e
e
e
Resistance to Warfarin Genetic resistance to warfarin among rodents, lagomorphs, and humans is discussed in Section 83.3.1.3. Two cases of intriguing warfarin resistance in humans were reported by Kempin (1983). Both patients under anticoagulant therapy could not kept within therapeutic range. The common factor that was found was heavy daily intake of broccoli (250450 mg/day). Broccoli is an important dietary source of vitamin K (200 J..Lg/100 gm). When the vegetable was removed from the diet, the anticoagulant therapy became effective. Biochemical Effects Warfarin has two actions: inhibition of synthesis of vitamin K-dependent factors (VII, proconvertin; IX, Christmas factor; and X, Stuart factor) and decrease of the production of prothrombin (factor 11) in the liver (Coon and Willis, 1972). In addition, warfarin induces capillary damage. There is unconfirmed evidence that these two actions are produced by the two moieties of the molecule. Thus 4-hydroxycoumarin inhibits the formation of prothrombin and reduces the clotting power of the blood, whereas there is some evidence that at sufficient dosage benzalacetone produces capillary damage and leads to bleeding upon the very slightest
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Rodenticides
trauma. Significantly enough, vitamin K has an antidotal action against both actions of warfarin up to a certain point (Varon and Cole, 1966). The basis for the change in vitamin KJ metabolism associated with poisoning and the alteration of this metabolism in resistant animals probably involves a warfarin-binding protein in the microsomal membranes of the liver. Thierry et al. (1970) found that ribosomes isolated from the livers of resistant rats bind only one-third to one-fifth as much warfarin as ribosomes from normal rats, regardless of whether warfarin is injected before the rats are killed for study or is added to the in vitro preparation. Lorusso and Suttie (1972) found that, when [I4C]warfarinat a concentration of 0.786 ppm was incubated with microsomal preparations, the concentrations reached were 42.0 and 17.7 pmol/mg of protein, depending on whether the preparations were prepared from normal or from warfarin-resistant rats, respectively. Furthermore, the warfarin was bound firmly to membranes of normal rats but loosely to those of warfarin-resistant rats. Vitamin K deficiency caused a 24% increase in the amount of warfarin bound, but this was overcome in animals given vitamin KI1 hr before being killed for in vitro study. Warfarin binding in vitro was reduced 90% in animals injected with warfarin 22 hr before being killed. Although the binding protein was a part of microsomal membranes, it seemed unrelated to cytochrome PA50. A protein which may be the same as the one just discussed has been isolated and shown to have a molecular weight of about 30,000. It may become adherent to ribosomes in the course of their preparation for biochemical study (Searcey and Graves, 1976). Binding of warfarin to cytochromes P-450 and P-448 occurs also and may help to explain changes in the rate of metabolism of warfarin following induction of microsomal enzymes by phenobarbital and other compounds. The stereochemical aspects of the metabolism of warfarin have been studied in great detail (Kaminsky et aI., 1976; Pohl et aI., 1976a, b, 1977). Formation of 7- and 8-hydroxywarfarin is promoted by other cytochromes. The same type of cytochrome is mainly responsible for the formation of each corresponding metabolite, regardless of how the activity of liver microsomal enzymes has been induced (Fasco et aI., 1979). Warfarin has been reported to inhibit (Biezunski, 1970) or to promote (Beracki and Bosmann, 1970) the synthesis of liver microsomal protein and other liver protein. The contradictory results may be explained by differences in procedure, but exactly how is unclear. It is also unclear what bearing the results have on the pharmacological action of warfarin. Warfarin causes a relative increase in vitamin KI oxide in the plasma or liver of people (Shearer et aI., 1973) and rats (Matschiner et al., 1970). The oxide is a naturally occurring compound. In vitamin K-deficient but otherwise normal rats, the oxide and vitamin KI are equally effective, but the oxide is not therapeutic in warfarin-treated rats. It has been proposed that coumarin and related anticoagulants act by inhibiting the conversion of the oxide back to the active vitamin and that the oxide per se is inhibitory. Involvement of the vitamin KI-
vitamin KI oxide cycle in the action of warfarin seems very likely, since the effect of warfarin on this cycle is greatly reduced in resistant rats. The hypothesis that warfarin inhibits prothrombin synthesis by causing accumulation of the oxide does not appear tenable (Caldwell et aI., 1974). However, it seems likely that the brevity of the action of vitamin K in the treatment of poisoning is the result of its irreversible conversion to the epoxide (Shearer and Barkhan, 1979). The superiority of vitamin KI over vitamin K3 in treating warfarin poisoning has been established experimentally (Pe numarthy and Oehme, 1978). For more detail than can be discussed here regarding vitamin K and vitamin K-dependent proteins a book edited by Suttie is available (1979). Although L-histidine at a dietary level of 400 ppm was without effect on rats, it potentiated the lethal action of warfarin (50 ppm) in both laboratory and field tests (Rao, 1979). The biochemical basis for this action of histidine should be explored. Effects on Organs and Tissues A possible oncogenic effect of prolonged warfarin therapy was speculated by Krauss (1982) based on a report by Gore and associates (1982) of an increased incidence of cancer in patients occurring 3 or more years after a pulmonary embolism. This highly speCUlative deduction made from a very limited number of cases has been criticized by Zacharski (1982) who cited results from two cohort studies (Annegers and Zacharski, 1980; Michaels, 1974). In those studies, no increased incidence of malignancies was observed in patients who had received long-term anticoagulant treatments and had been followed for several thousand patient-years. Pathology Animals killed by warfarin show most extreme pallor of the skin, muscles, and all the viscera. In addition, evidence of hemorrhage may be found in any part of the body but usually only in one location in a single autopsy. Such blood as remains in the heart and vessels is grossly thin and forms a poor clot or no clot. In a report of a boxer dog poisoned with a mixture of warfarin and calciferol, delayed necrosis of the tip of tongue and large areas of necrosed skin were seen. It is difficult, however, to attribute the damaging effect on the walls of the vessels to warfarin alone, since calciferol also can induce such lesions (Edlin, 1982). Treatment of Poisoning in Animals A diet containing selenium at a concentration of 2.5 mg/kg of feed was protective against the toxic effects of aflatoxin BI (a bifuranocoumarin) and warfarin in pigs given four daily oral doses of 0.2 mg/kg of body weight. Selenium is a component of glutathione peroxidase, an enzyme that prevents the production of free radicals (Davila et aI., 1983). 83.5.1.3 Toxicity to Humans Experimental Exposure When nine normal men and five normal women were given a single oral dose of warfarin at the rate of 1.5 mg/kg, maximal concentration in plasma was
83.5 Anti-vitamin K Compounds
reached in 2-12 hr. Maximal depression of prothrombin activity was between 36 and 72 hr. Their individual increases in prothrombin time were proportional to their half-time for disappearance of warfarin from the plasma. In other words, the pharmacological effect was greates in those with slower excretion. The half-times for disappearance from the plasma varied from 15 to 58 hr with a mean of 42 hr. Absorption of warfarin from the gastrointestinal tract was apparently complete; no warfarin was found in the stool even after massive doses, and plasma levels and prothrombin activity responses were virtually identical following oral and intravenous administration at the same rates (O'Reilly et a!., 1963). Having established the absolute configuration of the four warfarin alcohols, Ch an et a!. (1972) administered them to volunteers. Reduction of the alcohols was stereoselective. The rate of elimination of one of the isomers (R,S) was much slower than that of the others, and its effect was more sustained. The resulting metabolites were biologically active but not as active as warfarin itself. Six normal subjects were given a single dose of warfarin at the rate of 1.5 mg/kg. Three weeks later, the same people were given 200 mg of phenylbutazone three times a day for at least 8 days; on the fourth day, warfarin was repeated at 1.5 mg/kg. Compared to warfarin alone, administration of warfarin with phenylbutazone increased the prothrombin time even though the plasma concentration and biological half-life decreased. The result (in the face of an obvious inactivation of warfarin) was attributed to displacement of warfarin by phenylbutazone from binding to plasma albumin, making more free drug momentarily available to receptor sites in the liver (O'Reilly and Aggeler, 1968). The mutual displacement of phenylbutazone and warfarin from human plasma albumin has been studied in vitro (Solomon et al., 1968). As shown by study in seven volunteers, the action of triclofos (trichloroethyl sodium phosphate) is similar to that of phenylbutazone. A dosage of triclofos at the rate of 22 mg/kg/day prolonged the prothrombin time even though the dosage of warfarin was reduced. Trichloroacetic acid, a metabolie of triclofos, accumulated in the plasma to an average concentration of 80 ppm. The displacement of warfarin from albumin by trichloroacetic acid was sufficient to account for the observed potentiation of warfarin (Sellers et aI., 1972). At least in the rat, sodiumsalicylate has a similar effect (Coldwell et aI., 1974) but phenobarbital does not significantly influence the binding capacity of the plasma for warfarin (Ikeda et a!., 1968a). In a similar study with 10 male volunteers, both phenobarbital and glutethimide lowered the plasma warfarin concentration and reduced the half-life of warfarin by nearly 50%; chloral betaine had a slight effect also. Phenobarbital and glutethimide significantly reduced the hypoprothrombinemia response of warfarin, but results with chloral betaine were indistinguishable in this regard from results for placebo-treated and untreated controls (MacDonald et aI., 1969). The effects of cimetidine (a drug used to treat peptic ulcer) on the kinetics and dynamics of warfarin were studied in
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seven volunteers. It was shown clearly that cimetidine acted by inhibition of drug metabolism without significant effect on the binding of warfarin to plasma protein (Breckenridge et aI., 1979). Therapeutic Use Use of warfarin as a drug offers greater dosage and, therefore, greater opportunity for side effects than pest control operators encounter. Of course, bleeding is the most common complication of treatment. Most of it is clinically insignificant. Probably many cases remain unpublished. According to one study, the incidence among hospitalized patients was 10%, and it was 40% among ambulatory patients. The incidence of serious hemorrhage was estimated at 2-10% in hospitalized and ambulatory patients, respectively. A series of case reports illustrated some of the circumstances leading to serious hemorrhage. It was concluded that the complication can be kept at a minimum by careful selection of patients, informed and adequate supervision by the physician, and reliable laboratory control (Pastor et aI., 1962). Although the diagnosis in most cases of hemorrhage is obvious, there are exceptions. For example, two cases of intestinal hemorrhage leading to an initial diagnosis of acute abdomen have been reported (Cocks, 1960). Macular, papular, pruritic, or vesicular rashes due to warfarin are unusual, but those that do occur often are in patients who had taken the drug without untoward effect for 3 or more months. The skin returned to normal slowly after medication was stopped, but the dermatitis recurred within 2 or 3 days when medication was renewed (Schiff and Kern, 1968). Necrosis of the skin and subcutaneous tissues of localized areas has been attributed to warfarin only rarely and not always convincingly. For example, in a case reported by Vaughan et al. (1969), a totally unexplained illness suggestive of but not proved to be thrombophlebitis and pulmonary embolism preceded warfarin therapy by 6 weeks and may have been the underlying cause of the complication attributed circumstantially to warfarin. In two other cases, a more persuasive interpretation was made, namely that anticoagulants (heparin and warfarin) acted as neither the preparatory nor the provoking substance for the localized necrosis attributed to their use, but that underlying disease processes, including intravascular coagulation, sepsis, or localized inflammation, triggered a localized purpuric reaction that was then intensified by the warfarin therapy (Martin et aI., 1970). An entirely different kind of complication has involved potentiation of warfarin by disulfiram (Rothstein, 1968) or interference with its action by other drugs including griseofulvin (Cullen and Catalano, 1967) and phenobarbital (Robinson and MacDonald, 1966) or by insecticides. In one case, medical use of warfarin was nullified by use of 5% toxaphene and 1% lindane to dust sheep. Response to warfarin returned to normal within about 3 months after exposure to the insecticides (Jeffery et aI., 1976). Of course, discontinuing a drug that promotes the metabolism of warfarin has the same effect as introducing a drug such as disulfiram that interferes with metabolism of the anticoagulant.
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The success of cardiac value prostheses requires anticoagulation to prevent immobilization of the valve by thrombi and to minimize the chance of emboli. Installation of such prostheses in young women to correct rheumatic mitral or aortic stenosis increases their chance of surviving and reproducing, but it necessarily complicates any pregnancies they may have. Anticoagulants also may be used in women of childbearing age to treat thrombophlebitis, embolic disease, and a few other conditions. It has long been recognized that administration of any anticoagulant during pregnancy increases the danger of hemorrhage either during the course of gestation or during delivery. It gradually has become evident that warfarin also is teratogenic in humans (Beckert et al., 1975; DiSaia, 1966; Keber et al., 1968; Pettifor and Benson, 1975; Shaul et al., 1975; Sherrod and Harrod, 1978; Tejani, 1973). At least 29 cases of congenital anomaly have been attributed to warfarin (Hall et al., 1980). Most if not all of the recognized cases have involved nasal hypoplasia ranging from barely recognizable to very severe. Many of the babies had chondrodysplasia punctata, and this defect of cartilage development may be the basis not only of the nasal deformity but also of defects of the bones such as meningocele, deformities of the limbs, and a high arched palate seen much more rarely in babies of women treated with warfarin during the first trimester. Other teratogenic effects reported in one or more cases include microphthalmia, blindness, hydrocephalus, persistent truncus arteriosus, and mental retardation. Of 423 reported pregnancies in which coumarin derivatives were used, not over two-thirds resulted in apparently normal infants, one-sixth resulted in abortion or stillbirth, and one-sixth resulted in abnormal liveborn infants of which 29 showed fetal embryopathy. The critical period of exposure seemed to be between 6 and 9 weeks of gestation. Five cases of typical embryopathy and eight other cases showed central nervous system abnormalities following exposure to coumarin derivatives during gestation, but no critical period of exposure was evident (Hall et al., 1980; Stevenson et al., 1980). Twenty-nine of 423 pregnancies is a high incidence of teratogenic effect even if one takes into account that only medically reported cases were available for consideration. A different kind of evidence for the teratogenic action of warfarin involved a family with no history of consanguinity, birth defects, or mental retardation that produced one normal child in a pregnancy without warfarin and two deformed children in separate pregnancies in which warfarin was used (Sherrod and Harrod, 1978). On the other hand, inasmuch as defects occur in only about one-third of instances, none may be found in some small series of cases (Chong et al., 1984). The use of heparin during gestation does not result in a significantly better outcome of pregnancy than that obtained with warfarin. In 135 published cases, about two-thirds were apparently normal, one-eighth were stillborn, and one-fifth (of whom one-third died) were premature (Hall et al., 1980). Kaplan (1985) and Zakzouk (1986) reviewed the subject of warfarin-associated malformations and established that, based on the timing of warfarin exposure, second- and third-trimester exposure predisposes to central nervous system abnormalities
whereas first-trimester exposure is associated with the warfarin embryopathy: midface and nasal hypoplasia, optic atrophy, hypoplasia of the digits, and mental impairment. In addition, Kaplan reports a case of Dandy-Walker malformation associated with warfarin exposure confined to weeks 8-12 of gestation. Four previous cases of such association were known but the gestational exposure period was much longer. Not only has hereditary resistance of people to warfarin been observerd (O'Reilly, 1970) but also exceptional susceptibility, also presumably on a hereditary basis, has been reported (Solomon, 1968). Accidental and Intentional Poisoning A 32-year-old man was murdered by feeding him warfarin for 13 days. On the fourth day after intake started, the victim began having severe nosebleeds. Later, he bled from the mouth. Two days before death, he complained of pain in his limbs. His symptoms became worse and he died of circulatory failure on day 15 (Pribilla, 1996). The initial symptoms in an attempted suicide using warfarin were back pain and abdominal pain. The onset occurred 1 day after the sixth daily dose. A day after onset, vomiting and attacks of nose bleeding occurred. On the second day of illness, when admitted to hospital, the patient was observed to have a generalized petechial rash (Holmes and Love, 1952). In Korea, a family of 14 persons lived for a period of 15 days on a diet consisting almost entirely of corn (maize) meal containing warfarin. The first symptoms appeared 7-10 days after the eating of warfarin was begun. Massive bruises or hematomata developed at the knee and elbow joints and on the buttocks in all cases. Extensive gum and nasal hemorrhage usually appeared about a day later, and by days 15 blood loss was extensive (Lange and Terveer, 1954). A suicidal gesture that was reported and treated after only a single ingestion of a heaping tablespoonful of 0.5% wafarin produced no illness and not even an increase in prothrombin time (Kellum, 1952). There have been at least two attempted murders with warfarin (Ikkala et al., 1964; Nilsson, 1957). In each instance there were recurrent bouts of hemorrhagic difficulty, including hematuria, epistaxis, severe bruises without any history of trauma, and intestinal hemorrage. Abdominal or back pain was present. Each patient recovered promptly in hospital, but one had relapses for nearly a year and the other had bouts of poisoning of 2.5 years following repeated doses. An anticoagulant drug was suspected from the first in both cases, but finding the source proved difficult. Solution of each case was essentially epidemiological, but measurement of warfarin in the plasma was decisive in one case. Faced with the evidence, the daughterin-law of a 72-year-old woman and the wife of a 69-year-old man both confessed to the police. Although numerous accidental ingestions by children and adults have been reported to the New York Poison Control Center, no known injury from these ingestions has been observed (Jacobziner and Raybin, 1960).
83.5 Anti-vitamin K Compounds An outbreak of hemorrhagic disease due to the use of warfarin-contaminated talcum was described in Vietnam (Martin-Bouyer et aI., 1983). Of the 741 cases located in Ho-ChiMinh City, all in infants (55% under 2 months of age), 177 died. Eleven samples of baby powder were ana1yzed and concentrations of warfarin ranged from 1.7 to 6.5%. The percutaneous penetration of warfarin contained in the contaminated talc was studied in a young healthy female baboon, treated twice daily with a topical application of 3 gm of talc containing 3% warfarin (188 mg/kg/day of warfarin). One control animal was treated with uncontaminated talc. On the fifth day, the treated animal began to show signs of intoxication with profuse bleeding and died. At necropsy there were two large subcutaneous hematomata on the skull and the peritoneal cavity with filled with unclotted blood. On day 3 of treatment, a blood sample showed severe disturbances of the hepatic coagulation factors. Electron microscopy showed an increased number of swollen and misshapen mitochondria in the hepatocytes (Dreyfus et aI., 1983). Warfarin was administered to an 11-month-old baby girl by her psychologically disturbed mother. Upon hospitalization the parameters of coagulation were elevated (prothrombin time 53 sec, control 12 sec); the child had multiple hematomata and had a bloody discharge from her left ear. Treatment with vitamin KJ and infusion of fresh frozen plasma stopped the bleeding (White, 1985). Use Experience The safety record of warfarin used as a rodenticide has been excellent. One case of poisoning has been attributed to extensive, prolonged skin contact in the process of preparing and distributing baits. Unlike most solid bait, which is prepared by mixing ground grain with starch containing warfarin powder, this bait was prepared by pouring a 0.5% solution of the sodium salt over dried bread. The hands of the 23-yearold farmer who used this method were wet with the solution each of the 10 times he made bait during a 24-day period, and he did not wash his hands until several hours after each application. Two days after the last contact with rodenticide, gross hematuria appeared. Next day hematomata were noticed on the arms and legs; there was dull pain in both groins. The hematuria subsided after 3 days of rest but recurred along with nose bleeding when the man returned to work. When he was admitted to hospital, prothrombin, clotting, and bleeding times were abnormally long and anemia was severe (hemoglobin, 8.2%; red cell count, 2.9 million/mm3 ). The patient responded promptly to treatment with vitamin KJ (Fristedt and Sterner, 1965). Atypical Cases of Various Origins Mogilner et al. (1974) described three fatal cases of what they considered to represent Reye's syndrome. One of the young children was said to have ingested warfarin that she found by the road on her way to kindergarten 4 days before hospital admission; warfarin was found in the urine and traces were present in the blood. In another case, "significant amounts of warfarin were found in the urine and also in autopsy samples of liver and kidney tissue;"
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no source of the rotendicide was reported, and no information was given that would permit evaluation of the validity of the chemical analysis. There was no indication of warfarin in the third case. The authors considered that it was justified "to add warfarin to the long and varied list of etiological or precipitating factors in Reye's syndrome." The death of a 2-year-old boy was attributed to warfarin poisoning on his death certificate but probably was the result of beating, with pneumonia as a terminal event (Hayes and Vaughn, 1977). Dosage Response A total dose of 1000 mg of warfarin consumed in 13 days (about 1.1 mg/kg/day) was fatal (Pribilla, 1996). Serious illness followed the ingestion of 1.7 mg warfarin/kg/day for 6 consecutive days with suicidal intent. This would correspond to eating almost 1 pound of bait (0.025% warfarin) each day for 6 days. All signs and symptoms were caused by hemorrhage and, following multiple small transfusions and massive doses of vitamin K, recovery was complete (Holmes and Love, 1952). In the Korean cases, the dosage of the different individuals was determined to vary from about 1 to 2 mg warfarin/kg/day. As a result of this exposure and without benefit of treatment, 2 of the 14 persons died. A 19-year-old girl, who was in a state of shock and severe hemorrhage 2 days after the warfarin diet was discontinued, recovered following a blood transfusion and small daily doses of vitamin K. The remaining 11 members of the family recovered within a week after exposure, although only small daily doses of vitamin K were given and they all had shown marked signs of poisoning when they first accepted treatment. There was reason to think that those who died had received slightly higher dosages than those who survived (see Table 2.9 in Hayes, 1975). Recovery of the 12 survivors was complete. The entire episode was made possible only by a series of unusual events and by the extraordinary apathy of the family, resulting in their totally ignoring unmistakable signs of illness (Lange and Terveer, 1954). A single intravenous therapeutic dose of the sodium derivative (40-60 mg or about 0.7 mg/kg) in humans may produce some increase in prothrombin time within 2 hr and usually produces a substantial increase within 14 hr. The average maximal response is on the fourth day. Spontaneous recovery to normal occurs about 8 days after a single therapeutic dose. Thus significant depression of prothrombin level is maintained for 3-6 days. In the treatment of thromboembolic disease, a maintenance dose of about 2-10 mg/day is required to keep the prothrombin level between 10 and 30% of normal. Patients have been thus maintained for years. If human susceptibility to warfarin were different (as it is in a few genetically determined cases), the therapeutic dosage could be adjusted accordingly. It is interesting, however, that the upper limit of the usual maintenance dosage for humans (about 0.14 mg/kg/day) is an LD 95 for the rat. The inherently lesser susceptibility of humans to the compound undoubtedly contributes to its safety as a rodenticide. The threshold limit value of 0.1 mg/m 3 indicates that
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CHAPTER 83
Rodenticides
occupational intake of warfarin at the rate of 0.014 mg/kg/day is considered safe. Laboratory Findings Metabolites of warfarin including 5-, 6-, 7-, and 8-hydroxywarfarin and two a1iphatic side-chain a1cohols have been identified in the urine of normal volunteers who had received a single oral dose at the rate of 1.5 mg/kg (Chan et aI., 1972; Lewis and Traeger, 1970). This is, however, not a common way to confirm poisoning. Adequacy of treatment with warfarin usually is followed by measuring prothrombin time. In case of poisoning, the prothrombin time is greatly prolonged. The coagulation time is definitely increased by the Lee-White method and slightly increased by the capillary tube method. Bleeding time often is normal. Urine may be normal in appearance but may contain many red cells on microscopic examination, or it may be grossly hemorrhagic. The red cell count and hemoglobin gradually fall if bleeding continues. In terminal cases a state of shock develops. Sixty-nine euthyroid patients being treated with warfarin for thromboembolic disease showed no evidence of hyperthyroid condition, and 14 of them showed a hypothyroid tendency associated with an elevation of the thyroxine-binding capacity of plasma globulin. However, no clinical evidence of thyroid dysfunction was reported, and it was uncertain whether the small changes in laboratory tests of thyroid function were caused by warfarin (Braverman and Foster, 1969). Plasma levels of warfarin were 6.8 and 11.2 ppm 4 and 7 hr, respectively, after the ingestion of 50 mg of warfarin sodium in a suicide attempt. Plasma levels declined thereafter, and the half-time disappearance was calculated as 46 hr. Part of the dose was removed by gastric lavage soon after ingestion. This and other appropriate treatment prevented any increase in bleeding tendency (Cole and Bachmann, 1976). Pathology Apparently there is only one complete description of human pathology associated with uncomplicated warfarin poisoning, that of Pribilla (1996). The findings in that case were strikingly different from typical findings in the rat: exsanguination was less complete, as indicated by the fact that the liver was not tan in color and bleeding was far more generalized and not restricted to one or a few large hematomata. The two factors may be related in that bleeding into the organs may have interfered with their function and hastened death. In addition to generalized bleeding (due mainly to deficiency of coagulation), evidence of capillary damage and of parenchymal injury of the liver was found in the human case. In spite of the obvious differences between the findings in human and rat, the similarity was also striking because subserosal and intraseptal bleeding was prominent in the human case.
should be continued until the prothrombin time has reached normal. In a seriously ill patient, a small transfusion of carefully matched whole blood should be given initially and repeated daily until the patient has returned to normal. Such a patient should be given vitamin Kl also. If it were ever necessary to treat a patient in shock from blood loss resulting from warfarin poisoning, frequent small transfusions and a complete consideration of the blood chemistry would be in order. Any large hematomata should be the subject of a surgical consultation, but any surgical action should be taken only after the clotting power of the blood is restored to normal. The progress of the patient should be followed by the prothrombin test. Tests should be made at least twice daily until a return to normal is clearly established and stable. 83.5.2 COUMAFURYL 83.5.2.1 Identity, Properties, and Usses Chemical Name 3-[-1 (2-furanyl)-3-oxobutyl]-4-hydroxy2H -1-benzopyran-2-one in the chemical name. Structure
See Fig. 83.3.
Synonyms Coumafuryl is the common name approved by ISO. Other names for the compound include fumarin (BSI) and tomarin (Turkey). Trade names include Fumarin®, Fumasol®, Krumkil®, Lurat®, Ratafin®, Rat-a-way®, and Tomarin®. The CAS registry number is 117-52-2. Physical and Chemical Properties The empirical formula for coumafuryl is C17H140S, and the molecular weight is 298.28. It is a crystalline solid melting at 214°C. Use
Coumafuryl is an anticoagulant rat poison.
83.5.2.2 Toxicity in Laboratory Animals Coumafuryl is very similar to warfarin (see Fig. 83.3). The oral LD 50 for rats is quoted as 0.4 mg/kg (Wiswesser, 1976). Coumafuryl poisoning was observed in young chicks (less than 1 week old) with a mortality rate of 100%. Hemorrhage and unclotted blood was noted in the abdominal and thoracic cavities. At necropsy, crops and gizzards contained feed. Analysis of the content detected approximately 340 mg/kg of coumafuryl. Investigations found coumafuryl was present in the wood-straw mats in the chicken boxes (Munger et aI., 1993). 83.5.2.3 Toxity to Humans It is inevitable that a number of children and perhaps others
Treatment of Poisoning After blood has been taken for prothrombin and other differential diagnostic tests, vitamin Kl in a dose of 5-10 mg should be given three times on the first days of treatment irrespective of symptoms. The vitamin should be given intravenously slowly, usually by infusion. Smaller doses
have ingested coumafuryl. In fact, McLeod (1970) reported that coumafuryl and warfarin were among the pesticides most often ingested by persons (mainly children) admitted to a large hospital in New Orleans. However, these two compounds were the least hazardous pesticides in terms of morbidity.
83.5 Anti-vitamin K Compounds
Treatment of Poisoning Treatment is the same as that for warfarin (see Section 83.5.1.3). 83.5.3 DIPHACINONE 83.5.3.1 Identity, Properties, and Usses Chemical Name chemical name. Structure
2(Diphenylacetyl)indan-l,3-dione is the
See Fig. 83.3.
Synonyms Diphacinone (ANSI, BSI, ISO) is the common name in use except in Turkey and Italy, where diphacin is used, and in Russia where ratindan is used. Other nonproprietary names include dipazin and diphenacin. As a drug, the compound is known as diphenadione. Trade names for formulated baits containing diphacinone include Diphacine®, Ramik®, Promar®, and Gold Crest®. The CAS registry number is 82-66-6. Physical and Chemical Properties Diphacinone has the empirical formula C23H1603 and a molecular weight of 340.40. Its technical grade is a yellow crystalline powder melting at 145°C; it is slightly soluble in water (0.3 mglliter) and soluble in acetone (29 gmlliter) and toluene (73 gmlliter). Diphacinone is rapidly decomposed in water by sunlight. History, Formulation, and Uses The rodenticidal activity of diphacinone was described in 1952. It is formulated as prepared weather-resistant baits (pellets or meal) in concentrations of 50 mg/kg. A dry concentrate (1 gm/kg) for mixing with cereal bait is also available. All formulations are for professional application only. Diphacinone is used to control mice, rats, pairie dogs (Cynomys spp.), ground squirrels, voles, and other rodents. 83.5.3.2 Toxicity in Laboratory Animals In a study of bishydroxycoumarin, ethyl biscoumacetate, and 17 analogs of indandione in rabbits, diphacinone was found to be the most hypoprothrombinemic. A marked response lasting about 7 days was produced by a dosage of only 0.05 mg/kg. The acute oral LD 50 ranges from 0.3 to 2.3 mg/kg in the rat and from 3.0 to 7.5 mg/kg in the dog. It was found to be 14.7 mg/kg for cats and 150 mg/kg for pigs. In mice and rabbits, the oral LD 50 is 340 and 35 mg/kg, respectively. For the mallard duck, it is 3158 mg/kg. The LD 50 associated with 14 daily oral doses in rats was 0.1 mg/kg (Correll et aI., 1952). Acute percutaneous LD 50 for rats is less than 200 mg/kg. In a 21-day subchronic percutaneous study in rabbits, the no-effect level was 0.1 mg/kg daily. Diphacinone is neither a skin and eye irritant nor a skin sensitizer. An acute inhalation of diphacinone dust in the rat as shown an LC 50 of less than 2000 mg/m3 of air. Diphacinone is not mutagenic in the Ames test. Sprague-Dawley rats were fed for 21 days on a diet containing 1, 2, or 4 ppm diphacinone. All animals in the 2 and 4 ppm
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groups died and postmortem examination revealed massive internal hemorrhages. On day 21, the prothrombin clotting time of the animals in the 1 ppm group was not affected. A second study was performed in which the rats were fed for 90 days on a diet containing 0.0313, 0.625, 0.125, 0.25, or 0.5 ppm of diphacinone, or approximately 0.002, 0.003, 0.006, 0.013, and 0.025 mg/kg/day. One male in the 0.25 ppm group died on day 17 of treatment and another male in the 0.0625 ppm group on day 20 from a subdural hemorrhage. However, the mean prothrombin clotting times of the animals surviving the treatment period were not affected. The only parameter that showed some variation was the fibrinogen level, which was lower in the 0.5 ppm group (Elias and Johns, 1981). The most interesting aspect of the toxicity of diphacinone involves species difference. The oral LD 50 of the compound for vampire bats (Desmodus rotundus) was 0.91 mg/kg, whereas a dosage of 5 mg/kg produced no sign of illness in cattle (Elias et aI., 1978). The blood of beef cattle given a single intraruminal injection of the compound at a rate of 1 mg/kg became toxic to these bats and remained toxic for 3 days without harming the cattle. As indicated by examination before and 2 weeks after treatment, cattle dosed in this way on three ranches in Mexico experienced a 93% reduction in vampire bat bites. Bioassays of milk and liver indicated that there was no residue problem (Thompson et aI., 1972). Residue studies indicated that people may safely eat meat, including liver and kidney, from treated cattle (Bullard et aI., 1976).
Absorption, Distribution, Metabolism, and Excretion When 14C-Iabeled diphacinone was administered orally to mice, radioactivity reached its highest levels in the liver and lungs. The concentration in the liver reached its maximum in 7.5 and 3.0 hr in males and females, respectively (Cahill and Crowder, 1979). In another study, rats and mice were orally administered 14C-Iabeled diphacinone at dosages of 0.2 and 1.5 mg/kg. In rats, about 70% of the dose was excreted in the feces and 10% in urine in 8 days. The same elimination pattern was observed in mice. Eight days after the administration of the compound in rats and 4 days in mice, the liver had the highest level of residues, but kidneys and lung also contained significant levels of residues; brain, fat, and muscles had the lowest levels. Diphacinone is not extensively metabolized in rats, less than 1 % of the dose being expired as C02 hr. The metabolism pattern in rats involved mainly hydroxylation and conjugation reactions (Yu et al., 1982). Mode of Action Diphacinone inhibits the K-enzyme complex (liver-synthesized coagulation proteins: factors 11, VII, and X), and this inhibition phase lasts approximately 30 days in dogs, as opposed to the relatively short effect of warfarin (Mount and Feldman, 1983). The prolonged action of diphacinone may be due to protein binding in the liver or low excretion rate or a combination of both factors. Treatment of Poisoning in Animals In dogs, the results of the usual vitamin K oral therapy after diphacinone poisoning
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CHAPTER 83 Rodenticides
ware poor. Dogs treated with a sufficient quantity of vitamin K relapsed fatally several days after the corrective treatment stopped. The response to treatment seems to vary according to the amount of exposure to the rodenticide. The recommended therapeutic dose of vitamin KJ is 5 mg/kg of body weight in subcutaneous injections for at least 5 consecutive days (Mount and Feldman, 1983). 83.5.3.3 Toxicity to Humans Therapeutic Use Diphacinone has been used as a therapeutic agent because it has a relatively long duration of action. Its half-life in humans is 15-20 days. A single oral dose of diphacinone of 4 mg/person produced a clearly detectable reduction of prothrombin about 14 hr after ingestion and slightly more reduction a day later, with recovery to normal by the third day. A smaller, uncertain reduction was produced by 2 mg/person. A single 20-mg dose caused hypoprothrombinemia that was definite in 14 hr, marked in 48 hr, and persisted from 6 to 10 days. The recommended initial dose for therapy was 20 mg followed by daily doses of 2-4 mg (Field et aI., 1952). The drug was in use until a few years ago. Although there were no adverse effects except occasional nausea and not unexpected hemorrhagic complications at high dosage levels, caution was advised because of its close relation to phenidione, which has caused agronulocytosis, hepatitis with jaundice, nephropathy with actue renal tubular necrosis, severe exfoliative dermatitis, and massive generalized edema (American Medical Association (AMA), 1977). The drug ceased to be listed in the AMA Drug Evaluations of 1980. Treatment of Poisoning Treatment is the same as that for warfarin, but based on experimental data from animals it would seem advisable to increase the dose of vitamin K as well as the duration of the corrective treatment. 83.5.4 BRODIFACOUM 83.5.4.1 Identity, Properties, and Usses Chemical name 3-[3-(4'-bromo-[1,1'-biphenyl]-4-yl)-1,2,3, 4-tetrahydro-l-naphthylenyl]-4-hydroxy-2H -1-ben-zopyran-2one is the chemical name. Structure
See Fig. 83.4.
Synonyms Brodifacoum is the approved common name (ISOBIS). Trade names for the formulated material include Ratak®, Volak®, and Talon®. The code names are WBA 8119 and pp 581. The CAS registry number is 56073-10-0. Physical and Chemical Properties The empirical formula for brodifacoum is C31H23Br03 and the molecular weight is 523.4. It is an off-white to fawn-colored odorless powder with a melting point of 228-232°C. It is of very low solubility in water (less than 10 mg/liter at 20°C and pH 7). Brodifacoum is
slightly soluble in alcohols and benzene and soluble in acetone. It is stable at room temperature. It has a very low vapor pressure of less than 1.33 x 10- 7 kPa (l x 10-6 mm Hg) at 25°C. History, Formulation, and Uses The rodenticidal properties of brodifacoum were described in 1976. It is an indirect anticoagulant active against rats and mice including strains resistant to warfarin and other anticoagulants (Rennison and Hadler, 1975). It is also used to control other wild rodents. Brodifacoum is formulated as ready-to-used baits of low concentration (20 and 50 mg/kg of bait). A single ingestion is usually sufficient to kill. 83.5.4.2 Toxicity to Laboratory Animals Brodifacoum is extremely toxic to a number of mammalian species. The oral and dermal LD 50 values of the technical material are given in Table 83.5. In chicken, the oral LD 50 is reported to be 4.5 mg/kg and in the mallard duck it is 2.0 mg/kg. In a 42-day feeding study in rats, a concentration of 0.1 ppm did not induce any adverse effect (Worthing and Walker, 1983). Several cases of poisoning in domestic animals have been reported. One day after being seen to ingest brodifacoumcontaining bait, a 17-kg cocker spaniel developed depression and icterus accompanied by accelerated pulse and rapid and labored respiration. Despite supportive therapy, the dog died the same day. The autopsy confirmed the icter and showed approximately 1 liter of unclotted blood in the thoracic cavity and 100 ml in the pericardiac sac. Numerous hemorrhagic areas were seen in the serous membranes (Stowe et aI., 1983). A 4year-old cross-bred bitch was noticed to be depressed and weak the day following the laying of Talon® bait (0.005% brodifacoum); she was found dead in her kennel the next morning. The actual dose of brodifacoum ingested was unknown but it was estimated that the maximum quantity of bait eaten could have been 900 gm, resulting in an intake of 45 mg of active ingredient. At autopsy, the thoracic cavity contained approximately 1.8 liter of unclotted blood and a single clot was adherent to the base of the heart and to the aorta. Subcutaneous bruising was present on the rib cage. Brodifacoum was found in the liver at a concentration of 0.8 mg/kg (McSporran and Phillips, 1983).
Table 83.5 Single-Dose LD 50 for Brodifacoum LD50 Species
Route
Rat, M
oral
Rat
dermal
(mg/kg) 0.27 50.00
Mouse, M
oral
0.40
Guinea pig, F
oral
0.28
Rabbit, M
oral
0.30
Dog
oral
0.25-1.0
Cat
oral
,,=,0.25
Reference
83.5 Anti-vitamin K Compounds The acute oral toxicity of brodifacoum was examined in sheep. An LD 50 of 11 mg/kg was considered a good estimate (Godfreyetal.,1985). A white-winged duck (Cairina sculata) of a zoological setting was found with bilateral epistaxis and anemia. Brodifacoum was detected in the blood at a level of 0.002 ppm. The bird was treated with injectable and oral vitamin KI and transfused with 40 ml of whole blood and fully recovered (James et al.,1995). Absorption, Distribution, Metabolism, and Excretion Brodifacoum is absorbed through the gastrointestinal tract. When orally administered to male Sprague-Dawley rats at doses ranging from 0.1 to 0.33 mg/kg, brodifacoum exhibited a remarkably steep dose-response curve; 0.1 mg/kg failed to show an effect on the plasma prothrombin level within 24 hr, whereas 0.2 mg/kg reduced the prothrombin complex activity to 7% of normal values and 0.33 mg/kg reduced it to 4% of normal. Concentrations in the liver were rapidly established and remained relatively constant for at least 96 hr. The mean liver/serum concentration ratio is approximately 20. Disappearance from serum is slow with a half-life of 156 hr or even more. The slow disappearance from the plasma and liver and the large liver/serum ratio probably contribute to the higher toxicity of brodifacoum than of warfarin. These particular features may also explain the efficacy of brodifacoum against warfarin-resistant rats (Bachmann and Sullivan, 1983). Six weeks after intravenous administration of a single 1 mg/kg dose of brodifacoum to male New Zealand White rabbits, the prothrombin complex activity was still lower than 30% of normal (in the early part of the study, subcutaneous injections of vitamin K were given to prevent lethal hemorrhage). In the same study (Park and Leck, 1982), it was shown that in the rabbit, the maximal antagonism of vitamin KI by warfarin was produced by a dose of 63 mg/kg, whereas a similar result was obtained with only 1 mg/kg brodifacoum. It was shown that, in warfarin-resistant and warfarin-sensitive rats, brodifacoum produced the same rate of degradation of prothrombin complex activity as warfarin and significantly reduced the activity of clotting factors 11, VII, and X without affecting factor V. It was also demonstrated that brodifcacoum has the same mechanism of action as warfarin: reduction of vitamin K-dependent clotting factor synthesis by interruption of the vitamin K-epoxide cycle (Leck and Park, 1981). In mongrel dogs, the elimination of brodifacoum follows a classical experimental decay with a distributive half-life of 1.4 days and an elimination half-life phase of 8.7 days (Murphy et aI., 1985). Brodifacoum was fed to four dogs for 3 consecutive days, producing a cumulative dose of 1.1 mg/kg body weight. Serum brodifacoum concentrations were monitored. Inappetence and hemorrhagic tendencies were exhibited by day 5 postrodenticide exposure. One-stage prothrombin time, APTT, and ACT were 25% greater than time zero values at 24, 24, and 72 hours postdosing, respectively. All laboratory parameters returned to normal within 48 hours of initiating vitamin KI therapy
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(0.83 mg/kg orally for 5 days). Serum brodifacoum concentrations were highest (1065-1215 ng/mL) during the 3 days after dosing and were detectable (3.0-7.5 ng/mL) until day 24 brodifacoum postexposure. A mean brodifacoum elimination half-life of 6 ± 4 days was observed (Woody et aI., 1992). Factors Influencing Toxicty Pretreatment of rats daily by intraperitoneal injection of phenobarbital at 80 mg/kg for 2 consecutive days followed by a single administration of brodifacoum at 0.2 mg/kg by stomach tube reduced the anticoagulant effects, although the reduction was less marked than in the case of warfarin (Bachmann and Sullivan, 1983). It is also known that a very large number of drugs of different chemical structures can interact with coumarin anticoagulant therapy in humans (Koch-Weser and Sellers, 1971). Pathology Necropsies of poisoned dogs have shown that, in addition to large collections of unclotted blood, a number of lesions were also present, including bile statis with large amounts of brown pigment accumulated in the portal triads and in the macrophages of the periportal regions and congestion of the spleen with accumulation of golden brown pigment, believed to be hemosiderin, in the red pulp (Stowe et aI., 1983). Treatment of Poisoning in Animals Treatment should be as for other anticoagulant rodenticides: vitamin KI at dosage of 2.5-5.0 mg/kg and transfusion of fresh whole blood. Because of the long-lasting effect of brodifacoum, vitamin K therapy must be continued for at least 2-3 weeks. 83.5.4.3 Toxicity to Humans Brodifacoum has not been used therapeutically in humans. Cases of attempted suicide have been reported. A 31-year-old mentally disturbed woman ingested over a 2-day period approximately thirty 50-gm packages of Talon® (approximately 75 mg of brodifacoum). Two days later she was brought to the hospital's psychiatric unit, without any physical signs or symptoms. The routine laboratory tests showed a prothrombin time of 72 sec (control, 12 sec) and an activated partial thromboplastin time greater than 100 sec (normal, 25-35 sec). In spite of prolonged administration of large amounts of vitamin KI and repeated infusion of fresh frozen plasma, the depression of the prothrombin complex activity persisted for more than 45 days after the ingestion (Lipton and Klass, 1984). Jones et al. (1984) have reported a similar case in a 17 -year-old boy who attempted suicide by ingesting approximately 7.5 mg (0.12 mg/kg) brodifacoum. He was first seen for a gross hematuria, rapidly followed by epistaxis and gum bleeding. The prothrombin time and the activated partial thromboplastin time were considerably prolonged. The levels of plasma clotting factors 11, VII, IX, and X were decreased. Factor V was normal. Vitamin KI and plasma therapy were instituted and had to be continued for 55 days until the patient's coagulation remained normal and stable.
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CHAPTER 83 Rodenticides
Dosage Response From the two cases of poisoning reported above, it can be seen that the total doses ingested were in one case 75 mg and in the second one 7.5 mg and that the effects on the clotting factors were maximum in both cases, the clinical signs being almost absent in the woman who absorbed 75 mg of brodifacoum (although it was ingested over a 48-hr period). Therefore it would seem that, above a certain threshold, the response is maximum. This was also shown in rabbits given 1 and 10 mg/kg brodifacoum (Park and Leck, 1982). Murphy et al. (1985) have shown that in dogs serum concentrations below 12 mg/ml caused no measurable coagulopathic effects after cessation of vitamin K therapy. Plasma brodifacoum concentration was compared to prothrombin levels over time in a case of brodifacoum poisoning. Brodifacoum was eliminated according to a two-compartment model, with an initial half-life of 0.75 days and a terminal halflife of 24.2 days. On admission, the brodifacoum level was 731 microgramslL and the patient suffered severe urinary tract hemorrhage, requiring transfusion of blood products. Persistently increased prothrombin times necessitated treatment with phytonadione up to 80 mg/day for 4 months, until the brodifacoum level reached 10 microgramslL (Hollinger et al., 1993). The plasma concentration, plasma half-life, and mean retention time of brodifacoum (among other anticoagulant rodenticides) were determined in dogs in which preliminary diagnosis of anticoagulant poisoning had been made. Analysis was performed with high performance liquid chromatography (HPLC) on the plasma. In 7 dogs, the estimated half-time of brodifacoum ranged from 0.9 to 4.7 (median 2.4) days with a mean retention time of 1.9 to 3.7 (median 2.8 days) (Robben et al., 1998). The case of a voluntary ingestion of brodifacoum by a 39-year-old man was reported by Sheen et al. (1994). His prothrombin and partial thromboplastin times were respectively 150 and 113 seconds. Treatment with a daily dose of 200 mg of phytonadione for 5 months corrected the coagulopathy with no side effects occurring. Prolonged follow-up and vitamin K treatment were necessary in a child with spontaneous hemorrhage from her nose, mouth, and urinary tract following accidental ingestion ofbrodifacoum (Travis et al., 1993). 83.5.5 CHLOROPHACINONE 83.5.5.1 Identity, Properties, and Usses Chemical Name 2-[4-(chlorophenyl)phenylacetyl]-1 H -indene-l ,3(2H)-dione is the chemical name. Structure
See Fig. 83.3.
Synonyms Chlorophacinone is the approved common name (BSI-ISO). Trade names for the formulated products include Caid®, Liphadione®, Raviac® , Drat®, Quick®, Lepit®, Rozol®, and Saviac®. The CAS registry number is 3691-35-8.
Physical and Chemical Properties Chlorophacinone has the empirical formula C23H15CI03 and the molecular weight of 364.8. It forms a yellow crystalline solid with a melting point of 140°C. It is slightly soluble in water (100 mglliter at 20°C) and soluble in acetone, ethanol, and methanol. It is stable under normal storage conditions and noncorrosive. History, Formulation, and Uses Chlorophacinone is an anticoagulant rodenticide used to control rats, mice, voles, and other wild rodents. It is formulated as ready-to-use baits based on whole, cracked, or milled grain at concentrations of active ingredient ranging from 0.005 to 0.25%. It can also be used as a tracking powder. An oil concentrate is also available. 83.5.5.2 Toxicity to Laboratory Animals The acute oral LD 50 is reported to be 2 mg/kg in the rat, 1 mg/kg in the mouse, and 50 mg/kg in the rabbit. In the duck, the oral LD 50 is 100 mg/kg. Chlorophacinone is of low acute toxicity to wild birds (LD 50 of 430 mg/kg). The acute dermal LD 50 in the rabbit is 200 mg/kg (Sax, 1984). It is absorbed through the skin of the rabbit; a solution of 5 mg in 2 ml of liquid paraffin applied to 100 cm 2 of shaved skin of a rabbit caused a slight reduction of prothrombin (Worthing and Walker, 1983). Administration of 15 daily doses of 2.25 mg to gray partridges produced no detectable ill effects. Chlorophacinone is not an eye or skin irritant.
Absorption, Distribution, Metabolism, and Excretion Chlorophacinone is absorbed through the gastrointestinal tract. After oral administration, 90% is eliminated in the feces within 48 hr in the form of metabolites (Hartley and Kidd, 1983). Mode of Action Chlorophacinone is an anticoagulant agent depressing hepatic synthesis of prothrombin and clotting factors VII, IX, and X. Direct damage to capillary permeability occurs concurrently. The ultimate effect of these actions is to induce widespread internal hemorrhage. In addition, chlorophacinone is an uncoupler of oxidative phosphorylation. Unlike the coumarin derivatives, chlorophacinone may causes symptoms and signs of neurologic and cardiopulmonary injury in laboratory rats, which often lead to death before hemorrhage occurs. Chlorophacinone is characterized by its long-lasting depressive action on coagulation. 83.5.5.3 Toxicty to Humans Chlorophacinone has not been used therapeutically. Only two reports of human intoxication are known. They both involve suicidal attempts. One of them concerns a 37-year-old woman who ingested about 250 ml of a 0.25% concentrate formulation (about 625 mg of chlorophacinone). Despite intensive therapy with vitamin Kl (phytomenadione, natural form of vitamin K), the anticoagulant effect of chlorophacinone persisted for at least 45 days. An interesting fact is that it was discovered during this episode that the synthetic analog of vitamin K was ineffective (Murdoch, 1983). The second report concerns a 28-year-old
83.5 Anti-vitamin K Compounds
man who ingested an unknown amount of chlorophacinonebased rodenticide. Again, the most striking feature in this case was the unusually prolonged and severe anticoagulant effect, even under adequate therapy; it required 4 weeks for prothrombin level to come back to normal (Dusein et aI., 1984). Treatment of Poisoning Intoxication by chlorophacinone is treated by massive and prolonged administration of natural vitamin K.
83.5.6 DIFENACOUM 83.5.6.1 Identity, Properties, and Usses Chemical Name 3-[3-(1, l'-biphenyl)4-yl-1 ,2,3,4-tetrahydrol-naphthalenyl]-4-hydroxy-2H -1-benzopyran-2-one is the chemical name. Structure
See Fig. 83.4.
Synonyms Difenacoum is the approved common name. Trade names include Neoxorexa® and Ratak®. The CAS registry number is 56073-07-5. Physical and Chemical Properties Difenacoum has the empirical formula C3lH2403 and a molecular weight of 444.5. It is an off-white powder with a melting point of 215-219°C. It is slightly soluble in water (less than 10 mg/liter at pH 7) and soluble in organic solvents (50 g/liter in acetone and chloroform and 600 mg/liter in benzene). History, Formulations, and Uses The rodenticidal properties of difenacoum were first described in 1975. It is formulated as a 1 gm/kg concentrate and as a ready-to-use bait containing 50 mg of active ingredient per kilogram of bait. It is an indirect anticoagulant, more potent than the early compounds. It is used to control rats and mice resistant to other anticoagulants with varying degrees of activity.
83.5.6.2 Toxicity to Laboratory Animals The oral LD 50 is 1.8 mg/kg in male rats, 0.8 mg/kg in male mice, and 50 mg/kg in female guinea pigs. The LD 50 value for oral administration in pigs is reported to be above 80 mg/kg and it is 100 mg/kg in cats. The acute dermal LD 50 is 50 mg/kg in rats and 1000 mg/kg in rabbits. The cumulative oral LD 50 in male rats over a 5-day period is 0.16 mg/kg/day. Difenacoum and brodifacoum have been suspected of being responsible for secondary toxicity in barn owls feeding on rodents poisoned by these "second-generation" anticoagulants, a phenomenon which was not seen with warfarin baits (Wenz, 1984). Biochemical Effects Like brodifacoum and warfarin, difenacoum was shown to inhibit K-dependent steps in the synthesis of clotting factors 11, VII, IX, and X, and it is suspected that
1821
coumarin anticoagulants block the vitamin KI epoxide cycle by inhibiting the vitamin KI epoxide reductase. The latter is confirmed by the observation that difenacoum and brodifacoum produce an accumulation of tritiated vitamin KI epoxide in rats and rabbits administered tritiated vitamin KI (Park and Leck, 1982). Like brodifacoum, difenacoum has a much longer duration of action than warfarin. Treatment of Poisoning in Animals Prolonged administration of vitamin K lover several weeks is the treatment of choice. Since the effect of vitamin KI is usually delayed as it only permits the formation of new prothrombin, initial treatment with transfusion of fresh-frozen plasma or a small quantity of matched fresh blood is recommended in order to provide enough protrhombin to prevent further hemorrhage (Barlow et aI., 1982; Park and Leck, 1982).
83.5.6.3 Toxicity to Humans A case of an attempted suicide in a 17-year-old girl is reported in the literature. She was admitted to hospital having ingested 500 gm of the rat bait Neosorexa or about 25 mg of difenacoum. Upon admission she had a prolonged prothrombin time. She was treated with vitamin Kl for 45 days. The clotting activity returned to normal 30 days after the beginning of treatment (Barlow et aI., 1982). The coumarin anticoagulant difenacoum was detected by HPLC with multiwavelengh ultraviolet detection in plasma from a 41-years-old man who presented with a severe deficiency of vitamin K-dependent clotting factors of unknown aetiology. Plasma concentrations of difenacoum declined from 0.97 to 0.11 mgl- l in 47 days with a terminal half-life of 11.7 days. Subsequently, plasma concentrations of difenacoum and descarboxyprothrombin unexpectedly increased. Seven months after exposure, clotting times were still prolonged. The patient continued to have episodes of epistasix, haematoma, purpurae, and bruising and he required frequent treatment with fresh-frozen plasma in additional to oral phylloquinone (200 mg/days-l). Intermittent and unexpected increases in plasma concentrations of difenacoum and descarboxyprothrombin suggested that covert, repeated ingestion of the anticoagulant was the most likely cause of the poisoning. The measurement of low concentrations of plasma phylloquinone except following supervised ingestion of the vitamin indicated that as an outpatient, the subject was not compliant with treatment despite his protestations on the contrary. He continued to deny this even when confronted by laboratory findings and at no time did he ever admit to self-poisoning (McCarthy et aI., 1997). Treatment of Poisoning In case of threatening hemorrhage transfusion of fresh blood of fresh-frozen plasma is the initial step. Intravenous and oral administration of vitamin KI for a prolonged period of time (several weeks) with regular monitoring of coagulation is necessary.
1822
CHAPTER 83 Rodenticides
83.5.7 BROMADIOLONE 83.5.7.1 Identity, Properties, and Uses Chemical Name 3-[3-bromo[ 1,1 '-biphenyl-4-yl)-3-hydroxy-1-phenylpropyl]-4-hydroxy-2H -1-benzopyran-2-one is the chemical name. The CAS registry number is 28772-56-7. Structure
83.5.7.3 Toxicity to Humans It seems that so far no cases of human poisoning by bromadi-
olone have been reported in the literature. Treatment of Poisoning Intoxication by bromadiolone, like with other second-generation anticoagulant rodenticides, requires massive and prolonged administration of vitamin K.
See Fig. 83.3. 83.5.8 DIFETHIALONE
Synonyms Bromadiolone is the approved common name (BSI, E-ISO, F-ISO). Trade names include Deadline®, Lanirat®, Maki® and SuperCaid®. Physical and Chemical Properties Bromadiolone has the empirical formula C30H23Br04 and a molecular weight of 527.4. The technical material is a yellowish powdered mixture of two diastereoisomers of a minimum purity of 97% and a melting point of 200-21O°C. Its water solubility is in the order of 20 mg/l at 20°e. Solubility in organic solvents is (20°C) 730 g/l for dimethyl formamide, 25 mg/l for ethyl acetate, 8 mg/l for ethanol. It is stable under normal storage conditions. History, Formulations, and Uses Bromadiolone is a secondgeneration anticoagulant of the hydroxy-4-coumarin that was patented in 1967 for the control of commensal rats and mice, including those resistant to warfarin and first-generation anticoagulants, voles, and water voles. It is formulated as ready-to-use cereal and paraffine based baits containing 0.005% bromadiolone. 83.5.7.2 Toxicity to Laboratory Animals Single dose acute toxicity of bromadiolone is in the same order of magnitude as other second-generation anticoagulants: in the wild rat (Rattus norvegicus) 1.1-1.8 mg/kg and in the mouse (Mus musculus) 1.75 mg/kg (Buckle et aI., 1972). The toxicity to nonrodent species has been reported to be 0.3 mg/kg in the rabbit, 25 mg/kg in the cat, 0.15 to 1.0 in the dog, and 0.5 to 2.0 in the pig (Meehan, 1984). Birds appear to be somewhat less susceptible with acute oral LD 50s of 1600 mg/kg for quails (Tomlin, 1994). The acute percutaneous LD 50 is reported to be 2.1 mg/kg in the rabbit (Tomlin, 1994).
83.5.8.1 Identity, Properties, and Uses Chemical Name 3-[3-(4'-bromo[1,l'-biphenyl]-4-yl)-1,2,3, 4-tetrahydro-1-naphthalenyl]-4-hydroxy-2H -l-[benzothiopyran-2-one] is the chemical name. The CAS registry number is 104653-34-1. Structure
See Fig. 83.3.
Synonyms Difethialone is the approved common name (BSI). Trade names include Frap®, Baraki®, and Operats Plus®. Physical and Chemical Properties Difethialone has the empirical formula C31H23Br02S and a molecular weight of 539.5. It forms a whitish powder with a vapor pressure of 0.074 mPa at 25°e. It is practically un soluble in water: 0.39 mg/l at 25°e. Its solubility in ethanol, methanol, hexane, chloroform, and acetone in mg/l at 25°C is respectively 0.7, 0.47, 0.2, 40.8, and 4.3. The octanol/water partition coefficient is 1.41 x 105 . It is stable at temperatures up to 230°e. It is highly sensitive to phytolysis in aqueous solutions. It is strongly adsorbed in soils. History, Formulations, and Uses Difethialone was introduced as a second-generation anticoagulant rodenticide in 1986 for the control of commensa1 rats and mice including those resistant to firs-generation anticoagulants. It is formulated as ready-to-use whole grain cereals and husked oat grain baits containing 0.0025% active substance. 83.5.8.2 Toxicity to Laboratory Animals
Toxicity to Wildlife A recent survey of the effects of anticoagulant rodenticides on nontarget wild species in France has shown that this category of compounds is responsible for a very limited number of identified causes of death in most species: 1 to 3% (Bemy et aI., 1997).
The acute single dose oral LD 50 of technical grade difethialone (97.6% purity) has been reported as 0.56 mg/kg for rats, 1.29 mg/kg for mice, 4 mg/kg for dogs, and 2 to 3 mg/kg for pigs. The percutaneous acute LD 50 was determined in rabbits at 5.3 mg/kg, 6.5 mg/kg in male rats, and 5.3 mg/kg for females. By inhalation in rats exposed for 4 hours, the LC 50 is between 5 and 19.3 mg/m3. It is not irritant to the rabbit skin and only slightly irritant to the eyes. It is not a skin sensitizer. It has no mutagenic or teratogenic potential. The only effects seen in a 90-day feeding study in the rate were those expected on blood coagulation.
Mode of Action Bromadiolone is an anticoagulant with the same mechanism of action as the other second-generation rodenticides.
Absorption, Distribution, Metabolism, and Excretion In rats treated by oral administration difethialone has a short halflife in blood and a longer half-life in the liver. It is essentially
83.6 Vitamin D-Related Compounds
eliminated in the feces as the unchanged parent compound, indicating a very limited metabolization (Tomlin, 1994).
Mode of Action Difethialone is an anticoagulant with the same mode of action as the other second-generation compounds. 83.5.8.3 Toxicity to Humans It seems that so far no cases of human poisoning by difethialone have been reported in the literature.
Treatment of Poisoning Intoxication by difethialone, like with other second-generation anticoagulant rodenticides, requires massive and prolonged administration of vitamin K.
83.6 VITAMIN D-RELATED COMPOUNDS 83.6.1 ERGOCALCIFEROL 83.6.1.1 Identity, Properties, and Uses Chemical Name 9,1 0-Secoergosta-5, 7, 1O( 19),22-tetraen3-01 is the chemical name. Structure
See Fig. 83.5.
Synonyms The common name calciferol is approved by BPC (British Pharmacopeia Commission); ergocalciferol is approved by USP (U.S. Pharmacopeia). It is also known as vitamin D2, but for safety reasons it is prohibited in some countries to mention the identity of these compounds as vitamin D on rodenticides labels [Food and Agriculture Organization (FAO), 1979]. Other names include activated ergosterol, Derat Concentrate®, Deratol®, and Hi-Deratol®. The trade name is Sorexa c.R.® for a combination of calciferol and warfarin. Physical and Chemical Properties The empirical formula is C28H440 and the molecular weight is 396.63. It forms colorless
prismatic crystals. The melting point is 115-118 D C. It is insoluble in water and soluble in most organic solvents; solubility at 7D C is 69.5 g/l in acetone. It is slightly soluble in vegetable oils. Deterioration of pure, crystalline vitamin D2 is negligible after storage for 9 months in evacuated amber ampules at refrigerator temperature. However, calciferol tends to decompose in the presence of air and moisture. The stability of corn oil solutions is, however, satisfactory (Greaves et aI., 1974).
History, Formulations, and Uses The rodenticidal properties of calciferol were described by Greaves et al. (1974). The commercial rodenticide was introduced in the United Kingdom in 1974 as a combination of calciferol and warfarin formulated as a ready-to-use bait on canary seed (1 gm of calciferol and 250 mg of warfarin per kilogram). It is also available as an oil concentrate (20 gmlliter). Calciferol is used as a rodenticide for control of commensal rats and mice. Toxicity tests with calciferol combined with warfarin suggest that an additive effect between the compounds could exist. One interesting advantage of calciferol is that it is toxic to warfarin-resistant rodents. A second advantage is that it kills more rapidly-within 1 week instead of 1-3 weeks that are often required with anticoagulants (Greaves et aI., 1974). 83.6.1.2 Toxicity to Laboratory Animals Basic Findings The acute oral LD 50 of calciferol is 56 mg/ kg for rats and 23.7 mg/kg for mice. When administered daily for 5 consecutive days to laboratory rats, the LD 50 falls to 7 mg/kg/day. Lethal doses have also been reported by Gill and Redfern (1979) for the multimammate rats Mastomys natalensis; they range from 78 to 107 mg/kg (mean 96 mg/kg) in males and from 108 to 137 mg/kg (mean 119) in females when administered for 1 days at a concentration of 0.1 % in bait. Similar values were reported by Greaves et al. (1974) for wild rodents. When calciferol was given at doses of 100 mg/kg by stomach tube on one or more days, laboratory rats and mice became visibly ill within 3 days. The clinical signs are characterized by loss of appetite, listlessness, piloerection, hunched position, absence of reaction to external stimuli, weight loss, priapism, and frequent micturition (Gillman et aI., 1960).
Ergocalciferol
Figure 83.5
1823
Two fonns of vitamin D used as rodenticides.
.
ChQIecalciferol
1824
CHAPTER 83
Rodenticides
Absorption, Distribution, Metabolism, and Excretion The action of calciferol is to raise blood calcium levels by stimulating the absorption of calcium from the intestine and mobilizing skeletal reserves. This mechanism is slow and takes many hours to build up an effective level; the period of latency between the ingestion of calciferol and the development of hypercalcemia and occurrence of lethal lesions is of the order of several days, usually 4 or 5 (Greaves et aI., 1974). Calciferol has a long biological half-life in mammals and hypercalcemia induced by overdosage may continue for 6-9 months (Buckle et aI., 1972). The various tissues of the body store vitamin D for varying periods. The depletion of stores in the body is caused partly by its fecal excretion and partly by its destruction in the body. Biochemical Effects Calciferol is the most important factor for the optimal absorption of calcium. It is responsible for the synthesis of a protein that binds calcium in the intestinal mucosa, especially in the duodenum. In the absence of calciferol, calcium cannot be absorbed. Effects on Organs and Tissues All forms of vitamin Dare toxic when given in sufficiently large amounts. Excessive doses of calciferol mobilize the phosphorus and calcium from the tissues, thus broadly having an opposite effect to normal doses. The soft tissues tend to become calcified while the bone tends to be rarefied. The soft tissues most affected are the renal tubules and the media of the small renal arteries and of the large vessels, especially the aorta. The bronchi, lungs, heart and coronaries, and the stomach also are affected. In dogs, there is atrophy of the testes and the prostate while the parathyroids are smaller than normal. The histochemical effects of vascular injuries induced by calciferol orally administered to male and female Wistar rats for 5 consecutive days at doses ranging from 25,000 to 150,000 IV (1 mg of calciferol is equivalent to 40,000 IV) were studied by Gillman et al. (1960). By day 15 of the experiment-that is, day 10 after the last day of calciferol administration-necrosis of the spleen was observed in many rats that died or were sacrificed. Among the survivors, there were indications that the damaged spleen had been regenerated completely by day 25. The heart was severely injured very early in the experiments. Similarly, the coronary arteries were observed showing early dilation, injury to the internal elastic membranes, and associated early calcification. These changes were associated with sclerotic repair. In the aorta, there was an apparent relationship between the intensity of the reaction and the doses of calciferol received by the rats. By day 20, a very large concentration of calcium accumulated in the aorta (13-14% calcium compared to 22% in the femur). Similar observations were made by Grant et al. (1963) who also showed that this phenomenon was a three-step process involving early widespread alterations in many organs and tissues, followed by spontaneous recurrent resolution and reappearance of calcification for many months, even after a single episode of acute intoxication. These authors
also showed distinct differences in the time of onset and rate and extent of calcification of various tissues. Nikodemusz et al. (1981) have shown that in six male and six female common vole (Microtus arvalis), single acute oral dosing with calciferol (80-620 mg/kg) induced morphological changes representing varying degrees of parenchymal degeneration and calcification in the kidneys, lungs, and heart. Calcium deposits were observed in the esophageal and gastric mucosae, as well as in the aortic media. These lesions were not observed in the group receiving 80 mg/kg of calciferol. The approximate lethal dose by the oral route was estimated as 120 mglkg in males and 280 mglkg in females. The survival time ranged from 53 to 94 hr. Tarrant and Westlake (1984) have also shown that feeding laboratory-reared male rats (Rattus norvegicus) and male and female quail (Coturnix coturnixjaponica) a diet containing 0.1 % calciferol (the recommended field concentration) for 2 days induced in both species a similar pattern of calcium deposits in the kidneys, beginning on the second day after the animals were returned to a standard, calciferol-free diet. Effects on Reproduction Adult female New Zealand White rabbits had been intramuscularly treated with ergosterol in cottonseed oil in divided doses every other day for a total of 1.5 million IV. Three other groups of five rabbits each received intramuscular injections of ergosterol daily for the duration of the gestational period for a total of 2.5-3.5 million and 4.5 million ID. At autopsy all the females given 2.5 million IV and above died spontaneously within 65 days after their first injection of calciferol and all that were pregnant aborted during the first 12 days of pregnancy or delivered macerated fetuses. All aortas had various degrees of pathological changes, including those from females treated with 1.5 million IV. A total of 14 abnormalities of the aorta were observed in the 34 offspring whose mothers had been treated with excessive levels of ergocalciferol. Aortic lesions that appeared similar to the supravalvular aortic stenosis seen in humans were noted in six rabbits. The blood levels of ergocalciferol in the mothers and their offspring were seven and nine times greater than those in the control classes and their offspring, respectively, indicating that transplacental passage occurred (Friedman and Roberts, 1966). A similar experiment was carried out by Friedman and Mills (1969) on 15 pregnant New Zealand White rabbits given divided doses of intramuscular ergocalciferol every other day, starting on the second day of insemination and throughout the pregnancy for a total of 750,000 IV. Characteristic malformations were observed in the offspring. These were represented by premature closure of cranial structures, small skulls compared with the controls, and narrowing of the body of the mandible. The maxillary and mandibular central incisors showed severe enamel hypoplasia. Some cases of anodontia and abnormal palatal shape were noted. Similar craniofacial malformations are frequently associated with supravalvular aortic stenosis in children. Several additional studies of the effects of vitamin D on reproduction and fetal development in rats and rabbits have been reported. Latore (1961) found that excess ergocalciferol
83.6 Vitamin D-Related Compounds reduces fertility in rats when administered on days 0-7 of gestation but not when administered on days 8-21. Nebel and Ornstein (1966) have demonstrated that ergocalciferol administered to rats (Rattus rattus) in daily doses of 20,000 IV for 1-3 weeks significantly affected the genital cycle, fertility, and early pregnancy from both the morphological and functional points of view in direct relation to the beginning and duration of ergocalciferol administration. Ornoy et al. (1968) showed that vitamin D2 (ergocalciferol) given to pregnant albino rats at daily doses of 4000, 20,000, or 40,000 m from day 9 to days 21 of pregnancy crosses the placental barrier and induces alterations of the mineral composition in the fetal bones at the 40,000 IV level only. In the same treated group, placentas, fetuses, and fetal bones were found to be smaller than those in the untreated groups. However, it seems from the experimental data that pregnant rats are more tolerant to rather high levels of ergocalciferol than nonpregnant females (Ornoy et aI., 1968; Potvliege, 1962). Factors Influencing Toxicity In commercially available preparations, calciferol is often associated with warfarin in a 4: 1 ratio because it has been shown that this mixture produced a marked increase in mortality in Norway rats (Greaves et aI., 1974), thus suggesting an additive effect of the two compounds. At the LD 50 level, toxicity of the mixture is intermediate between the toxicities of calciferol and warfarin. Treatment of Poisoning in Animals It seems that the only case of poisoning of a domestic animal reported is that of a 4-year-old boxer dog seen several days following ingestion of a warfarin-calciferol mixture. Despite treatment with vitamin K, antibiotics and vitamins, the end of the tongue became necrotic and had to be removed and large areas of the skin were also necrosed, indicating generalized vascular damage. The dog made a slow recovery (Edlin, 1982). It is likely that treatment with vitamin K had prevented the hemorrhagic manifestations of warfarin toxicity from showing but that the vascular injuries due to calciferol were responsible for the necrotic effects. 83.6.1.3 Toxicity to Humans Therapeutic Use As a vitamin, calciferol is used to prevent and to cure rickets, tetany, spasmophilia and osteoporomalacia. Several decades ago, vitamin D was recommended for the treatment of lupus vulgaris (skin tuberculosis limited to the face); massive doses were used, sufficient to provoke a mild degree of hypervitaminosis. The purpose was to induce calcification of the subcutaneous lesions in order to stop their progression. In the case of rickets, the recommended preventive dose was 5001500 m/day, and the curative dose was 1000-3000 m/day; it was estimated that 10,000 IV represents a toxic dose. The minimal toxic overdose does not appear to be many times greater than the optimum curative dose (Harris, 1955). Many cases of vitamin D2 as well as D3 poisoning in humans have been reported. The oldest ones were analyzed by Bicknell and Prescott (1953). More recent cases have been reported, all
1825
related to accidental or inadvertent overdosing. A 69-year-old women with hypoparathyroidism was treated with a twiceweekly dose of 50,000 m of calciferol. However, for 4 weeks before admission to hospital she had mistakenly taken 300,000 IV (7.5 mg) a day. On examination she was lethargic and mentally confused, with muscular hypotonia. In this report, two other similar cases concerning elderly women are reported; one had received 100,000 m (2.5 mg) calciferol orally on alternate days for 3 months and the other had received 400,000 m (10 mg) a day by mouth and 600,000 IV of calciferol once a week by injection during the 2 months before admission to hospital. All three patients had elevated serum calcium. All three of them were treated with intravenous injections of porcine calcitonin, which caused serum calcium to fall back to a normal level within 2-3 days (Buckle et aI., 1972). Davies and Adams (1978) have also reported eight cases of severe vitamin D poisoning. In six patients the therapy was unnecessary, and in two others inadequate supervision of treatment resulted in overdosage. Paterson (1980) reported 21 cases of hypercalcemia due to vitamin D poisoning; among them two patients died while intoxicated. Overall, the clinical symptoms induced by chronic poisoning include, in various associations and intensity, loss of appetite, loss of weight greater than would be expected from the loss of appetite, nausea, vomiting, and constipation or diarrhea. Abdominal pain may be so severe as to lead to unnecessary laparotomies. Headaches are usual and one special form has been noticed. This is a tightness across the back of the head which goes on to acute sensitiveness of the scalp. Mental confusion and loss of memory may also be seen. Epileptiform fits are a rare complication. Metastatic calcification has been described along with vascular and renal calcification. An unusual, albeit typical case has been reported by Cohen et al. (1979): a case of a deafness due to a long-term overdo sage for the treatment of pseudohypoparathyroidy (2.5 mg of calciferol daily for 4 years). The patient had a 3-month history of deafness, weight loss, anorexia, and weakness. She had extensive calcification of the tympanic membranes, corneas, kidneys, and blood vessels. She had also, of course, high serum calcium, which was successfully treated with calcitonin and prednisolone and a lowcalcium diet. She was discharged from hospital symptom-free except for the deafness. Hypercalcemia is the earliest sign of vitamin D overdosage. It was suggested at one point that increased intake of vitamin D through consumption of fish liver in Norway might be correlated with an increased probability of myocardial infarction (Linden, 1974). However, this hypothesis, based on a retrospective study of a number of patients with myocardial infarction, angina pectoris, and degenerative joint diseases, was later criticized because of serious bias and shortcomings in the methodology (Lindahl and Lindwall, 1975). A prospective study, performed in the same area of northern Norway, has not confirmed the existence of a higher risk of myocardial infraction related to vitamin D intake or status of the population (Vik et aI., 1979). Use Experience and Dosage Response There is no known report of accidents occurring with calciferol used as a roden-
1826
CHAPTER 83
Rodenticides
ticide. It is difficult to estimate the minimum toxic dose in humans; however, from the reported cases of overdosage, it would seem that 0.15 mg/kg/day for 3 weeks may lead to clinical signs of poisoning. Treatment of Poisoning Treatment of acute and chronic overdosages with calciferol requires that hypercalcemia be brought down to a nonnallevel quickly; intravenous injection of calcitonin is the specific treatment to be applied under close monitoring of the serum calcium level. Steroid therapy is also often effective but slower, nonnocalcemia being achieved after 5-7 days. Other methods are available but they are not devoid of problems: Chelation with sodium ededate is only temporary in its effect and is nephrotoxic. Sodium phosphate is also effective but may be responsible for metastatic calcifications (Buckle et aI., 1972). One case of vitamin D2 poisoning in an elderly woman was successfully treated by inducing the hepatic microsomal enzymes with 500 mg/day of glutethimide. The mechanism of action is still unknown (Iqbal and Taylor, 1982), but this approach is rather slow since the serum calcium level did not fall back to nonnal values until day 12 of treatment. 83.6.2 CHOLECALCIFEROL 83.6.2.1 Identity, Properties, and Uses Chemical name 9, 1O-secocholesta-5,7, 1O( 19)-trien-3-betaol is the chemical name. Structure
See Fig. 83.5.
Synonyms Some synonyms are activated 7-dehydrocholesterol, oleovitamin D3, cholecalcifero natural vitamin D3. The trade name for the rodenticide is Quintox®. Physical and Chemical Properties The empirical formula is C27H440 and the molecular weight is 384.62. It fonns tine needles. The melting point is 84-86°C. It is practically insoluble in water, soluble in the usual organic solvents, and only slightly soluble in vegetable oils. It is oxidized and inactivated by moist air within a few days. However, the fonnulated product is stable over 1 year at ambient temperatures in sealed packages. History, Formulations, and Uses Cholecalciferol is formulated for use ad a rodenticide as a grain bait containing 750 ppm (0.075%) of active ingredient and commercialized under the trade name Quintox®. It is a single-feeding and multifeeding rodenticide used for controlling anticoagulant-resistant rats and mice. There is a period of time between feeding and death similar to but perhaps shorter than the observed with anticoagulant rodenticides. The rodenticidal activity of cholecalciferol is comparable to that of ergocalciferol. 83.6.2.2 Toxicity to Laboratory Animals The two fonns of calciferol are equally toxic to most mammals. The acute oral LD 50 of cholecalciferol is 43.6 mg/kg for Rattus
norvegicus and 42.5 mg/kg for mice (Mus musculus). In the dog, the oral LD 50 is 88 mg/kg. However, deaths have been observed with doses of 10 to 20 mg/kg (El Bahri, 1990). Absorption, Distribution, and Excretion Cholecalciferol after absorption from the intestine is transported to the liver, where it is metabolized to 25-hydroxycholecalciferol by an NADPH-dependentreaction. This metabolite is then transferred to the kidney and converted to 24-, 25-, or 1,25-dihydrocholecalciferol by mitochondrial mixed-function oxidases (McClain et aI., 1980). After their intestinal absorption, ergocalciferol (vitamin D2) and cholecalciferol (vitamin D3) undergo an identical metabolic C pathway (Fournier et aI., 1985). The metabolism and phannacokinetics of one metabolite of cholecalciferol, 24,25-dihydrocholecalciferol, have been reviewed by Jarnargin et al. (1985); the excretion curve shows an initial fast phase with a plasma half-life of 0.55 hr and a second slow phase with a plasma half-life of 73.8 hr in the rat. The clearance from plasma, liver, and kidney but not intestine follows a two-compartment model. The most potent form of vitamin D3, 1,25-dihydrochlolecalciferol, has been shown to be responsible for the stimulation of intestinal absorption of calcium and the metabolism of calcium in bone. However, Frolick and Deluca (1973) have shown that when given orally to rats, 1,25dihydrocholecalciferol is rapidly modified during its passage through the intestine, thus reducing its physiological activity to a large extent. Effects on Organs and Tissues Moderately excessive doses of cholecalciferol, 100, 300, 2000, and 4000 IU/kg of feed, given for 4 months to experimental Yorkshire pigs rapidly produced gross arterial lesions (fibromuscular interstitial thickening of the coronaries, especially at the branching sites). Macrophages, plasma cells, and mast cells were observed to accumulate in the subendothelial space. The extent of these lesions, resembling those commonly seen in atherosceloris in humans, was more or less dose-related (Toda et aI., 1985). Since both fonns of vitamin D (ergocalciferol and cholecalciferol) follow the same metabolic pathway, it is expected that they are responsible for inducing the same lesions. Effects on Reproduction Calcitriol (l,25-dihydrocholecalciferol), which is the most biologically active metabolite of cholecalciferol, has been administered to pregnant rats and rabbits at daily doses of 0.02-0.08 and 0.30 ).1.g/kg from days 7-15 of gestation in the rat and from days 7-18 in the rabbit. In rats no adverse effect on fertility, litter parameters and offspring was observed. Hypercalcemia and hypophosphatemia were observed in pregnant rats of the middle an high-dose groups, as well as hypercalcemia in pups. Calcitriol induced maternal mortality and fetotoxicity in rabbits treated with 0.3 ).1.g/kg/day. Two litters at the highest dose and one litter at the middle dose levels contained fetuses with multiple abnormalities (McClain et aI., 1980).
83.7 Miscellaneous Synthetic Organic Rodenticides
83.6.2.3 Toxicity to Humans Experimental Oral Exposure The active metabolite 1,25dihydrocho1ecalciferol was administered either orally or intravenously to four healthy volunteers and three patients with hypoparathyroidism. After an oral dose, the highest serum concentration of radioactivity was reached after 4 hr. The route of administration had little apparent effect on the serum concentration or on the rapid phase of elimination, but the slow phase of excretion was longer after oral administration. The highest urinary excretion rate was observed during the first 24 hr and was little affected by the route of administration. The half-life of 1,25-dihydrochole-calciferol was 3-5 days. On average, 40% of the dose of 1,25-dihyrocholecalciferol was excreted within 10 days (Mawer et aI., 1976). Therapeutic Use Cholecalciferol being the natural form of vitamin D, the therapeutic uses are those outlined for ergocalciferol. Accidents and Use Experience These are the same as those mentioned for ergocalciferol. There is no known report on an accident specifically attributable to cholecalciferol used as a rodenticide. However, intoxication from overdosage during therapeutic use of vitamin D3 is well known and signs and symptoms have been well studied (Navarro et al., 1985).
83.7 MISCELLANEOUS SYNTHETIC ORGANIC RODENTICIDES As far as is known, the miscellaneous synthetic organic rodenticides (see Fig. 83.6) are unrelated to one another or to other groups of pesticides either pharmacologically or chemically. The two that have been studied in humans are remarkable. Chloralose is an anesthetic, albeit one with the property of inducing myoclinic seizures. With the possible exception of this compound, anesthetics have nothing to offer for killing vertebrate pests. Far too many animals become anesthetized before consuming a fatal dose, and they later recover. In fact, the possibility of recovery seems to be taken into account in the British practice of recovering birds affected by the compound. Norbomide is of interest as one of the most selective poisons known. Its limitation is not any lack of toxicity to species of the genus Rattus but the problem of secondary bait refusal (Greaves, 1966). 83.7.1 CHLORALOSE 83.7.1.1 Identity, Properties, and Uses Chemical Name 1,2-0 -(2,2,2-trichloroethylidene )-a- D-glucofuranose is the chemical name. Structure
See Fig. 83.6.
If
H-C-O
11
1827
~
fH-CCl 3
c-o I
HO-C-H
~
-
I
H-C-OH I H-C-OH
I
H-C-OH
I
H (X-Chloralose Figure 83.6
OH
I
'c
00
NH 0
Norbormidc
Miscellaneous synthetic organic rodenticides.
Synonyms Chloralose (BSI) is the common name in use. Other nonproprietary names include a-chloralose, a-D-glucochloralose, anhydroglucochloral, chloro-alosane, and glucochloral. Trade names include Alphakil® and Somio®. The CAS registry number is 15879-93-3. Physical and Chemical Properties Chloralose has the empirical formula CSHll C1306 and a molecular weight of 309.54. It forms a crystalline powder that melts at 187°C. It is soluble in ether and glacial acetic acid, slightly soluble in chloroform, and almost insoluble in petroleum ether. Its solubility in water at 15°C is 0.44%. Chloralose reduces Fehling's solution only after prolonged heating. It is hydrolyzed into its two components by acids. History, Formulations, and Uses Chloralose has been in use in Europe for many years. Its narcotic properties are employed to immobilize depredating birds and render them easier to kill by other means. Baits contain about 1.5% of the compound. It has been reported that chloralose is also used on seed grain as a bird repellent. It is used against mice in baits of up to 40 gmlkg. All baits should contain a warning dye. Chloralose is used in medicine as a soporific and formerly was used as an anesthetic. 83.7.1.2 Toxicity to Laboratory Animals The oral LD 50 values of chloralose in rats and mice are in the range of 300-400 mg/kg. Cats are more susceptible (100 mg/kg) and dogs more resistant (600-1000 mg/kg) (Cornwell, 1969). The compound is more toxic to many birds than to mammals. Oral LD 50 values have been determined for the starling (75 mg/kg), redwing blackbird (32 mg/kg), yellow-headed blackbird (133 mg/kg), crow (42 mg/kg), pigeon (178 mg/kg), house finch (56 mg/kg), house sparrow (42 mg/kg), mallard duck (42 mg/kg), mourning dove (42 mg/kg), and whitecrowned sparrow (56 mg/kg) (Schafer, 1972). Hanriot and Richet, who advocated chloralose as a soporific, found that dogs survived oral dosages as high as 610 mg/kg but were killed by dosages of 660 mg/kg and greater. Cats survived oral dosages of 65 mg/kg or lower but one was killed by a
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CHAPTER 83
Rodenticides
dosage of 71 mg/kg and all were killed by 140 mg/kg or more. Dogs survived intravenous injection of 120 mg/kg or less but were killed by 150 mg/kg. A dosage of 12.5 mg/kg produced severe symptoms in a cat (Hanriot and Richet, 1897). After some delay, animals poisoned by chloralose show incoordination, vertigo, tremor, and failure to recognize objects. The sense of pain is lost but there is increased reactivity to touch, sound, or electric shock. If stimulated, the animals respond reflexively and with full force. If artificial respiration is withheld, animals that have received a sufficient dosage die of respiratory failure (Hanriot and Richet, 1897). Thus it was recognized very early that although response to chloralose is somewhat similar to that to chloral, it is also similar to the response to strychnine in the sensitization to external stimuli. Chloralose is metabolized to chloral, CH(OHh-CCI3 (Cornwell, 1969), oxidized to trichloroacetic acid, and reduced to trichloroethanol (Marshall and Owens, 1954; Owens and Marshall, 1955). The latter metabolite is responsible for much of the hypnotic effect of chloral hydrate; all tissues studied so far are capable of forming it from chloral hydrate (Butler, 1949). Trichloroethanol combines with glucuronic acid in the liver to form the pharmacologically inactive urochloralic acid, which is readily excreted in the urine (Lees, 1972). Chloralose was not tumorigenic in two strains of mice that received it at the highest tolerated level for 18 months (Innes et aI., 1969). Treatment of Poisoning in Animals Administration of analeptic drugs or stimulants of the central nervous system such as methylamphetamine (0.5-4 mg/kg of body weight, orally or intramuscularly) or ephedrine (2.5 mg/kg of body weight subcutaneously) has been recommended and successfully applied in poisoned dogs (Bennett, 1972; Smith and Boyd, 1972), although his had already been criticized (Shepherd, 1971) on pharmacological and biochemical grounds. In addition, supportative therapy to correct any hypothermia and respiratory problems may be indicated in severely poisoned animals. 83.7.1.3 Toxicity to Humans Therapeutic Use Chloralose was introduced in 1888 and 1893 as an anesthetic and soporific. Its anesthetic use was soon dropped, presumably because an effective dosage tended to cause excessive muscular activity. However, this same feature was regarded as an advantage in connection with a soporific. Thus Sollman (1901) considered chloralose preferable to chloral except for insomnia due to exaggerated reflex irritability. He pointed out that chloralose is a stronger hypnotic, heightens the reflexes, has less action on the heart, and produces practically no local irritation. Chloralose has gone out of fashion in the United States because its action is somewhat delayed compared to that of chloral (Sollman, 1942) and perhaps because activation of reflexes was not considered an appropriate property of a soporific. Accidental and Intentional Poisoning Most reported cases of poisoning by chloralose have occurred in France, where the
compound is used medically as a soporific and sedative and also is used as a poison to kill crows and rats. Tempe and Kurtz (1972) listed 60 brands of poison based on chloralose, of which 31 also contained ANTU (see Section 83.4.1.3). In their series of 22 acute intoxications by chloralose, only one was caused by the compound intended for use as a drug; the others were caused by rat poisons. Most cases of poisoning have involved attempted-generally unsuccessful-suicide. A few cases of mild accidental poisoning of children have been recorded (Gaultier et aI., 1962). The characteristic effect of chloralose is coma, which may bc preceded by vomiting, vertigo, trembling, and a sensation of inebriation. In massive intoxication, the coma may appear in some minutes but usually appears in one to several hours after ingestion of chloralose (Favarel-Garrigues and Boget, 1968; Tempe and Kurtz, 1972). The patient may be calm and limp or may be agitated. Cornette and Franck (1970) saw only cases in agitated coma with varying degrees of myoclonia. The myoclonia was always reinforced by stimulation and it occurred predominantly in the arm or leg that was stimulated. There was a bilateral, symmetrical seizure, occasionally with hypersalivation and incontinence of urine. However, tonic spasm or tonic clonic convulsions were not seen. These authors never encountered a case of hypotonic or "calm" coma seen by some others. The attributed the difference to better treatment, but it would seem difficult to exclude differences in dosage as a cause. Favarel-Garrigues and Boget (1968) specifically noted that this form of coma usually but not always occurred in massive intoxications, and one almost always saw hyperreactivity, clonic jerks, and a state of agitation in such cases during recovery. Tempe and Kurtz (1972) reported similar experience. The reflexes are active in agitated coma. In 9 of the 22 cases reported by Tempe and Kurtz (1972), 6 showed a positive Babinski test bilaterally. Reflexes are diminished or absent in massive intoxication. Authors have not agreed on whether the most severe seizures caused by chloralose are truly epileptic (Moene et aI., 1969) or are bilateral, synchronous myoclonic disturbances (Cornette and Franck, 1970). That the latter view is correct seems to depend not on any difference in clinical severity but on a lack of correlation of the EEG with the physical disturbance. Hypersecretion of the respiratory tract apparently is the most life-threatening aspect of intoxication. It may occur in the absence of ANTU but is more common and more severe when ANTU is involved. This hypersecretion may appear early, suddenly, and severely, but it also disappears in less than an hour. Unlike pulmonary edema, the secretion is not sanguineous and it is poor of albumin. The mild appearance on X-ray contrasts with the clinical severity (Favarel-Garrigues and Boget, 1968). Even in the apparent absence of injection, the temperature may be elevated as high as 41 QC (Moene et aI., 1969). Reawakeing requires several hours in mild poisoning, 1214 hr in severe poisoning, and as much as 96 hr in massive poisoning. The return of consciousness may be gradual but usually is sudden and may be accompanied by headache, stiffness, and weakness. Recovery without sequelae is the rule (Favarel-
83.7 Miscellaneous Synthetic Organic Rodenticides Garrigues and Boget, 1968; Tempe and Kurtz, 1972). Moene et ai. (1969) reported a death caused by uncomplicated poisoning, but it is indicative of the toxicity of chloralose that the victim succeeded in his suicide with this compound only on his third attempt within a period of about 4 months. Death due to circulatory collapse occurred on the evening of the fifth hospital day. Dosage Response Eleven patients, who were thought to have taken doses ranging from 640 to 2880 mg, all survived with appropriate treatments (Cornette and Franck, 1970). In another case involving chloralose intended as a rat poison, a dose thought to be 3000-4000 mg was survived (Boudouresque et al., 1966). Other reports of nonfatal dosages have fallen within the range of 2000-9000 mg, but one kind of rat poison came in 20-gm packets and at least one person may have survived that dose. The dose in a fatal case was unknown (Moene et ai., 1969; Gras et ai., 1975). Tempe and Kurtz (1972) considered that toxic signs could result from 400 mg but that most cases resulted from ingestion of about 1000 mg. Laboratory Findings In two cases in which the dosage was known with considerable certainty, about 45% of the amount ingested was recovered from urine passed within the first 24 hr, about 90% in the form of the glucose conjugate (Gras et ai., 1975). The EEG picture in acute poisoning by chloralose is characteristic, involving slow waves and numerous spikes that usually are bilaterally symmetrical and synchrounous. The pattern is changed dramatically by diazepam, which converts it to delta activity without the rapid rhythms commonly seen with that medication. The tracing always become normal, often within 24 hr. The acute EEG record does not correspond to the myoclonic movements that, as observed visually or by EMG, are usually isolated, brief asymmetrical, and asynchronous (Boudouresque et ai., 1966; Cornette and Franck, 1970; Tempe and Kurtz, 1972). Treatment of Poisoning The myoclonic seizures respond to intravenous injection of 10 mg of diazepam, but this may have to be repeated once or twice. Recovery without sequelae is the rule (Boudouresque et ai., 1966; Cornette and Franck, 1970; Tempe and Kurtz, 1972). 83.7.2 NORBORMIDE 83.7.2.1 Identity, Properties, and Uses Chemical Name 6-(a-hydroxy-a-2-pyridylbenzyl)-7-(a-2pyridy lbenzylidene)-norbor-5-ene-2,3-dicarboximide is the chemical name. Structure
See Fig. 83.6.
Synonyms Norbormide (ANSI, BSI, ISO) is the common name for this compound. Trade names include Shoxin® and
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Raticate®. Code designations include McN-l,025 and S-6,999. The CAS registry number is 991-42-4. Physical and Chemical Properties Norbormide has the empirical formula C33H2SN303 and a molecular weight of 511.55. It is a white crystalline powder melting at 190-l98°C. Its solubility at 30°C in ethanol is 14 mg/liter; in chloroform, more than 150 mg/liter; in diethyl ether, 1 mglliter; and in 0.1 N CHI, 20 mglliter. Norbormide is stable at room temperature and to boiling. It is hydrolyzed by alkali and is noncorrosive. History, Formulations, and Uses Norbormide was introduced in 1964 by the McNeil Laboratories, Inc. It is a selective rodenticide, lethal to rats but not to other rodent species. It usually is concentrated in prepared baits of cereal at 5-10 gm/kg. Baits containing the compound should contain a warning dye. Basic Findings Norbormide shows a remarkable selectivity both in toxicity and in pharmacological effect. Oral LD 50 values for both wild and domestic Norway rats ranged from 5.3 to 15.0 mg/kg (Greaves, 1966; Niu, 1970; Roszkowski, 1965; Roszkowski et ai., 1964). Corresponding values for roof rats and Hawaiian rats were 52 and about 10 mg/kg, respectively. Oral LD 50 values were much higher in other rodents and lagomorphs-for example, hamster, 140 mg/kg; guinea pig, 620 mg/kg; mouse, 2250 mg/kg; and rabbit, about 1000 mg/kg. The oral toxicity was low in all other species tested; in the dog, cat, monkey, sheep, pig, and chicken, no effect was detectable at 1000 mg/kg (Niu, 1970; Roszkowski, 1965; Roszkowski et ai.,1964). Rats given an overdose of norbormide died within 15 min to 4 hr. At first, the animals assumed a hunched position. Later there was locomotor impairment due to weakness but not paralysis of the hind legs. Struggling, labored breathing and, in some instances, a mild convulsion preceded death (Roszkowski et aI., 1964). Many analogs have been studied and none was found as toxic to rats as norbormide (Poos et ai., 1996). Dogs survived daily doses of norbormide corresponding to a dietary level of 10,000 ppm for 15-60 days, but they lost appetite and looked ill. Dogs tolerated a dosage corresponding to a dietary level of 1000 ppm for 60 days without ill effect (Roszkowski et aI., 1964). Even in laboratory rats, susceptibility to the compounds was greatly reduced when it was mixed in the diet rather than being given by stomach tube. This may be explained in part by tolerance. The oral LD 50 determined after a I-day rest period was less than doubled in rats pretreated for 1-7 days at the rate of 2 mg/kg/day. The degree of tolerance was small, but it was statistically significant. When the rest period was 5 days, no tolerance remained (Roszkowski, 1965). In any event, primary bait refusal can be a very serious problem in the use of norbormide (Maddock and Schoff, 1967). Mode of Action and Cause of Death Some sex and species differences in susceptibility could be explained by differ-
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CHAPTER 83 Rodenticides
ences in metabolism or absorption. For example, the oral LD 50 values for male and female laboratory rats were 5.3 and 15.0 mg/kg, respectively, although the intravenous values (0.65 and 0.63 mg/kg, respectively) did not differ significantly. The very low susceptibility of mice to oral doses (LD 50, 2250 mg/kg) was due in part to poor absorption, as indicated by the intraperitoneal LD 50 of only 390 mg/kg. However, differences in absorption could not account for many observed species differences. For example, anesthetized dogs showed no detectable response to intravenous injection at the rate of 40mg/kg. Species differences in response of the peripheral blood vessels to norbormide seemed to account for most of the observed species differences in its toxicity. The compound caused an extreme, irreversible vasoconstriction in laboratory rats, and this was considered the cause of death. The effect was demonstrated by direct observation of flow experiments, or both, in the ear, eye, skin, mesentary, and heart and undoubtedly occurred in other organs. It resulted from either systemic or appropriate local administration. Only vessels of relatively small caliber were visibly constricted. Spiral rat aortic segments and duodenal strips did not respond to norbormide. The mechanism of action was best demonstrated in the heart. Isolated myocardial strips showed no loss of contraction or responsiveness when norbormide was added to the bath. However, when norbormide was injected into the aorta of isolated rat hearts, the coronary flow rate decreased greatly and the heart slowed and developed arrhythmia. The effects were not inhibited or reversed by sodium nitrite or other vasodilators. Following vasoconstriction and presumably as a result of vascular injury, an erythematous but not typically inflammatory lesion began to develop in the rat skin 6 hr after intradermal injection of 0.1 ml of 0.1 % solution. The lesion became maximal in 24-48 hr. Sometimes an area of central necrosis was observed. A concentration of 0.01 % produced the effect only inconsistently. Except in rats, vasoconstriction was not seen even at high dosage levels. Why rats respond differently remains obscure (Roszkowski, 1965). A number of the observations just discussed were confirmed by Niu (1970), who reached the same general conclusion regarding the importance of vasoconstriction in the rat. An oral or intraperitoneal dosage of 1020 mg/kg caused a doubling of blood glucose levels and a decrease of liver and muscle glycogen when coma began 0.5-2 hr after treatment. The same dosage had no effect on the glucose levels of two strains of mice, nor did it produce illness. Insulin counteracted the hyperglycemic effect or norbormide in rats but did not protect against toxic manifestations and death, suggesting that the hyperglycemia is secondary (Patil and Radhakrishnamurty, 1973). 83.7.2.2 Toxicity to Humans Because the toxicity of norbormide to different species corresponds to its ability to cause peripheral vasoconstriction in
them, it was a meaningful test to inject three volunteers intradermally with 0.1 ml of 0.1 % solution. Skin treated in this was showed no response not seen in controls (Roszkowski, 1965). In a study that has been cited for many years but apparently was first published by Hayes (1975), Dr Kazwya Kamoya of the Showa Medical School in Tokyo administered norbormide to volunteers orally at doses ranging from 20 to 300 mg. No sign or symptom was produced. It was concluded that body temperature and blood pressure decreased slightly and temporarily following the larger dose. Actually, the largest fall in temperature observed at any dose was 0.7°C, and this occurred after doses of 20 to 80 mg, whereas the largest fall after a dose of 120 mg or more was only 0.3°C. Systotic (but not diastolic) blood pressure possibly fell after 120 mg of norbormide and certainly fell after larger doses. The largest decreases recorded were from 1332176 to 100174 after 200 mg and from 120/80 to 96/80 after 300 mg. The lowest values were measured 1 hr after ingestion, and the values were essentially normal at 2 hr in each instance. It seems unlikely that poisoning by norbormide will occur. If it does, treatment must be symptomatic.
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Chakrabarti, S. K. (1978). Influence of heavy metals on the in vitro interaction between human serum albumin and warfarin. Biochem. Pharmacol. 27, 2957-2959. Chan, K. K., Lewis, R. J., and Trager, W. F. (1972). Absolute configuration of the four warfarin alcohols. 1. Med. Chem. 15, 1265-1270. Chapman, c., and Phillips, M. A. (1955). Fluoroacetamide as a rodenticide. J. Sci. FoodAgri. 6,231-232. Chappelka, R. (1980). The rat poison Vacor. Letter to the editor. N. Engl. J. Med. 302, 1147. Chenoweth, M. B. (1949). Monofluoroacetic acid and related compounds. Pharmacol. Rev. 1,383-384. Chenoweth, M. B., and Gilman, A. (1946). Studies on the pharmacology of fluoroacetate. J. Pharmacol. Exp. Ther. 87,90-103. Chenoweth, M. B., Kandel, A., Johnson, L. B., and Bennett, D. R. (1951). Factors influencing fluoroacetate poisoning. Practical treatment with glycerol monoacetate. J. Pharmacol. Exp. Ther. 102,31-49. Chi, C. H., Chen, K. W, Chan, S. H., Wu, M. H., and Huang, J. J. (1996). Clinical presentation and prognostic factors in sodium monofluoroacetate intoxication. J. Toxicol. Clin. Toxicol. 34(6),707-712. Chong, M. K. B., Harvey, D., and DeSwiet, M. (1984). Follow-up study of children whose mothers were treated with warfarin during pregnancy. Br. J. Obstet. Gynaecol. 91, 1070-1073. Chung, H. M. (1984). Acute renal failure caused by acute monofluoroacetate poisoning. Vet. Hum. Toxicol. 26,29-32. Cocks, J. R. (1960). Anticoagulants and the acute abdomen. Med. J. Aust. 1, 1138-114l. Cohen, H. N., Fogelman, I., Boyle, I. T., and Doig, J. A. (1979). Deafness due to hypervitaminosis D. Lancet 1, 985. Colamussi, v., Bonari, R., and Benini, F. (1970). Minor poisoning from fluoroethanol (description of three cases). Arcisp. S. Anna Ferrara 23, 447-458. Coldwell, B. B., Buttar, H. S., Paul, C. J., and Thomas, B. H. (1974). Effect of sodium salicylate on the fate of warfarin in the rat. Toxicol. Appl. Pharmacol. 28,374-384. Cole, E. R., and Bachmann, F. (1976). Spectrophotometric assays for warfarin sodium and dicumarol. Arch. Intern. Med. 136,474-479. Coon, W W, and Willis, P. W (1972). Some aspect of the pharmacology of oral anticoagulants. Clin. Pharmacol. Ther. 11,312-336. Cornette, M., and Franck, G. (1970). Clinical and electroencephalographical aspects of acute chloralose poisoning. Review of 11 recent cases. Rev. Neurol. 123,268-272 [in French]. Comwell, P. B. (1969). Alphakil-A new rodenticide for mouse control. Pharm. J. 202, 74-75. Correll, J. T., Coleman, L. L., Long, S., and Willy, R. F. (1952). Diphenylacetyl1,3-indandione as a potent hypoprothrombinemic agent. Proc. Soc. Exp. BioI. Med. 80, 139-143. Cullen, S. I., and Cata1ano, P. M. (1967). Griseofulvin-warfarin antagonism. J. Am. Med. Assoc. 199,582-583. Davies, J. M., Thomas, H. F., and Manson, D. (1982). Bladder tumours among rodent operatives handling ANTU. Br. Med. J. 285, 927-93l. Davies, M., and Adams, P. H. (1978). The continuing risk of vitamin D intoxication. Lancet 2, 621-623. Davi1a, J. c., Edds, G. T., Osona, 0., and Simpson, C. F. (1983). Modification of the effects of aflatoxin Bland warfarin in young pigs given selenium. Am. 1. Vet. Res. 44, 1877-1883. Deckert, F. W (1973). Warfarin metabolism in the guinea pig. I. Pharmacological studies. Drug. Metab. Dispos. 1,704-710. Deckert, F. W., Moss, J. N., Sambuca, A. S., Seigel, M. C., and SteigerwaIt, R. B. (1977). Nutritional and drug interactions with Vacor rodenticide in rats. Fed. Proc., Fed. Am. Soc. Exp. BioI. 36, 990. Deckert, F. W, Godfrey, W J., Lisk, D. C., Steigerwalt, R. B., and Udinsky, J. R. (1978a). Metabolic interactions with RH-787 (Vacor rodenticide). Fed. Proc., Fed. Soc. Exp. BioI. 37, 424. Deckert, F. W, Hagerman, L. M., Lisk, D. C., Steigerwalt, R. D., and Udinsky, J. R. (1978b). The disposition of ( 14 C)-RH-787 (Vacor rodenticide) in rodents and dogs. Toxicol. Appl. Pharmacol. 45,313.
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Rodenticides
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CHAPTER
84 Methyl Bromide Vincent J. Piccirillo NPC Incorporated
84.1 INTRODUCTION Methyl bromide is a broad spectrum pesticide primarily used for soil fumigation, commodity/quarantine treatment, and structural fumigation. Since human exposure is more likely to occur by the inhalation route, the majority of the toxicologic evaluations for methyl bromide are inhalation studies. Ingestion of fumigated commodities is a secondary route of human exposure. Chronic dietary studies in rats and dogs show no concern for long term oral ingestion of methyl bromide. This chapter briefly describes many of the published studies and elaborates the results from a number of contemporaneous methyl bromide toxicity studies by the oral and inhalation routes which were conducted to support the pesticide registration and other regulatory needs of the U.S. Environmental Protection Agency as well as State and international regulatory bodies. A primary focus in this chapter is methyl bromide induced neurotoxicity. From a risk characterization standpoint, clinical observations of neurotoxicity are considered as the primary endpoint of concern from inhalation exposure. Review of the overall toxicity of methyl bromide shows that methyl bromide induced toxicity is a function of both the concentration and the duration of exposure. This is an important consideration for human exposure assessments.
84.2 CHEMICAL PROPERTIES AND PESTICIDAL USES OF METHYL BROMIDE Methyl bromide (CH3Br, bromomethane, CAS no. 74-83-9) is a colorless, odorless gas at normal temperature and pressure. Methyl bromide is produced by the interaction of methanol (CH30H) and hydrogen bromide (HBr). It is made commercially but also is produced natually by marine algae and other plants and as a by-product of the combustion of plant materials (i.e., forest fires). Under increased pressure or below 3°C, methyl bromide is a clear to straw colored liquid and it is usually shipped as a liquified, compressed gas. Methyl bromide has a boiling point of 38.5° Fahrenheit and is nonflammable in air. Methyl bromide formulations contain chloropicrin, an irritant and lacrimator, as a warning agent. Handbook of Pesticide Toxicology Volume 2. Agents
Methyl bromide is a broad spectrum pesticide primarily used for soil fumigation, commodity/quarantine treatment and structural fumigation. It is also used as an intermediate in the manufacture of other chemicals. Methyl bromide has been used as a fumigant for more than 50 years and is strictly controlled by the U.S. Environmental Protection Agency (EPA) under the Federal Insecticide, Fungicide, and Rodenticide Act. Its application and use are also controlled by various state regulatory authorities. For soil fumigation, methyl bromide is injected directly into the soil, which is then covered with plastic sheeting. The sheeting is sealed, kept in place for several days, and then removed. Soil fumigation with methyl bromide enhances the quality of the crops and increases yield by eliminating fungal diseases, nematodes, weed seeds, and other soil borne pests. The primary crops grown in methyl bromide treated soil are peppers, strawberries, tomatoes, and grapes. Methyl bromide is widely used for fumigating postharvest commodities, such as wheat and cereals, spices, nuts, and dried and fresh fruits to eradicate pest infestations. Fumigation typically occurs where the commodities are stored, such as in ship holds, grain elevators, warehouses, special fumigation chambers, and on shipping piers/docks. Commodity fumigation typically involves the use of specially designed and permanently installed chambers into which the methyl bromide is released. After treatment, mechanical ventilation is used to continuously aerate the commodity until the concentration of methyl bromide in the vented air is at established safety levels. Another type of commodity fumigation involves the sealing of the commodity under a tarpaulin followed by injection of the methyl bromide. Aeration and ventilation occurs after the tarpualin is removed. In structural fumigation, all openings in the structure are sealed. All types of commercial and residential structures may be fumigated with methyl bromide to control or eradicate pests, such as termites. The structure is covered by a "tent" or tarpaulin and the methyl bromide gas is released inside the structure. After a specified period, the tarpaulin is removed and the structure is aerated until the concentration of methyl bromide inside the structure reaches safe levels.
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Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
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CHAPTER 84 Methyl Bromide
Methyl bromide may only be applied and used by professional, certified applicators. All persons working with methyl bromide are required to be knowledgeable about its hazards and trained in the use of required respiratory protection equipment, detector devices, and emergency procedures. Applicators and other persons in the fumigation area must wear appropriate personal protective equipment (PPE) as required by the label and V.S. EPA regulations. Such PPE typically includes full eye/face shields, safety shoes, and respirators. If the concentration of methyl bromide in the work area exceeds establishes safety levels, all persons in the fumigated area must wear approved, self-contained breathing apparatus or evacuate the area. The placarding and posting of warning notices at all entrances to an area undergoing fumigation is required. No one is permitted in a structure or area undergoing fumigation, unless they are involved in the fumigation and are wearing appropriate PPE. Re-entry into fumigated areas or structures is prohibited until the air concentration of methyl bromide is shown to be at safe levels. Individuals living in close proximity to fumigated fields, greenhouses, or structures are unlikely to be exposed to unsafe levels of methyl bromide because the application restrictions and the rapid dissipation of methyl bromide in the atmosphere. Additional regulatory controls further limit this possibility. For example, in California, maximum air concentration levels have been established for state-mandated buffer zones surrounding fumigated areas.
84.3 TOXICOLOGY OF METHYL BROMIDE The toxicology of methyl bromide has been extensively reviewed (ATSDR, 1991; WHO, 1995). This chapter briefly describes many of the published studies and elaborates the results from a number of contemporaneous toxicity studies with methyl bromide which were conducted to support the EPA reregistration and to provide specific data to meet various state registration/regulatory requirements. A primary focus in this chapter is methyl bromide induced neurotoxicity. From a risk characterization standpoint, clinical observations of neurotoxicity are considered as the primary endpoint of concern from inhalation exposure. Reviewing of the overall toxicity of methy I bromide shows that methyl bromide induced toxicity is a function of both the concentration and the duration of exposure. This is an important consideration in evaluating potential human risks.
84.3.1 ACUTE TOXICITY 84.3.1.1 Oral The acute LD50 for methyl bromide in rats was reported as 214 mg/kg (Danse et aI., 1984). Prior to conducting longer term toxicity studies, an acute oral toxicity study was conducted which compared liquid methyl bromide to a microencapsulated form. Similar oral toxicity was noted for both forms of methyl bromide; the oral LD50 values were 104 mg/kg for
liquid methyl bromide and 133 mk/kg for microencapsulated methyl bromide (Kiplinger, 1994). In beagle dogs, 500 mg/kg produced severe signs of toxicity and vomiting followed by death within 24 hours of dosing. A 50 mg/kg dose elicited signs of toxicity and vomiting of reddish material but no deaths. At low doses of 5 and 3 mg/kg, dogs vomited shortly after dosing. An oral LD50 study could not be conducted since dogs vomited the dose (Naas, 1990).
84.3.1.2 Inhalation Overt toxicity (i.e., death) from acute inhalation exposures to methyl bromide has been extensively evaluated in rodents (Alexeeff and Kilgore, 1985; Irish et aI., 1940; Japanese Ministry of Labour, 1992; Zwart, 1988). The majority of the acute inhalation studie in mice and rats demonstrate that methyl bromide exposure related effects and mortality are a function of both the concentration and the duration of exposure. Inhalation LC50 values for methyl bromide in mice have been reported as 1700 ppm for a 30 minute exposure (Bakhishev, 1973), 1200 ppm for a 60 minute exposure (Alexeeff and Kilgore, 1985),397 ppm for a 120 minute exposure (Balander and Polyak, 1962), and 405 ppm for a 240 minute exposure (Yamano, 1991). Similarly, inhalation LC50 values for rats were 2833 ppm for 30 minute exposure (Bakhishev, 1973), 1880ppm for a 60 minute exposure (Zwart, 1988; Zwart et aI., 1992), 781 ppm for a 240 minute exposure (Kato et aI., 1986), and 302 ppm for an eight hour exposure (Honma et aI., 1985). A number of acute inhalation study provide results which clearly demonstrate the concentration and the duration of exposure relationship for methyl bromide. In an acute inhalation study, groups of F344 rats were exposed to methyl bromide at concentrations of 150, 225, 338, 506, 760, or 1140 ppm for four hours (Japanese Ministry of Labour, 1992). This single exposure resulted in decreased locomotor activity, ataxia, nasal discharge, lacrimation, diarrhea, irregular breathing, and bradypnea in rats exposed to 338 ppm and greater. No clinical signs of toxicity were evident in animals exposed to 225 ppm methyl bromide or less. Histologic evaluations revealed metaplasia of the olfactory epithelium for rats exposed to 225, 338, and 506 ppm methyl bromide. Honma et al. (1985) conducted a series of acute inhalation toxicity studies to evaluate methyl bromide-induced effects on locomotor activity, body temperature, body weight gain, and enhancement of thiopental-induced sleep. Rats were exposed to methyl bromide concentrations of 63, 125, 188, or 250 ppm for eight hours. At 63 ppm and greater, enhanced thiopental sleep potentiation, measured by time to loss of righting reflex upon thiopental injection, was noted. Body weight gain and body temperature were decreased in rats exposed to methyl bromide concentrations of 125 ppm and greater. Neurotoxicity, indicated by reduced locomotor activity, was seen at concentrations of 188 and 250 ppm methyl bromide. These effects were reversible within 24 hours of exposure. In a study that evaluated histologic changes from acute inhalation of methyl bromide (Hurtt et aI., 1987), Fischer 344
84.3 Toxicology of Methyl Bromide
rats were exposed to methyl bromide concentrations of 0, 90, 175, 250, or 325 ppm on a six hours/day, five consecutive day regimen. Diarrhea was noted by the end of the second day of exposure for 250 and 325 ppm animals. By the end of the third exposure, animals from these groups showed ataxia. Two of the 325 ppm rats exhibited tremors and/or convulsions during the fourth exposure. Subsequently, three animals from this group succumbed after the fourth exposure. Clinical signs of neurotoxicity were not observed in the 250 and 325 ppm animals after a single (or second) exposure to methyl bromide. Irish et at. (1940) exposed rats and rabbits to methyl bromide concentration ranging from 108 to 12,850 ppm. The study results showed clear concentration and exposure duration dependence. Rabbits tolerated exposure to 220 ppm methyl bromide for 20 hours but exposure at this concentration for 32 hours resulted in 100% mortality. At 2570 ppm, rabbits survived a 1 hour exposure to while 100% mortality was observed after 2.2 hours exposure. Groups of mice were exposed for four hours via whole body exposure to methyl bromide atmospheric concentrations of 100, 150,225,338,506, or 760 ppm (Japanese Ministry of Labour, 1992). Concentration dependent clinical signs of toxicity consisting of decreased locomotor activity, tremors, convulsions, diarrhea, dyspnea, and bradypnea were seen at concentrations of 506 and 760 ppm methyl bromide. The acute inhalation NOAEL for clinical signs of neurotoxicity in mice exposed to methyl bromide was 338 ppm. Alexeeff and Kilgore (1985) conducted a series of onehour inhalation exposure studies in which mice were exposed to methyl bromide concentrations ranging from 225 to 1530 ppm. Based on all signs of neurotoxicity, the NOEL was 560ppm. In a series of inhalation studies (Newton, 1994a), beagle dogs received one to four days exposure to methyl bromide. The purpose of this study was to determine tolerable inhalation exposure levels to be used in a four-week inhalation study. In the initial phase of the study, three males and three females were exposed for six to seven hours to methyl bromide concentrations of 233 (one male), 314 (one male, one female), 345/350 (one male, one female), or 394 ppm (one female). Signs oftoxicity were observed at all concentrations, therefore, the one-day NOAEL was <233 ppm. In the second phase of the study, dogs were exposed to either 55 ppm (one male, one female), 156 (one male, one female), 268 ppm (one male, two females),or 283 ppm (two males, one female) for seven hours/day for up to four days. The 268 ppm and 283 ppm dogs were exposed to methyl bromide for two days and developed clinical signs of toxicity. Therefore, the two-day NOAEL for beagle dogs exposed to methyl bromide was <268 ppm. The 55 ppm and 156 ppm dogs were exposed to methyl bromide seven hours/day for four consecutive days. No effects were seen in either the 55 or 156 ppm dogs during days 1 and 2 of exposure. However, the 156 ppm animals showed decreased activity during exposure on days 3 and 4 and irregular gait during the postexposure period on day 4.
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84.3.2 SUB CHRONIC TOXICITY 84.3.2.1 Oral In a subchronic toxicity study (Danse et al., 1984), Wistar rats were dosed with methyl bromide via gavage at doses of 0, 0.4, 2, 10, or 50 mg/kg/day for 90 days. At 50 mg/kg/day, marked, diffuse hyperplasia of the epithelium of the forestomach was seen in all animals. Squamous cell carcinoma of the forestomach was diagnosed in 13 of 20 animals receiving methyl bromide at 50 mg/kg/day. Upon subsequent evaluation of the histology slides, it was concluded that the forestomach lesions represented inflammation and hyperplasia rather than malignant lesions. Inflammatory lesions of the forestomach were also seen in animals treated with 2 and 10 mg/kg/day. 84.3.2.2 Inhalation Table 84.1 summarizes the results from a number of subchronic inhalation studies in rats and mice. For comparison purposes, the study results have been presented to show the overall study NOEL, the NOEL for neurotoxicity, and the LOEL for neurotoxicity. Beagle dogs were exposed to either 55 ppm (one male, one female), 156 (one male, one female), 268 ppm (one male, two females), or 283 ppm (two males, one female) for seven hours/day for up to four days (Newton, 1994a). Clinical signs of toxicity were seen in the 268 and 283 ppm dogs after two exposures. The 156 ppm animals showed decreased activity during exposure on days 3 and 4 and irregular gait during the post exposure period on day 4. The 55 ppm concentration was the four-day NOEL. In a four-week inhalation study (Newton, 1994b), beagle dogs (four/sex/group) were exposed for five days/week, seven hours/day to methyl bromide at concentrations of 0, 5, 10, 25, 50, or 100 ppm. No clinical evidence of neurotoxicity was seen in any group throughout the four weeks of exposure. After four weeks, four ofthe controls (two females and two males) and all dogs in the 5 ppm group continued the test for an additional two weeks and the exposure concentration for the 10 ppm group was increased to 150 ppm. Dogs exposed to 150 ppm methyl bromide showed severe body weight loss over the first few days of exposure. After five or six exposures, evaluation of the 150 ppm animals by a veterinary neurologist revealed ataxia, a base-wide stance, intention tremor, nystagmus, marked depression, and inability (unwillingness) to stand and perform postural responses. Due to the severity of these effects, the dogs were sacrificed. Neurologic evaluation revealed no treatment related neurologic effects for dogs exposed at the lower methyl bromide concentrations. Microscopic findings were limited to the 150 ppm group in which significant neurologic effects were seen and consisted of vacuoles in the granular layer of the cerebellum. 84.3.3 GENETIC TOXICITY Methy I bromide has been tested in numerous in vitro and in vivo genetic toxicity studies with variable results.
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CHAPTER 84 Methyl Bromide
Table 84.1 Summary of Subchronic Inhalation Toxicity Studies with Methyl Bromide Species
Exposure
Neurotoxicity NOEL
Neurotoxicity
Study
(Strain)
(hours/days/weeks)
Overall NOEL
(effect)
LOEL
Subchronic neurotoxicity
Rat
6/5113
30ppm
30ppm
70ppm
6/5/3; 61711
18 ppm
18 ppm
51 ppm
4/5/6
<150ppm
200ppm
300ppm
(Norris et aI., 1993) Subchronic toxicity (NTP, 1992) Subchronic toxicity (Kato et aI., 1986) Subchronic toxicity (Haber et aI., 1985)
(SD) Rat (SPF Wistar) Rat (SD) Rat
(dec. body weight) 6/5113
30ppm
(F344/N)
60ppm
120ppm
(dec. body weight)
(NTP, 1992) Subchronic toxicity (Japanese Ministry, 1992) Subchronic toxicity (Wilmer et al., 1983) Subchronic toxicity (NTP, 1992) Subchronic toxicity (Japanese Ministry, 1992)
Rat
6/5/13
7.5 ppm
Rat
6/5/13
6.4ppm
(Wistar) Mouse
293 ppm
42ppm
no neurotoxicity
(liver pathology) 6/5/13
20ppm
(B6C3Fl) Mouse
117 ppm (clinical pathology)
(F344/DuCrj)
80ppm
120 ppm
(hematology) 6/5/13
(Crj:BDFl)
Methyl bromide has been reported to induce mutagenic effects in bacterial tests with Salmonella typhimurium (TAIOO and TAI535), E. coli, and Klebsiella pneumoniae (DjalaliBehzad et aI., 1981; Kramers et aI., 1985a; Moriya et aI., 1983; Simmon et aI., 1977). No evidence of mutagenicity was seen when methyl bromide was tested in a modified Ames test using an in situ impingement test system but a significant response was seen with the SOS repair test (Ong et aI., 1987). A sex-related recessive lethal assay was conducted with male strain Oregon K Drosophila melanogaster (McGregor, 1981). The Drosophila were exposed to methyl bromide concentrations of 20 or 70 ppm for 5 hours and then allowed to mate on days I, 3, or 8 following exposure. The FI progeny were mated brother to sister, I to 4 days after emergence. The resulting F2 generation was then examined for the absence of wild-type males. No compound-related increases in the frequency of lethal mutations in the F2 generation were noted. In a second sex-related recessive lethal assay, Drosophilia melanogaster of the Berlin K strain were exposed to methyl bromide concentrations ranging from 18 to 192 ppm for varying exposure intervals (Kramers et aI., 1985a, b). As was noted for the acute inhalation studies in mammalian species, mutagenic responses were related to both exposure concentration and duration. No increase in mutation frequency was seen in Drosophila exposed to 192 ppm for six hours at 192 ppm. Exposure at 155 ppm resulted in all flies dying during the fourth day of exposure. At lower concentrations, prolonged exposure resulted in mutagenic responses. Exposure at 125 ppm for five days (six hours per day) and at 50 ppm for 15 days (six hours per day) were considered mutagenic.
30ppm
60 ppm (body weight,
no neurotoxicity
hematology, urinalysis)
Exposure of L5178Y mouse lymphoma cells to methyl bromide concentrations ranging from 7.7 to 7710 ppm resulted in dose-related increases in 6-thioguanine- and bromodeoxyuridine-resistant mutants (Kramers et aI., 1985a). Sister chromatid exchanges (SeEs) and chromosomal aberrations in human lymphocytes exposed to methyl bromide were evaluated. Exposure of human Iymphocyte cultures to an atmosphere of 4.3% methyl bromide for 100 seconds increased the frequency of SeEs from 10.0 to 16.8 per cell (Tucker et aI., 1986). When human lymphocytes were treated with methyl bromide (0-24 ).l.g/ml) for 30 minutes, dose-related increases in SeEs and chromosomal aberration were found. Metabolic activation (S9) significantly induced chromosomal aberrations (Garry et aI., 1990). Drosophila melanogaster (third instar larvae trans-dihybrid for two recessive wing hair mutations) were exposed to methyl bromide vapor concentrations ranging up to 5140 ppm for one hour in a mitotic recombination assay in somatic cells (somatic wing spot assay). Wings of surviving adults were evaluated for the presence of cellular clones with malformed wing hairs. Methyl bromide induced mitotic recombination as exhibited by the observation of small and large single as well as twin spots (Katz, 1985, 1987). A rodent micronucleus (MN) study was conducted in BDFl mice and F344 rats (lO/sex/group) exposed via vapor inhalation to methyl bromide concentrations of 0, 154,200,260,338, or 440 ppm for six hours/day, five days/week for two weeks (Araki et.al, 1995). Bone marrow of rats and peripheral blood of mice were evaluated for MN induction. In mice, significantly increased incidences of micronuclei in bone marrow polychromatic erythrocytes (peE) were observed in males at 154 and
84.3 Toxicology of Methyl Bromide 200 ppm and in females at 154 ppm; smaller increases in MN frequency were observed in normochromatic erythrocytes. Peripheral blood showed significant increases in MN at 200 ppm in males and 154 ppm in females. Due to excessive mortality, mice exposed to methyl bromide concentrations of 260 ppm and greater were not assayed. In rats, a statistically significant increase ofMN in PCE was seen for males exposed at 338 ppm. A nonstatistically significant increase of MN in PCEs was seen in female rats exposed at 260 and 338 ppm. Rats exposed at 400 ppm were not assayed due to excessive mortality. Methyl bromide was selected by the National Institutes of Occupational Safety and Health for evaluation in a Tier II Mutagenic Screening (McGregor, 1981). The testing program included: (1) unscheduled DNA synthesis (UDS) assay in human diploid fibroblasts, (2) sex-linked recessive lethal test in Drosophila melanogaster, (3) cytogenetic test in bone marrow cells of male and female rats, (4) sperm abnormality test in male mice, and (5) dominant lethal test in male rats. Summary results from this screening battery show that: • Human diploid fibroblasts exposed for three hours to methyl bromide concentrations of up to 70% in air over a minimal volume of culture medium in a UDS resulted in no increase in UDS. • Drosophila melanogaster exposed to methyl bromide concentrations of 20 or 70 ppm for five hours did not have an increased frequency of sex linked recessive mutations. • Cytogenetic analysis of bone marrow cells derived from male and female Sprague Dawley rats that were exposed for one or five days (seven hours/day) to methyl bromide concentrations of 20 or 70 ppm showed no treatment related increases in the frequency of chromosomal aberrations in any of the methyl bromide exposed groups. • Evaluation and characterization of sperm from male B6C3Fl mice exposed to 20 or 70 ppm methyl bromide for seven hours/day on five consecutive days then sacrificed five weeks later showed no significant increase in the frequency of abnormal sperm. • In the dominant lethal study, male Sprague Dawley rats were exposed to methyl bromide concentrations of (air), 20, or 70 ppm for seven hours per day for five consecutive days then allowed to breed with two virgin females weekly for 10 weeks. The females were sacrificed on Day 14 after presumed mating. Examination of the ovaries and the uterine contents showed no evidence of genotoxicity.
°
In a separate study, methyl bromide was evaluated for its ability to induce single strand breaks in rat testicular DNA using alkaline elution techniques (Bentley, 1994). In this study, groups of 10 male Fischer 344 rats were exposed to methyl bromide vapor concentrations of 0, 75, 150, or 250 ppm for six hours per day over five consecutive days. The negative control group was exposed to room air only. A positive control group received a single intraperitoneal injection of 50 mg/kg methyl methanesulfonate in phosphate buffered saline. Five animals
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from each group were sacrificed at 1 and 24 hours postexposure. Significant toxicity was seen in the 250 ppm rats. Two males from this group died and a third male was sacrificed in extremis with the I-hour post treatment animals. Surviving rats showed decreased body weight and clinical signs of toxicity characterized by ataxia, spasms, diarrhea, lethargy, and prostration. At 150 ppm, the male rats showed slight body weight loss, nasal and/or ocular discharge and wet and/or stained perineum during the five-day exposure. A statistically significant increase in the mean elution rate of testicular cell DNA was observed at both sacrifice times only in rats exposed to the highly toxic 250 ppm concentration. 84.3.4 DEVELOPMENTAL AND REPRODUCTIVE TOXICITY
In a developmental toxicity study (Sikov et aI., 1981), pregnant female Wistar rats were exposed to methyl bromide concentrations of 0, 20, or 70 ppm for seven hours/day on gestation days 1 to 19. In addition, some groups were exposed pregestationally to 20 or 70 ppm on a for three weeks (five days per week) immediately prior to mating. The distribution of dose groups (pregestational exposure concentration/gestational exposure concentration) was 0/0, 0120, 0/70, 20/0, 20120, 70/0, and 70/70 ppm. Cesarean sacrifice was performed on gestation day 19. No clinical evidence of maternal toxicity, fetotoxicity or developmental toxicity was observed in any exposure scenario. Artificially inseminated New Zealand White rabbits (24 per group) were exposed daily on gestation days 1 through 24 to methyl bromide concentrations of and 20 ppm; a group of inseminated rabbits exposed to 70 ppm methyl bromide were terminated due to excessive mortality and neurotoxicity characterized by convulsions and paresis in the hindlimbs seen after one week of treatment (Sikov et aI., 1981). Control and 20 ppm exposed rabbits were sacrificed on gestation day 30. No fetoxicity nor developmental toxicity was noted for the 20 ppm group. In probe studies (Breslin et aI., 1990a), pregnant rabbits were exposed to methyl bromide concentrations of 0, 10, 30, or 50 ppm in one study and concentrations of 0, 50, 70, or 140 ppm in a second study. Exposure was for six hours/day on gestation days 7 to 19. Evidence of toxicity was observed only in the 140 ppm group does. All does exposed to methyl bromide at this concentration showed lethargy and decreased food consumption after eight exposures. With continued exposure signs of neurotoxicity were apparent and resulted in sacrifice of the does on gestation day 17. No apparent embryotoxicity was observed at any exposure level. The subsequent developmental toxicity study in rabbits was conducted in two phases (Breslin et aI., 1990a, 1990b). In the initial phase, pregnant New Zealand White rabbits were exposed for six hours/day to methyl bromide concentrations of 0, 20, 40, and 80 ppm on days 7 through 19 of gestation. In the second phase, pregnant does were exposed to or 80 ppm only. Cesarean delivery was performed on day 28 of gestation. In the first phase, maternal toxicity, evidenced by decreased bodyweight gain and clinical
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CHAPTER 84
Methyl Bromide
signs of neurotoxicity, was observed in three of the does from the 80 ppm group. The clinical signs consisted of right-sided head tilt, ataxia, slight lateral recumbency, and lethargy. In the second study, a significant decrease in body weight during gestation was the only evidence of maternal toxicity in the 80 ppm group. Developmental effects were limited to the maternally toxic 80 ppm group only. In phase 1, fetal findings consisted of low incidences of omphalocele, hemorrhaging with or without hydrops (edema), retroesophogeal right subclavian artery, gall bladder agenesis, and fused sternebra. Fetal effects in phase 2 were limited to decreased fetal weight, hemorrhaging with or without hydrops and gall bladder agenesis. Male and female Sprague Dawley rats were exposed to methyl bromide by whole body inhalation exposure, six hours/ day, five days/week at concentrations of 0, 3, 30, or 90 ppm in a two-generation reproduction study (American Biogenics Corporation, 1986). Two litters were produced for each generation. No deaths nor noteworthy antemortem clinical finding were observed over the course of the study. The 90 ppm FO males had significantly decreased body weights at five of the 10 premating intervals and at final sacrifice. No other decreases in body weights were observed among the FO generation or during the Fl generation prior to the gestational period for the F2a litter. A slight depression of body weight was noted during the gestation and lactation periods for the 90 ppm Fl dams. Reproductive performance was not altered by methyl bromide exposure and there were no significant differences in pup survival. No methyl bromide related anomalies were noted for the progeny. Gross pathologic examination revealed no treatment related lesions in either the parental animals or their progeny. Mean brain weight for the 90 ppm males (PO and Fl) and females (Fl) were decreased Increased liver to body weight ratio for the 90 ppm PO males and females and increased heart to brain weight ratios for the 90 ppm Fl females were noted. No other significant differences were seen in the parent organ weight data. No significant differences in the Fl b progeny body and organ weights were noted but statistically significant decreases in final body weights were observed for the 90 ppm F2b males and females and the 30 ppm F2b females. F2b progeny organ weights were significantly reduced for the 90 ppm female brain, heart, kidney, and liver weights, the 30 ppm female liver weight, and the 30 and 90 ppm female liver to brain weight ratio. The 30 and 90 ppm F2b female brain to body weight ratio was increased. There were no other significant differences noted for progeny. Microscopic examination of the reproductive organs and abnormal tissues revealed no treatment related lesions. 84.3.5 CHRONIC TOXICITY AND ONCOGENICITY-INHALATION Wistar rats were exposed (whole body) to methyl bromide at atmospheric concentrations of 0, 3, 30, or 90 ppm on a six hour/day, five days/week basis for 29 months (Reuzel et aI., 1987, 1991). At the 90 ppm concentration, decreased survival was noted for both the males and females from the end of the
second year through termination at 29 months. Also, at this concentration, body weights, especially for females, were lower than the control group from week 4 and throughout the remainder of the study, and decreased absolute brain weight was noted for females. No differences in hematology, clinical pathology, or urinalysis were seen at either the 3 month or one year intervals. No treatment related evidence of neoplasia was observed in the study. Treatment related nonneoplastic pathology consisted of an increased incidence of thrombi in the heart, and myocardial degeneration for both sexes from the 90 ppm group. Irritation of the nasal cavity characterized by hyperplasia of the olfactory epithelium was seen in a time-related fashion for all methyl bromide treated groups. Dose-dependent increases in the incidences of degenerative and hyperplastic changes of the nasal olfactory epithelium were observed. The lesions were characterized as very slight, slight, or moderate. A statistically significant increase was found between controls and the lowdose group (3 ppm) at the end of the exposure period (29 months). However, the frequency of this lesion also increased in the controls (age dependence) from 12 through 24 months to 29 months of age. In addition, all but one of the lesions in the 3 ppm exposure level group were described as slight or very slight. Moreover, one moderate lesion of the nasal mucosa was also observed in a control animal at the 24-month sacrifice interval. The NOEL for this lesion was >90 ppm after 12 months of exposure, 3 ppm after 24 months of exposure, and <3 ppm after 29 months of exposure. The Gotoh et al. (1994) study directly correlate with the Reuzel et al. (1987, 1991) study at the 24-month interval. Gotoh et al. exposed F344 rats to methyl bromide concentrations of 0,4,20, and 100 ppm on a six hour/day, five day/week basis for 104 weeks. After 24 months of exposure, increased incidences of necrosis and respiratory metaplasia of the olfactory epithelium were seen for male rats exposed to 100 ppm methyl bromide; these findings were marginally increased for the female rats at 100 ppm. Metaplasia was noted for 22% of the control males and 6% of the control females at the 24-month terminal interval. These results show that metaplasia produced in the rat olfactory epithelium was a threshold response upon chronic inhalation exposure to methyl bromide and that a high control incidence of metaplasia is noted in aged rats. B6C3F1 mice were exposed via inhalation (whole body) to concentrations of 0, 10, 33, or 100 ppm methyl bromide on a six hours/day, five days/week basis for two years (NTP, 1992). The 100 ppm exposure concentration clearly exceeded an acceptable maximum tolerated dose for carcinogenicity testing. This exposure concentration was terminated after 20 weeks due to debilitating neurotoxicity and mortalities; these animals were exposed to untreated air for the remainder of the twoyear study period. Interim sacrifice of 10 mice per sex/treatment level was performed after 6 and 15 months of exposure. Neurobehavioral testing was performed on selected animals every 3 months. Clinical signs of neurotoxicity, consisting of tremors, paralysis, gait disturbances, and abnormal posture, were noted for 100 ppm males (78%) and females (43%). Similar findings were seen for a few (2 to 3%) of the 33 ppm exposed
84.3 Toxicology of Methyl Bromide animals. After 3 months of exposure, neurobehavioral changes were noted for the 100 ppm males and females. Neurobehavioral testing also revealed changes in the 10 and 33 ppm groups after 6 months of exposure. Decreased body weights were observed in females dosed at 33 ppm and in both sexes dosed at 100 ppm. Exposure related histologic changes were generally limited to the 100 ppm animals and consisted of findings in the brain (degeneration of the cerebrum and cerebellum), heart (degeneration and cardiomyopathy), sternal dysplasia, and either necrosis or metaplasia of the olfactory epithelium. 84.3.6 CHRONIC TOXICITY AND ONCOGENICITY-DIETARY Methyl bromide was evaluated for chronic toxicity and oncogenicity in a 24-month dietary toxicity study (Mertens, 1997) in Sprague Dawley rats. Because of the volatile nature of methyl bromide and the feeding characteristics of rats, it was not possible to conduct the study using feed fumigated with methyl bromide. For purposes of this study, methyl bromide was microencapsulated and mixed into the rodent diet. Methyl bromide dietary concentrations were 0.5, 2.5, 50, and 250 ppm (0.02, 0.11, 2.20, and 11.10 mg/kg/day for males and 0.3, 0.15, 2.92, and 15.12 mg/kg/day for females, respectively). Basal diet and placebo (microcapsules without methyl bromide) control groups were treated on a comparable regimen. No methyl bromide related effects were seen on survival, clinical condition, hematology, serum chemistry, urinalysis, organ weights, ophthalmologic assessments, or macroscopic and microscopic pathology evaluations. Food consumption, mean body weights, and mean body weight gains were reduced in the 250 ppm males and females during the rapid growth phase for the animals during the first 12 to 18 months of the study. During the first 18 months of the study, mean body weight gain for males were 9% to 21 % lower than the male control groups while mean body weight gain for females was 7% to 22% lower than the female control groups. Typical of chronic toxicity studies, food consumption and body weight gains during the second year of the study were comparable to controls as the mature animals reached adult body weight plateau. No evidence of oncogenicity was seen in this study. In a 12-month dietary safety study (Newton, 1995), beagle dogs were exposed to methyl bromide fumigated feed at dietary concentrations of 0, 0.5, 1.5, or 5 ppm (0, 0.06, 0.13, and 0.27 mg/kg/day for males and 0, 0.07,0.12, and 0.26 mg/kg/day for females, respectively). Prestudy trials were conducted to determine the methyl bromide fumigation concentrations and postfumigation intervals required to achieve the desired concentrations over a I-hour feeding period. No toxicologically significant methyl bromide effects were seen in clinical observations, body weight, body weight gain, food consumption, clinical pathology, urinalysis, ophthalmology, absolute or relative organ weights, and macroscopic or microscopic pathology. Based on the results of this study, the NOEL for methyl bromide when administered via fumigated feed to beagle dogs
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was greater than 5 ppm (>0.27 mg/kg/day for males and >0.26 mg/kg/day for females). 84.3.7 NEUROTOXICITY In an acute neurotoxicity study (Driscoll and Hurley, 1993), male and female CD (Sprague Dawley) rats were exposed via inhalation for six hours to methyl bromide at concentrations of 0, 30, 100, or 350 ppm. Animals were assessed for clinical signs and changes in body weights. Neurobehavioral evaluations were performed within three hours of exposure and at 7 and 14 days postexposure. These evaluations included the functional observation battery and motor activity assessments. After 15 days, animals were euthanized, necropsied, and examined for gross pathologic changes, and brains were weighed. In addition, microscopic evaluations were performed on central and peripheral nervous tissue. All animals survived to study termination. No methyl bromide induced effects were noted for body or brain weights. Neurobehavioral effects were observed only in the 350 ppm exposed group and were limited to the three-hour postexposure assessment. Effects noted in male and female rats consisted of decreased arousal, increased incidences of drooping or half-shut eyelids, piloerection, decreased rearing, depressed body temperature, and markedly decreased motor activity. The 350 ppm males had a decreased tail pinch response while females from this group showed increased urination and abnormal air righting response. No treatment related histological findings were seen in nervous system or nasal tissues. In a subchronic inhalation neurotoxicity study (Norris et aI., 1993) CD (Sprague Dawley) rats were exposed to methyl bromide concentrations of 0, 30, 70, or 140 ppm. Exposure was six hours/day, five days/week for 13 weeks. At the 140 ppm concentration, two male rats died during the first month. Clinical signs observed for these rats included convulsions, tremors, hyperactivity, rapid respiration, and salivation. Mean body weights were significantly lower than the controls. Neurologic evaluations for males revealed increased hind limb splay (weeks 4, 8, 13), abnormal air righting reflex (week 13), and decreased forelimb grip strength (week 13). Female rats demonstrated lower arousal scores (weeks 8, 13), decreased rearing (weeks 4,8, 13) and significantly decreased motor activity (week 13). Mean absolute brain weights were significantly lower for both sexes; no differences were noted for the relative brain weights, indicating that lower absolute brain weight was a reflection of the generally lower body weights for the treated animals. Gross lesions were limited to moderate to severe brain hemorrhage in the two 140 ppm male animals which died. Microscopic lesions in the brain were found in these two males and in one 140 ppm male that survived the 13-week exposure. Microscopic lesions in the brain were seen in these three males and consisted of neuronal necrosis in the hippocampus, necrosis and malacia in the cerebral cortex and basal ganglia, and malacia and/or necrosis in the thalamus and midbrain. The lesions were more severe for one of the males which was found dead
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CHAPTER 84 Methyl Bromide
and the male that survived the 13-week exposure; both of these animals were noted with convulsions during the study, suggesting that some of the microscopic brain findings may have been secondary effects of brain swelling related to the convulsions. One additional 140 ppm male had slight neuronal edema in the hippocampus. Other lesions in the 140 ppm group consisted of minimal regenerative dysplasia of the olfactory epithelium of the nasal cavity in three males and three females and minimal peripheral nerve degeneration in two males and two females. In the 70 ppm group, lower mean body weights and weight gain were seen for females from week 9 onward of the study. Neurologic findings were limited to slightly decreased forelimb grip strength (week 13) in males and decreased motor activity (week 13) in females. Although the mean absolute brain weight for females was statistically significantly decreased (5% lower than the control group), no difference was seen for the relative brain weight and no microscopic pathology was seen. At 30 ppm, the mean absolute brain weight for females was statistically significantly lower than control (5% lower than the control group); however, no difference was seen in relative brain weight and no microscopic pathology in the brain was seen. Peripheral nerve degeneration was observed in one female rat. This finding was considered incidental since nerve degeneration was not seen in animals from the 70 ppm group. 84.3.8 SPECIFIC TARGET ORGAN EFFECTS
Methyl bromide is an unusual respiratory toxicant in the rat in that it is specifically toxic to the olfactory epithelium while other nasal epithelia are unaffected (Hurtt et al., 1988). Within the olfactory epithelium, methyl bromide only affects specific cell types. The major components of the rat olfactory epithelium are the basal cells, the long ducts of Bowman's glands, sensory cells, and the sustenacular or support cells. Studies using histochemical techniques clearly showed that methyl bromide specifically induced degeneration of the sensory and sustentacular cells while sparing the basal cells from which the sensory and sustenacular cells are regenerated. Hurtt et al. (1988) evaluated the time course for the regeneration of the olfactory epithelium following short term exposures to methyl bromide. Male rats were exposed to 200 ppm methyl bromide for six hours/day for one to five days. Air-exposed animals served as controls. In a companion study, animals were exposed to 0, 90, or 200 ppm for six hours and olfactory function assessed by the ability of food deprived animals to locate buried food pellets. Destruction of the olfactory epithelium was evident after a single six-hour exposure to 90 or 200 ppm. As discussed previously, severe effects were seen in the sustentacular and mature sensory cells while basal cells remained intact. Regeneration of the olfactory epithelium was seen as early as day 3 of exposure despite continued exposure at these high methyl bromide concentrations. The recovery of the olfactory epithelium was essentially complete by 10 weeks after exposure. The rapid recovery would be expected since the nonaffected basal cells regenerate sensory and sustentacular cells.
Olfactory function, as measured by food finding activity, was impaired in animals exposed to 200 ppm only. Recovery of this function was evident by four to six days postexposure, much earlier than the time course for histological recovery. In another study (Hastings et al., 1994), morphologic and biochemical (carnosine content of the olfactory bulb, a biomarker for integrity of the olfactory epithelium) evaluations were conducted to further explore methyl bromide exposure effects and recovery. Prior to treatment, rats were food-deprived and trained to find buried food pellets as a measure of olfactory function. The rats were exposed to a methyl bromide concentration of 200 ppm on a four hour/day, four days/week, two-week regimen. After a single exposure, extensive damage to the olfactory epithelium, reduced carnosine content, and impaired olfactory function were observed. Even though exposure continued, olfactory function began to improve after the first exposure. This recovery proceeded even though persistent thinning and disorganization of the olfactory epithelium and decreased carnosine levels in the olfactory bulb were present. Regeneration of the olfactory epithelium was complete approximately 30-40 days after the last exposure.
84.4 METABOLISM Inhalation was the primary route of exposure for the majority of the methyl bromide toxicology studies and is the most probable route of exposure for humans. As a result, the metabolism of methyl bromide has been almost exclusively evaluated by inhalation. 84.4.1 ABSORPTION
Medinsky et al. (1984, 1985) evaluated the uptake of methyl bromide upon six-hour exposure of rats to concentrations of 1.6,9.0, 170, or 310 ppm. Methyl bromide uptake was found to be linear over all exposure concentrations with the exception of the 310 ppm concentration. It was noted that suspected nasal irritation at 310 ppm may have reduced the total amount of methyl bromide inhaled due to decreased tidal and minute volumes (Medinsky et al., 1985). Total methyl bromide absorbed was 9 or 40 ).l.mol/kg body weight after exposure to 1.6 ppm (50 nmoVliter) or 9.0 ppm (300 nmoVliter), respectively. Uptake at the lower levels was approximately 48%. Uptake at 5700 (170 ppm) and 10,400 (310 ppm) nmoVliter was 37% and 27%, respectively (Medinsky et al., 1985). Andersen et al. (1980) found methyl bromide to exhibit rapid, first-order uptake kinetics. Saturation was not reached until concentrations causing animal death were achieved. Gargas and Andersen (1982) also demonstrated that methyl bromide uptake and metabolism followed first order over a broad range of exposure concentration (100-3000 ppm). Honma et al. (1985) also found methyl bromide to be rapidly absorbed and distributed. Methyl bromide concentrations in blood, liver, adipose, and brain reached maximum levels within one hour after the start of exposure and maintained almost the same levels
84.4 Metabolism
during the eight hour exposure. Equilibrium between methyl bromide air and tissue concentrations occurred rapidly. 84.4.2 DISTRIBUTION
Methyl bromide is rapidly and widely distributed in tissues immediately after exposure. Bond et al. (1985) investigated the tissue distribution of 14C methyl bromide. Radioactivity was widely distributed in tissues immediately following exposure with highest levels of 14C found in lung, adrenal, kidney, liver, and nasal turbinates. Low concentrations of l4C were also detected in other tissues. The liver was the only tissue examined immediately after exposure that contained a large percentage (about 17%) of the absorbed methyl bromide. Approximately 80% of the initial amount of tissue radiolabel was eliminated by 65 hours. Elimination halflives varied from 1.5 to 8 hours with the exception of the liver with an elimination half-life of 33 hours. Using non-radiolabelled methyl bromide, Honma et al. (1985) investigated the distribution of methyl bromide into liver, fat, brain, muscle, kidney, and blood upon inhalation exposure of rats to 250 ppm for 8 hours. Methyl bromide concentrations in all tissues listed reached maximum levels within 1 hour after the start of the exposure and was found at approximately the same tissue concentration through the remainder of exposure period. The highest tissue concentration was found in fat. Methyl bromide was rapidly eliminated from rat tissues following the cessation of exposure, with a half-life of about 30 minutes in the early postexposure period. At 48 hours postexposure, methyl bromide was not detected in any tissue examined. Medinsky et al. (1985) evaluated tissue distribution of methyl bromide 66 hours after exposure to range of vapor concentrations. Significant levels of 14C (approximately 20% of the absorbed methyl bromide dose) remained in tissues 66 hours after exposure. The highest level (approximately 20% of the 14C) was associated with the liver. Other tissues having appreciable concentrations of 14C included the lungs, nasal turbinates, and kidneys. Very low concentrations of 14C (less than 1 ).lmol of methyl bromide equivalents/g of tissue) were found in nervous tissues (spinal cord and brain). Tissue disposition after oral or intraperitoneal (IP) injection of 250 ).lmol/kg 14C-methyl bromide was also evaluated (Medinsky et aI., 1984). Approximately 14-17% of the radioactivity administered was found in the tissues and carcass 72 hours after oral and IP dosing, respectively. Analysis of individual tissues indicated liver was the major organ for retention of radioactivity after administration of methyl bromide. Other tissues containing significant amounts of radioactivity (> 10 nmol or 1% of the total dose) included kidney, testes, lung, heart, stomach, and spleen. As expected, the only tissue showing significantly higher radioactivity level after oral administration, as compared to IP injection, was the stomach.
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84.4.3 IDENTIFICATION AND QUANTITATION OF METABOLITES
In all methyl bromide metabolism studies, carbon dioxide was the major metabo1ite. Approximately 47% of the total methyl bromide dose was excreted as 14C02 in expired air (Bond et aI., 1985). About 1% of the total 14C-methyl bromide absorbed was exhaled as 14C-methyl bromide. Smaller quantities of the radiolabelled material were excreted in the urine and feces, with about 22% and 2% of the total absorbed 14C-methyl bromide excreted by these routes, respectively. All radioactivity in excreta was identified as methyl bromide degradates/metabolites. No evidence of parent chemical was found in any of the excreta samples (Bond et aI., 1985). Gargas and Andersen (1982) showed that bromine ion is released in the initial metabolism of a variety of brominated hydrocarbons. The study results showed that bromine ion is retained in extracellular fluid, is stable to further biotransformation, and is very slowly excreted. The first-order rate constant for the release of bromine from methyl bromide is 0.32/kg/hour. The elimination of bromine from rat tissue is slower than that of methyl bromide (Honma et aI., 1985). Peak concentrations of bromine in blood, kidneys, and liver occurred four to eight hours after methyl bromide exposure, and the half-life of bromine in these tissues was about five days. There was no correlation between the duration of bromine retention and observed signs of neurotoxicity (Gargas and Andersen, 1982). In a methyl bromide inhalation study in mice, Alexeeff and Kilgore (1985) followed mouse tissue bromine concentrations through 7 days postexposure. No bromide ion was detected in any tissue one week after exposure to concentrations up to 2.72 mg/l air. Over 95% of the bromide ion in exposed mice was eliminated within 2.5 days. The bromide ion levels were highest in liver and kidney and lowest in whole blood. Lung and brain bromide ion levels were intermediate. Honma et al. (1985) also investigated bromine and methanol concentrations in response to treatment with methyl bromide. Peak concentrations of bromine in blood, kidneys, and liver occurred four to eight hours after methyl bromide exposure, and the half-life of bromine in these tissues was about five days. Methanol production was not significant. 84.4.4 EXCRETION
Excretion of methyl bromide is primarily as exhaled C02. Very low levels of parent methyl bromide are found in expired air. Greater than 85% of the total amount of 14C that was exhaled as C02 and excreted in urine and feces was eliminated within 24 hours (Bond et aI., 1985). C02 excretion exhibited a biphasic elimination pattern with 85% of the 14C02 excreted with a half-life of about 4 hours and 15% excreted with a half-life of about 11 hours. The elimination half-life of 14C in urine was approximately 10 hours and in feces was approximatley 16 hours. Analysis of excreta samples in which there was sufficient radioactivity showed no evidence parent methyl bromide.
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CHAPTER 84 Methyl Bromide
84.5 HUMAN EXPOSURE The potential routes of human exposure to methyl bromide are oral (through consumption of fumigated food products), dermal (skin contact), or inhalation (exposure to methyl bromide gas). Extensive studies have shown that residues of methyl bromide found in crops grown on fumigated soils are virtually nondetectable. In addition, methyl bromide concentrations in commodities treated postharvest, decrease rapidly after required aeration, and are nondetectable after relatively short periods of time. Tolerance levels for the metabolite of methyl bromide (inorganic bromide) in treated foods have been established by the EPA. These tolerances further ensure that humans are not exposed to unsafe levels of methyl bromide's metabolite in foods. In summation, there is no significant likelihood of oral exposure to methyl bromide through consumption of treated food. Human exposure is more likely to occur by inhalation. People living in close proximity to fumigated fields, greenhouses, or structures are protected from the risk of significant inhalation exposure through special notice requirements, safety precautions, and the use of buffer zones. The potential for dermal or inhalation exposure to methyl bromide is highest for applicators and other personnel who are involved in manufacturing, filling, handling, or application of methyl bromide. Strictly applied safety measures in manufacturing and filling installations limit the potential risk of exposure to plant personnel. In addition, fumigators/applicators are protected from dermal and inhalation exposures through adherence to strict safety procedures and the use of protective equipment.
REFERENCES Alexeeff, G. v., and Kilgore, W. W. (1983). Methyl bromide. Residue Rev. 88, 101-153. Alexeeff, G. v., and Kilgore, W. W. (1985). Determination of acute toxic effects in mice foIIowing exposure to methyl bromide. 1. Toxicol. Environ. Health 15, 109-123. American Biogenics Corporation (1986). "Two-Generation Reproduction Study via Inhalation in Albino Rats Using Methyl Bromide." Unpublished report from Study 450-1525. Andersen, M., Gargas, M., Jones, R., and Jenkins, L. (1980). Determination of the kinetic constants for metabolism of inhalated toxicants in vivo using gas uptake measurements. Toxicol. Appl. Pharmacol. 54, 100-116. Anger, W. K., Setzer, J. V. Russo, J. M., Brightwell, W. S., Wait, R. G., and Johnson, B. L. (1981). Neurobehavioral evaluation of soil and structural fumigators using methyl bromide and sulfuryl fluoride. NeuroToxicology 7, 137-156. Araki, A., Kato, E Matsushima, T., Ikawa, N., and Nozaki, K. (1995). Methyl bromide-micronuclei induction of methyl bromide in rats and mice by subchronic inhalation test. Environ. Mut. Commun. 17,47-56. ATSDR (1991). "Toxicological Profile for Bromomethane." U.S. Department of Health and Human Services, Public Health Services Agency for Toxic Substances and Disease Registry, Publication 91-06. Bakhishev, G. N. (1973). Relative toxicity of aliphatic halohydrocarbons to rats. Farmakol Toksikol. 8, 140-142. [In Russian] Balander, P. A., and Polyak, M. G. (1962). Toxicological characteristics of methyl bromide. 1. Gig. I Toksikol. 60, 412-419.
Bentley, K. S. (1994). "Detection of Single Strand Breaks in Rat Testicular DNA by Alkaline Elution FoIIowing In Vivo Inhalation Exposure to Methyl Bromide." Unpublished report from E. I. DuPont Haskell Laboratories, Project 9714-001: MBIPI2I1ALKlHASK: 999. Bond, J. A., Dutcher, J. S., Medinsky, M. A., Henderson, R. E, and Bimbaum, L. S. (1985). Disposition of [l4C]methyl bromide in rats after inhalation. Toxicol. Appl. Pharmacol. 78,259-267. Breslin, W. J., Zablotny, C L., Bradley, G. J., Nitschke, K. D., and Lomax, L G. (I 990a). "Methyl Bromide Inhalation Teratology Probe Study in New Zealand White Rabbits." Unpublished study from Dow Chemical Company Toxicology Laboratory. Breslin, W. J., Zablotny, C L., Bradley, G. J., and Lomax, L. G. (I 990b). "Methyl Bromide Inhalation Teratology Study in New Zealand White Rabbits." Unpublished study from Dow Chemical Company Toxicology Laboratory. Danse, L H. J. C, van Velsen, E L, and vander Heijden C A. (1984). Methyl bromide: Carcinogenic effect in the rat forestomach. Toxicol. Appl. Pharmacol. 72, 262-271. Djalali-Behzad, G., Hussain, S., Ostermann-Golker, S., and Segerbaeck, D. (1981). Estimation of genetic risks of alkylating agents. VI. Exposure of mice and bacteria to methyl bromide. Mutat. Res. 84, 1-9. Driscoll, C D., and Hurley, J. M. (1993). "Methyl Bromide: Single Exposure Vapor Inhalation Neurotoxicity Study in Rats." Unpublished report from Bushy Run Research Center, Project 92N 1197. Eustis, S. L., Haber, S. B., Drew, R. T., and Yang, R. S. H. (1988). Toxicology and pathology of methyl bromide in F344 rats and B6C3FI mice following repeated inhalation exposure. Fundam. Appl. Toxicol. 11,594-610. Gargas, M., and Andersen, M. (1982). Metabolism of inhaled brominated hydrocarbons: validation of gas uptake results by determination of stable metabolite. Toxicol. Appl. Pharmacol. 66, 55-68. Garry, V. E, Nelson, R. L., Griffith, J., and Harkins, M. (1990). Preparation of human study of pesticide applicators: sister chromatid exchanges and chromosomal aberrations in cultured human Iymphocytes exposed to selected fumigants. Teratolog. Carcinog. Mutagen. 10,21-29. Gotoh, K, Nishizawa, T., Yamaguchi, T, Kanou, H., Kasai, T., Ohsawa, M., Ohbayyashi, H., Aiso, S., Ikawa, N., Yamamoto, S., Noguchi, T., Nagano, K., Enomoto, M., Nozaki, K., and Sakabe, H. (1994). Two year toxicological and carcinogenesis studies of methyl bromide in F344 rats and BDFl mice-Inhalation studies. In "Proceedings: Second Asia-Pacific Symposium on Environmental and Occupational Health." Haber et al. (1985). Hastings, L., Andringa, A., and Miller, M. A. (1994). Exposure of the olfactory system to toxic compounds: structural and functional consequences. Inh. Toxicol. 6, 437-440. Hine, C H. (1969). Methyl bromide poisoning: A review of ten cases. 1. Occup. Med. 11, 1-10. Honma, T., Miyagawa, M., Sato, M., and Hasegawa, H. (1985). Neurotoxicity and metabolism of methyl bromide in rats. Toxicol. Appl. Pharmacol. 81, 183-191. Hurtt, M. E., Morgan, K. T., and Working, P. K. (1987). Histopathology of acute toxic responses to selected tissues from rats exposed by inhalation to methyl bromide. Fundam. Appl. Toxicol. 9,352-365. Hurtt, M. E., Thomas, D. A., Working, P. K., Monticello, T. M., and Morgan, K. T. (1988). Degeneration and regeneration of the olfactory epithelium following inhalation exposure to methyl bromide: Pathology, cell kinetics and olfactory function. Toxicol. Appl. Pharmacol. 94,311-328. Irish, D. D., Adams, E. M., Spencer, H. C, and Rowe, V. K. (1940). The response attending exposure of laboratory animals to vapors of methyl bromide.l. Ind. Hyg. Toxicol. 22,218-230. Japanese Ministry of Labour (1992). "Toxicology and Carcinogenesis Studies of Methyl Bromide in F344 Rats and BDF Mice (Inhalation Studies)." Unpublished report from the Industrial Safety and Health Association, Japanese Bioassay Laboratory, Tokyo. Kato, N., Morinobu, S., and Ishizu, S. (1986). Subacute inhalation experiment for methyl bromide in rats. Ind. Health 24, 87-103. Katz, A. J. (1985). Genotoxicity of methyl bromide in somatic cells of Drosophila larvae. Environ. Mutagen. 7, 13.
References
Katz, A. J. (1987). Inhalation of methyl bromide gas induces mitotic recombination in somatic cells of Drosophila melanogaster. Mutat. Res. 192, 131-135. Kiplinger, G. A. (1994). "Methyl Bromide: Acute Oral Toxicity Comparison Study of Microencapsulated Methyl Bromide and Liquid Methyl Bromide in Albino Rats." Unpublished report from WIL Research Laboratories, Project WIL-490 11. Kramers, P. G. N., Voogd, C. E., Knaap, A. G. A. C., and Van der Heijden, C. A. (1985a). Mutagenicity of methyl bromide in a series of short term assays. Mutat. Res. 155,41-47. Kramers, P. G. N., Bissumbhar, B., and Mout, H. C. A. (1985b). Studies with gaseous mutagens in Drosophila melanogaster. In "Short Term Bioassays in the Analysis of Complex Environmental Mixtures IV" (M. D. Waters, S. S. Sandhu, J. Lewtas, L. Claxton, G. Straus, and S. Nesnow, eds.), pp. 6573. Plenum, New YorkILondon. McGregor, D. B. (1981). "Tier II Mutagenic Screening of 13 NIOSH Priority Compounds. Report 32. Individual Compound Report: Methyl Bromide." National Institute of Occupational Safety and Health, Cincinnati, OH, PB83-130211. Medinsky, M., Bond, J., Dutcher, J., and Birnbaum, L. (1984). Disposition of [14C]-methyl bromide in rats after inhalation. Toxicology 32, 187-196. Medinsky, M., Dutcher, J., Bond, J., Henderson, R., Mauderly, J., Snipes, M., Mewhinney, J., Cheng, Y., and Birnbaum, L. (1985). Uptake and excretion of [l4C]-methyl bromide as influenced by exposure concentration. Toxicol. Appl. Pharmacol. 78,215-225. Mertens, J. J. W. M. (1997). "A 24-Month Chronic Dietary Study of Methyl Bromide in Rats." Unpublished report from WIL Research Laboratories, Project WIL-49014. Moriya, M., Ohta, T., Watanabe, K., Miyazawa, T., Kato, K., and Shirasu, Y. (1983). Further mutagenicity studies on pesticides in bacterial reversion assay systems. Mutat. Res. 116, 185-216. Naas, D. J. (1990). "Acute Oral Toxicity Study in Beagle Dogs with Methyl Bromide." Unpublished report from WIL Research Laboratories, Inc., Project WIL-49006. Newton, P. E. (1994a). "An Up-and-Down Acute Inhalation Toxicity Study of Methyl Bromide in the Dog." UnpUblished report from Pharmaco-LSR, Project 93-6067. Newton, P. E. (1994b). "A Four Week Inhalation Toxicity Study of Methyl Bromide in the Dog." Unpublished report from Pharmaco LSR, Project 936068. Newton, P. E. (1995). "A Chronic (12-Month) Toxicity Study of Methyl Bromide Fumigated Feed in the Dog." Unpublished report from PharmacoLSR, Project 94-3186.
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Norris, J. c., Driscoll, C. D., and Hurley, J. M. (1993). "Methyl Bromide: Ninety-Day Vapor Inhalation Neurotoxicity Study in CD Rats." Unpublished report from Bushy Run Research Center, Project 92NII72. NTP (1992). "Toxicology and Carcinogenesis Studies of Methyl Bromide (CAS No. 74-83-9) in B6C3Fl Mice (Inhalation Studies)." National Toxicology Program Technical Report Series 385. Ong, J. M., Stewart, J., Wen, Y., and Whong, W. (1987). Application of SOS umu-test for the detection of genotoxic volatile chemicals and air pollutants. Environ. Mutagen. 9, 171-176. Reuzel, P. G. J., Kuper, C. F., Dreef-van der Meulen, H. C., and Hollanders, V. M. H. (1987). "Chronic (29-Month) Inhalation Toxicity and Carcinogenicity Study of Methyl Bromide in Rats." Unpublished Report From CIVO Institutes TNO. Reuzel, P. G. J., Dreef-van der Meulen, H. c., Hollanders, V. M. H., Kuper, C. F., Feron, V. J., and van der Heijden, C. A. (1991). Chronic inhalation toxicity and carcinogenicity study of methyl bromide in Wistar rats. Food Chem. Toxicol. 29,31-39. Sikov, M. R., Cannon, W. C., Carr D. B., Miller, R. A., Montgomery, L. F., and Phelps, D. W. (1981). "Teratologic Assessment of Butylene Oxide, Styrene Oxide and Methyl Bromide." Battelle Pacific Northwest Laboratories, Contract 210-78-0025, Division of Biomedical and Behavioral Science, National Institute for Occupational Safety and Health, U. S. Department of Health and Human Services. Simmon, V. F., Kauhanen, K., and Tardiff, R. G. (1977). Mutagenic activity of chemicals identified in drinking water. In "Progress in Genetic Toxicology" (D. Scott, B. A. Bridges, and F. M. Sobels, eds.), pp. 249-258. ElsevierlNorth-Holland Biomedical Press, Amsterdam. Tucker, 1. D., Xu, J., Stewart, J., Baciu, P. c., and Ong, T. (1986). Detection of sister chromatid exchanges induced by volatile genotoxicants. Teratog. Carcinog. Mutagen. 6, 14-21. WHO (1995). "Environmental Health Criteria 166, Methyl Bromide." Published under the joint sponsorship of the United Nations Environment Programme, the International Labour Organisation, and the World Health Organization, Geneva. Wilmer et al. (1983). Yamano, Y. (1991). Experimental study on methyl bromide poisoning in mice. Acute inhalation study and the effects of glutathione as an antidote. Jpn. J. Ind. Health 33: 23-30. [in Japanese] Zwart, A. (1988). "Acute Inhalation Study of Methyl Bromide in Rats." CIVO Rep. V88. 127127. CIVO Institutes, TNO, Zeist, The Netherlands. Zwart, A., Arts, J. H. E., Ten Berge, W. F., and Appelman, L. M. (1992). Alternative acute inhalation toxicity testing by determination of the concentration-time-mortality relationship: Experimental comparison with standard LC50 testing. Regul. Toxicol. Pharmacol. 15, 278-290.
CHAPTER
85 1,3-Dichloropropene w. T. Stott and B. B. Gollapudi Dow Chemical Company
85.1 CHEMISTRY AND FORMULATIONS 85.1.1 CHEMICAL NAME 1,3-Dichloro-l-propene is the chemical name. The pesticide is typically marketed either as a relatively pure eis-isomer or as a mixture of eis- (E) and trans- (Z) isomers.
85.1.2 STRUCTURE
H
Cl
>=
CICH2
trans-
keted under the trademarks Dorlone® (admixture with 1,2dibromoethane), D-D® Soil Fumigant, Nemex®, Telone®, and Vidden D®. More recent trademarks include Vorlex® (admixture with methylisothiocyanate), Di-Trapex®, D-D® Super, Telone® 11 Soil Fumigant, and Telone® C-17 Soil Fungicide and Nematicide (admixture with chloropicrin). Purified cisisomer 1,3-dichloropropene has also been marketed under the trademarks Telone-eis® and Nematrap®. Formulations have typically contained 1-2% of an acid scavenger. Older formulations were often stabilized using epichlorohydrin while epoxidized soybean oil has been utilized in more recent formulations.
CICH2 Cl
A
H
H
85.2 USES
cis-
85.1.3 SYNONYMS 1,3-Dichloropropene also is known as a-chloroallyl chloride and 1,3-dichloropropylene. The CAS registry number for 1,3dichloropropene is 542-75-6. The number for the trans-isomer is 10061-02-6; that of the eis-isomer is 10061-01-5.
85.1.4 PHYSICAL AND CHEMICAL PROPERTIES 1,3-Dichloropropene has the empirical formula C3H4Cl2 and a molecular weight of 110.98. It is a white to amber-colored liquid with a sweet-penetrating odor. The density at 25°C is 1.217. The boiling points of the cis- and trans-isomers are 104 and 112°C, respectively. The flash point is 28°C. The solubility in water at 25°C is approximately 2 gikg. The compound is miscible with acetone, benzene, carbon tetrachloride, heptane, and methanol.
Introduced in 1945, 1,3-dichloropropene is a soil fumigant nematocide, for preplanting control of parasitic plant nematodes in numerous food and nonfood crops including deciduous fruit and nuts, vines, strawberries, field crops, vegetables, tobacco, tree nurseries, and numerous other specialty crops. Formulated 1,3-dichloropropene is injected into the soil using chisels prior to crop planting at a minimum depth of 10-12 inches below the soil surface. The injected zone is subsequently capped off with soil which is often then covered with plastic to help maintain concentrations in the soil. Injected 1,3-dichloropropene is believed to volatilize, move through the soil air space, and redissolve into the film of water that surrounds soil particles, where it may exert its toxic effect on soil nematodes. 1,3Dichloropropene in the soil is lost via chemical hydrolysis in water, metabolism by soil biotic a, and evaporation. Efficacy is thus dictated not only by target organism sensitivity, but also by the vapor pressure, diffusion coefficient, distribution in air, water, and soil phases of the soil matrix, and the temperature and moisture content of the treated soil.
85.3 HAZARD IDENTIFICATION
85.1.5 FORMULATIONS
85.3.1 ACUTE TOXICITY
Commercial formulations, containing varying amounts of a mixture of cis- and trans-isomers, have historically been mar-
1,3-dichloropropene is irritating to eyes and skin of animals. As reviewed by Torkelson (1994), application of mixed isomers of
Handbook of Pesticide Toxicology Volume 2. Agents
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Copyright © 200 I by Academic Press. An rights of reproduction in any fonn reserved.
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CHAPTER 85
1,3-Dichloropropene
1,3-dichloropropene to the skin of rabbits up to 4 hours (occluded application site) caused a moderate erythema and moderate to severe edema. Mixed isomers of 1,3-dichloropropene also caused a marked redness and slight to moderate chemosis of the conjunctivae immediately following instillation of 0.1 mL into the eyes of rabbits. These effects were gradually reversible and, in most instances, washing with water was effective in averting injury. Similar dermal and ocular effects have been reported for cis-l ,3-dichloropropene (Gardner, 1989) and a formulation of 1,3-dichloropropene containing approximately 30% 1,2-dichloropropane (D-D®) (Hine et aI., 1953). Both mixed isomer 1,3-dichloropropene and cis-l ,3-dichloropropene have tested positive in Guinea pig skin sensitization assays (Gardner, 1989; lones, 1988a; Torkelson, 1994). The acute oral, dermal, and inhalation lethality of both mixed isomers of 1,3-dichloropropene and cis-l ,3-dichloropropene in laboratory animals have also been established (Gardner, 1989; Hine et al., 1953; lones, 1988b, c; lones and Collier, 1986a, b; Nitschke et aI., 1990a; Torkelson, 1994). The oral LDso of mixed isomers of 1,3-dichloropropene in rats ranges from 130 to 713 mg/kg in males and 110-250 to 510 mg/kg in females dependent upon vehicle and strain used. The oral LDso for cis-l ,3-dichloropropene in rats has been calculated to be 85 to 126 and 117 in males and females, respectively. The dermal LDso has been reported as greater than 1211 mg/kg mixed isomers in both sexes of rats (unoccluded application site); 1000 and 1300-2000 mg/kg mixed isomers in male and female rats, respectively (occluded application site); 333-540 mg/kg mixed isomers for both sexes of rabbits (occluded application site); and 758-1090 mg/kg cis-l,3-dichloropropene for both sexes of rats (occluded application site). The acute 4-hour LCso value for mixed isomers of 1,3-dichloropropene vapor in rats was 855-1035 ppm for males and 904 ppm for females. Exposed animals had a distinct "garlic" odor and suffered eye and nasal irritation. The acute 4-hour LCso value for cis-l,3dichloropropene vapor in rats was 670 and 744 ppm for males and females, respectively. Grossly observable eye and respiratory tract irritation was absent following a 2-week observation period. Brief exposure to concentrations in excess of 2700 ppm mixed isomer vapor also caused severe lung, liver, and kidney injury.
85.3.2 REPEATED DOSE TOXICITY The subacute and subchronic toxicity of 1,3-dichloropropene has been examined both orally, via gavage or mixed in feed, and via inhalation. Subacute and subchronic oral toxicity studies on relatively modem formulations have been carried out in rats, mice, and dogs by stabilizing the I ,3-dichloropropene by microencapsulation in a starch-sucrose matrix and then mixing the encapsulated material into the feed of test animals. Studies reported by Stott et al. (1988) have demonstrated the ready bioavailability of this material once ingested by test animals. Relatively short-term oral dietary toxicity studies employing microencapsulated 1,3-dichloropropene were conducted as
a prelude to the subchronic studies summarized below (Haut et aI., 1992a, b). Male and female Fischer 344 rats and B6C3Fl mice were administered dosages of 10, 25, 50, or 100 (rats) or 175 (mice) mg/kg/day of microencapsulated 1,3dichloropropene (mixed isomers) via their diet for 2 weeks. The body weights of both sexes of rats ingesting 2:50 mg/kg/day and male mice and female mice ingesting 2: 100 and 175 mg/kg/day, respectively, were decreased. Histopathological changes were restricted to rats and consisted of hyperplasia and hyperkeratosis of the nonglandular mucosa of the stomachs of both sexes of rats ingesting 2: 50 mg/kg/day and a single male ingesting 25 mg/kg/day. In a subsequent subchronic rat study, male and female Fischer 344 rats were administered dosages of 5, 15, 50, or 100 mg/kg/day of microencapsulated 1,3-dichloropropene (mixed isomers) via their diet for 13 weeks (Haut et aI., 1996). The body weights of males and females ingesting 2:5 and 2: 15 mg/kg/day, respectively, were decreased. A number of changes in serum biochemical parameters and decreases in organ weights accompanied the depressed body weights of these animals. Histopathological changes were restricted to basal cell hyperplasia and/or hyperkeratosis of the nonglandular mucosa of the stomach of both sexes ingesting 2:50 mg/kg/day and a single male ingesting 15 mg/kg/day. These changes were at least partially reversible upon ingestion of control feed for 4 weeks. In a subsequent subchronic mouse study, male and female B6C3Fl mice were administered dosages of 15, 50, 100, or 175 mg/kg/day of microencapsulated 1,3-dichloropropene (mixed isomers) via their diet for 13 weeks (Haut et aI., 1996). A doserelated decrease in the body weights of males and females ingesting 2:50 mg/kg/day was observed. Histologic changes consistent with decreased cytoplasmic glycogen and with decreased lipid content were observed in the liver of all treated mice and the kidneys of high dose group mice. No treatmentrelated histopathologic effects were reported. Male and female Beagle dogs were administered microencapsulated 1,3-dichloropropene via their diets for 13 weeks at concentrations which resulted in mean dosages of 5, 15-16, or 41 mg/kg/day (Stebbins et aI., 1999). The body weights of both sexes of dogs were decreased in a dose-related manner relative to controls. The primary effect of 1,3-dichloropropene ingestion was upon erythroid parameters measured in peripheral blood. Calculated erythroid indices and morphologic changes in stained peripheral blood of male and females ingesting 2:15-16 mg/kg/day indicated the presence of a hypochromic, microcytic anemia. In an early oral toxicity study, male and female Wistar rats were administered dosages of 1, 3, 10, or 30 mg/kg/day of a roughly 78% pure mixed isomer 1,3-dichloropropene formulation 6 days/week for 13 weeks via oral gavage (see Stott et al., 1988; Til et aI., 1973). The kidney weights of both sexes of high dose group animals and males administered 10 mg/kg/day mixed isomer 1,3-dichloropropene were elevated relative to controls. However, no gross or histopathologic changes or alterations in hematologic indicies, urinaly-
85.3 Hazard Identification
sis, or serum enzymes accompanied these changes. It was not clear whether toxicity was dictated by 1,3-dichloropropene or of some impurity, possibly 1,2-dichloropropane, which was present at a concentration of nearly 20%. The latter chemical reportedly causes increased liver and kidney weights and hepatic histopathologic changes in rats (Bruckner et aI., 1989; IPCS, 1993). The toxicity of inhaled 1,3-dichloropropene in several formulations has also been evaluated over the years. Formulations of >90% purity have been studied in which both sexes of Fischer 344 rats and CD-l mice were exposed 6 hours/day, 5 days/weck to mixed-isomer vapor concentrations of 5, 10, or 30 ppm for 4 weeks; 10, 30, or 90 ppm for 13 weeks; and 10, 30, 90, or 150 ppm for 13 weeks (Coate, 1979a, b; Stott et aI., 1988). No treatment-related effects were observed in rats or mice following a 4-week exposure to up to 30 ppm vapor. However, decreases in body weights were noted and the nasal mucosa and urinary bladder (female mice) were identified as potential target tissues of inhaled 1,3-dichloropropene for 13 weeks. Nasal effects consisted of degeneration of olfactory epithelium and/or hyperplasia of respiratory epithelium in both sexes of rats (::::30 ppm) and mice (::::90 ppm) and respiratory metaplasia in olfactory regions of mice (150 ppm). Bladder effects consisted of hyperplasia of the transitional epithelium in female mice only (::::90 ppm). Subsequent studies were undertaken in which Fischer 344 rats were similarly exposed to cis-l ,3-dichloropropene vapor concentrations of 10, 60, or 150 ppm for 2 weeks or 10, 30, or 90 ppm for 13 weeks (Nitschke et al., 1990b; Nitschke and Lomax, 1990). These latter studies also identified body weight changes and/or histopathological changes in the nasal respiratory and olfactory epithelium in both sexes of rats (::::6090ppm). In contrast, in an early inhalation study conducted in 1958, male and female rats, guinea pigs, rabbits, and dogs (females only) were exposed to 1 or 3 ppm mixed isomer 1,3dichloropropene vapor 7 hr/day, 5 days/week, for 6 months (Torkelson and Oyen, 1977). The only effect of exposure which was reported was a "foamy vacuolation or proximal tubule epithelium" in high exposure group male rats. The significance of this change has been questioned in view of findings from more recent studies and the diagnosis was characteristic of nephropathy endemic in the site-bred rats used during this early period (Stott et aI., 1988). No effects were reported in female rats or other animals exposed to 1 ppm vapor. Exposure of rats and mice to 5, 15, or 50 ppm D-D®, an admixture of 1,3-dichloropropene and 1,2-dichloropropene, for 12 weeks revealed liver and kidney weight changes in high exposure male and female rats, respectively, and diffuse hepatocellular swelling in high exposure male mice (urinary bladders were not examined) (Parker et al., 1982). As noted, increased liver and kidney weights and hepatic histopathologic changes have been a consistent finding in 1,2-dichloropropene toxicity studies (Bruckner et aI., 1989; IPCS, 1993).
1851
85.3.3 EFFECTS ON REPRODUCTION
Mixed isomer vapors of 1,3-dichloropropene were not embrytotoxic or teratogenic in bred rats or inseminated rabbits exposed to 20, 60, or 120 ppm vapors, 6 hours/day, during gestation days 6-15 (rats) or 6-18 (rabbits) (Hanley et aI., 1987). Maternal toxicity was evidenced at all exposure levels. Exposure of male and female rats to 10, 30, or 90 ppm 1,3-dichloropropene vapors for two generations did not adversely affect reproduction or neonatal growth or survival even though 90 ppm proved to be a toxic exposure level (Breslin et aI., 1989). Consistent with these results, no treatment-related changes in testes weight, sperm count, or sperm morphology occurred in mice 30 days after being injected i.p. with 10, 19, 38, 75, 150, 300, or 600 mg/kg/day 1,3-dichloropropene daily, for 5 days (Osterloh and Feldman, 1993). Finally, exposure of male and female rats to 14,32, or 96 ppm D-D®, a low purity 1,3-dichloropropene formulation containing approximately 30% 1,2-dichloropropane, 6 hours/day, for 10 weeks did not affect animal mating behavior or fertility (Linnett et aI., 1988). 85.3.4 ABSORPTION, DISTRIBUTION, METABOLISM, AND EXCRETION
Toxicity and pharmacokinetic data indicate that 1,3-dichloropropene is absorbed from the skin, respiratory tract, and gastrointestinal tract. Both inhaled and ingested eis- and transisomers were rapidly eliminated from the bloodstream of rats in a biphasic manner consisting of a prominent initial phase with a half-life of approximately 4-7 minutes followed by a slower phase with a half-life of 22-43 minutes (Stott and Kastl, 1986; Stott et al., 1988). The predominant routes of excretion of radioactivity in male and female rats following a single or repeated oral dose(s) of eis, trans, or mixed 1,3-dichloropropene were via the urine (eis, 82-84%; trans, 56-61 %; mixed, 5161%), feces (eis or trans, 2-3%; mixed, 17-21%), and expiration of C02 (eis, 2-5%; trans, 23-24%; mixed, 15%) (Bartels et aI., 1999; Dietz et aI., 1984a; Hutson et aI., 1971; IPCS, 1993). Excretion and distribution of 1,3-dichloropropene was independent of dose in rats and mice administered up to 50 and 100 mg/kg, respectively (Dietz et aI., 1984a). In both species, greater than 80% of the administered dosages were excreted within 24 hours of dosing. There were no remarkable sexrelated differences in excretion routes, kinetics of excretion, or tissue distribution of administered radioactivity in rats (Bartels et aI., 1999; Hutson et aI., 1971). Humans also rapidly metabolize and excrete inhaled 1,3-D vapor (Waechter et aI., 1992). Human volunteers exposed to 1.0 ppm mixed isomer vapor for 6 hours asorbed roughly 80% of inhaled 1,3-D. Blood concentrations of 1,3-D rapidly fell following postexposure, resulting in no quantifiable concentrations by the first sampling time point post exposure, 10 minutes, establishing a blood half-life of less than this value. Approximately 90% of the estimated dose of inhaled Telone 11 was excreted within 36 hours of exposure. Dermal absorption of
1852
CHAPTER 85
1,3-Dichloropropene
1,3-D vapor does not appear to be a factor in exposed humans as whole-body uptake has been estimated to be roughly only 2-5% of that absorbed via inhalation (Kezic et aI., 1996). The mercapturic acid conjugate of 1,3-dichloropropene and its further oxidation product, a sulfoxide, were the primary excretion products identified in the urine of treated animals (Bartels et aI., 1999; Climie et aI., 1979; Dietz et aI., 1984a; Fisher and Kilgore, 1988a). Approximately 32-36% of acute 5 or 50 mg/kg oral doses of mixed isomers were excreted in male rats in an isomeric ratio of cis- to trans-mercapturate of approximately 4: 1 (Bartels et aI., 1999; Dietz et aI., 1984a). Repeated dosing of rats at 5 mg/kg/day for several weeks appeared to increase the percentage excretion of these metabolites to roughly 43% of the dose (Bartels et aI., 1999). In contrast, Onkenhout et al. (1986) reported that approximately 45-55% of a range of dosages from 0.05 to 4.5 mg/kg mixed isomers administered via interperitoneal injection to rats were excreted as the mercapturate conjugate at an isomeric ratio of only 1.2: 1. Several additional glutathione conjugate degradates have been observed in the urine of rats and, to a greater extent, mice (Bartels et aI., 1999; Dietz et aI., 1984a). These have included the mercaptoacetate, mercaptopyruvate, and cysteine conjugates of 1,3-dichloropropene. The mercapturic acid conjugate of 1,3-D has also been identified in the urine of human subjects exposed to mixed isomer vapors of 1,3-D under field application or laboratory conditions and has been utilized as a biomonitor of 1,3-D exposure (Kezic etal., 1996; Osterloh et aI., 1984; Osterloh and Feldman, 1993; van Welie et aI., 1991; Waechter et aI., 1992). Consistent with the earlier rat data, Waechter et al. (1992) found that humans excrete approximately 45% and 14% of absorbed cis- and trans-isomers of 1,3-dichloropropene, respectively, as the cisand trans-mercapturates, a 3.2: 1 ratio. Based upon these data, it has been proposed that 1,3dichloropropene is primarily metabolized in rats and mice, and likely humans, by conjugation with glutathione and by hydrolysis of the 3-position chlorine (Dietz et aI., 1984b; Hutson et aI., 1971; IPCS, 1993; Onkenhout et aI., 1986). The end product of the latter pathway is C02. Several other minor metabolites suggestive of oxidative metabolism (epoxidation) have also been reported in the liver of rats administered a very high, lethal, dosage of 1,3-dichloropropene (700 mg/kg) via i.p. injection (Schneider et aI., 1998). However, a disproportionately lower yield of epoxide was reported at a lower, nonleathal, i.p. dosage and none was detected upon oral dosing (Bartels et aI., 1999). The toxicity of 1,3-dichloropropene thus appears to reflect the balance between inherent chemical reactivity and competing enzymatic activation, and spontaneous and enzymatic detoxification pathways, some or all of which are saturable. A relatively small amount of mixed isomers have been found to bind to macromolecules in the forestomach and, to a lesser extent, in the glandular stomach of rats and mice following acute oral dosing with 1,3-dichloropropene (Dietz et aI., 1984b). Macromolecular binding correlated with a dose and time-related depression in the nonprotein sulfhydryl content, presumably of glutathione, as a result of direct or enzymatic,
of stomach and liver tissues of these latter animals. Similar exposure-related decreases in the sulfhydryl content of a number of tissues of rats inhaling 1,3-dichloropropene vapor for one hour have also been reported (Fisher and Kilgore, 1988b). Initial losses in sulfhydryllevels, however, may be offset somewhat by a significant rebound effect which was observed in livers of rats and lungs of mice repeatedly administered 1,3dichloropropene via oral gavage and inhalation, respectively (W. Stott, unpublished data). Indeed, no DNA adducts were reported by Gollapudi et al. (1999) in a relatively sensitive 32P-postlabeling assay of these same hepatic and pulmonary tissues and Schneider et al. (1998) reported a seven-fold decrease in formation of epoxide in vitro in the presence of glutathione. Significantly, the net activities of glutathione-Stransferase isozymes to metabolize 1,3-dichloropropene also appear to determine responses in in vitro assays of genotoxicity. In general, target organisms utilized in in vitro assays, especially bacteria, conjugate 1,3-dichloropropene very poorly relative to mammalian tissue extracts, especially when the latter are fortified with physiological levels of glutathione (Creedy et aI., 1984; Stott, et aI., 1992). 85.3.5 SHORT TERM ASSAYS 1,3-Dichloropropene has been tested in a wide variety of genotoxicity assays with variable results. A number of early in vitro assays reporting positive responses for 1,3-dichloropropene in bacteria (De Lorenzo et aI., 1977; Eder et aI., 1982; Neudecker et aI., 1977; Stolzenberg and Hine, 1980) were confounded by the presence of mutagenic impurities and/or stabilizing agent (e.g., epichlorohydrin) or by the generation of mutagenic oxidation products during gas chromatographic purification procedures (Talcott and King, 1984; Watson et aI., 1987). Subsequent studies, using 1,3-dichloropropene purified by passage through a silicic acid column, were weakly positive only in the presence of liver microsomes (100,000 x g pellet) from rats induced with polychlorinated biphenyls. Addition of cytosolic fractions, which contain glutathione-S-transferase, and physiological levels of glutathione, eliminated activity. Metabolically active fractions (S9) obtained from the lungs or kidneys of naIve mice or from the lungs of mice repeatedly exposed to 1,3-dichloropropene vapor also did not metabolize purified 1,3-dichloropropene to a mutagen in Salmonella mutagenicity assays (Gollapudi et aI., 1999; Stott, et aI., 1992). Assays employing mammalian cell lines have also resulted in mixed evidence of genotoxicity. Negative results have been obtained using an epoxidized soybean oil-stabilized mixed isomer formulation of 1,3-dichloropropene in the Chinese hamster ovary (CHO) HGPRT forward mutation and rat hepatocyte unscheduled DNA synthesis assays (Gollapudi et aI., 1999). 1,3-Dichloropropene samples of unknown purity or stabilizing agent were negative in chromosomal aberration assays in CHO cells, and rat liver cell lines (Dean et aI., 1985; Loveday et aI., 1989) and negative sister chromatid exchange assays in V79 lung fibroblasts with S9 have also been reported (Loveday et aI., 1989). However, similar samples of single or mixed
85.3 Hazard Identification
isomeric 1,3-dich10ropropene induced unscheduled DNA synthesis in HeLa cells (Schiffman et aI., 1983), sister chromatid exchanges in V791ung fibrob1asts (von der Hude et aI., 1987), DNA fragmentation and repair in V79 cells and rat and human hepatocytes (Dean et aI., 1985; Martelli et aI., 1993), chromosomal aberrations in CHL and CHO cells (Loveday et aI., 1989), and mutations at the tk locus in L5178Y mouse lymphoma cells (Myhr and Caspary, 1991). In contrast to the results of in vitro genotoxicity assays, in vivo assays of 1,3-dichloropropene have been generally negative. Negative results were obtained in mouse bone marrow micronucleus assays using relatively high oral or i.p. dosages of 1,3-dichloropropene (Gollapudi et al., 1999; Shelby et aI., 1993). However, Kevekordes et al. (1996) reported that 1,3-dich10ropropene induced micronuclei in bone marrow erythrocytes of female, but not male, mice. Inhaled 1,3dichloropropene vapors did not cause dominant lethal effects in rat germ cells in two separate assays (Gollapudi et aI., 1998; Linnett et aI., 1988) nor did they cause point mutations in somatic tissues (lung and liver) of Big Blue™ transgenic mice (Gollapudi et aI., 1999). In addition, negative results have been obtained for a 1,3-dichloropropene formulation (stabilizer unknown) in several host-mediated bacterial mutagenicity assays in mice (Shirasu et aI., 1981; Sudo et aI., 1979). A 32p_ postlabelling assay of liver tissue from rats dosed orally and of lung tissue of mice exposed via inhalation (target tissues for tumor formation) for the potential formation of DNA adducts was also negative (Gollapudi et aI., 1999). Positive results, however, have been reported to cause single strand breaks in the DNA of several tissues and DNA repair in hepatocytes of rats following oral or intraperitoneal dosing (Ghia et aI., 1993; Kitchin and Brown, 1994). Finally, when fed to Drosophila at a high concentration, an epoxide-stabilized 1,3-dichloropropene formulation caused an increased incidence of sex-linked recessive lethal mutations but not reciprocal translocations (Valencia et aI., 1985). 85.3.6 CHRONIC TOXICITY AND ONCOGENICITY ASSAYS Chronic toxicity and oncogenicity studies of 1,3-dichloropropene via several routes have been conducted. In a recent study using an epoxidized soy bean oil stabilized and weakly in vitro mutagenic formulation of 1,3-dichloropropene, Fischer 344 rats, and B6C3F1 mice were administered dosages of 2.5, 12.5, or 25 mg/kg/day and 2.5, 25, or 50 mg/kg/day, respectively, as a microencapsulated preparation via their diet, 7 days/wk, for 2 years (Stott et aI., 1996). In both sexes of rats and mice, body weights were decreased and hyperplasia of the nonglandular stomach mucosa was reported to occur in a dose-related manner. An increased incidence of foci of altered cells was also noted in the livers of treated rats following 24 months dosing. The only tumorigenic response observed in rats was an increase in the incidence of benign liver tumors in high dose males and females and intermediate dose males. No tumorigenic response was reported in either sex of mice.
1853
In contrast, an early study involving the administration of an older, highly in vitro mutagenic, mixed-isomerepichlorohydrinstabilized formulation of 1,3-dichloropropene to rats (25 or 50 mg/kg/day) and mice (50 or 100 mg/kg/day) via gavage, 3 days/week, for up to 2 years resulted in increases in several benign and malignant tumor types in both species (NTP, 1985; Yang et aI., 1986). These included forestomach and liver tumors in male rats (25 or 50 mg/kg/day or both), forestomach tumors in female rats (50 mg/kg/day), and forestomach, lung, and urinary bladder tumors in female mice (50 or 100 mg/kg/day or both). The gavage bioassay in male mice was judged to be an "inadequate study of carcinogenicity" due to excessive early mortality of controls. Nontumorigenic responses were limited to hyperplasia of the nonglandular portion of the stomachs of mice and rats and hyperplasia of the urinary bladder epithelium of mice. The chronic toxicity of orally administered epoxidized soybean oil stabilized formulation of 1,3-dichloropropene has also been evaluated in male and female Beagle dogs administered 0.5, 2.5, or 15 mg/kg/day of microencapsulated 1,3-dichloropropene via their diets for one year (Stebbins et aI., 1999). The primary effect in both sexes of dogs ingesting 15 mg/kg/day 1,3-dichloropropene was a regenerative, hypochromic, microcytic anemia. Histologic changes in bone marrow and spleen consistent with increased hematopoiesis and extramedullary hematopoisis were consistent with this diagnosis. The latter changes along with increases in reticulocytes in these animals confirmed the regenerative nature of this effect. Anemia was observed following 3 months of dosing and remained relatively constant or improved somewhat over the remainder of the dosing period. The only other treatment-related effect observed in the study was a slight inflammation of the tongue of several high dose males suggestive of irritation by ingestion of 1,3-dichloropropene. Inhalation exposure of rats and mice to 5, 20 or 60 ppm of an epoxidized soybean oil stabilized formulation of 1,3dichloropropene 6 hours/day,S days/week, for 2 years resulted in non tumorigenic lesions of the nasal mucosa in both sexes of rats and mice exposed to 60 ppm and female mice exposed to 20 ppm, the urinary bladder epithelium of both sexes of mice exposed to 60 ppm and the forestomach of male mice exposed to 60 ppm (Lomax et aI., 1988). Slight changes in the morphology of renal and hepatic tissues of male and female mice exposed to 60 ppm, respectively, indicative of decreased lipid and glycogen content, respectively, were also observed. An increased incidence of benign lung tumors in high exposure group male mice was the only tumorigenic response observed (44% vs 18% in controls; historical incidence of adenomas = 7-32%). Cis-l,3-dichloropropene was negative in a mouse skin initiation-promotion bioassay when tested with phorbol myristate as a promoter and was not carcinogenic following repeated dermal application of 122 mg to the backs of Ha : ICR Swiss mice 3 times/week for up to 85 weeks (Van Duuren et al., 1979). An increase in local fibrosarcomas (6 of 30 mice vs 0 of 30 controls) was reported in mice following repeated subcutaneous injections of 3 mg/animal/week (lIweek) for up to 83 weeks.
1854
CHAPTER 85
1,3-Dichloropropene
85.4 DOSE-RESPONSE The toxicity of 1,3-dich10ropropene in a number of laboratory animals displays both a clear dose-response and clearly defined no-ob served-effect levels (NOELs). This is true for both acute and repeated dose (subacute, subchronic, and chronic) nonneoplastic treatment-related effects as well as neoplastic effects in rodents upon chronic exposure (see above discussion of individual studies). A number of toxicity studies with their lowest effect levels (LEL), NOELs, and target tissues/effect at the LELs are summarized in Table 90.1. It can be seen that the most sensitive target tissues/effects (i.e., those observed at LELs) are consistent between studies of differing durations for a given species of test animal. Increasing the duration of the dosing period from 2 to 13 weeks to 2 years did not appear to significantly change the potential toxicity of 1,3-dich10ropropene relative to nonneoplastic pathological effects. Changes in NOELs or no-observed-adverse-effect levels (NOAELS) obtained in studies in which 1,3-dichloropropene was ingested or inhaled were generally within less than half a log unit of each other. An exception to this was the series of oral dietary toxicity studies in mice in which body weight depression was the major treatmentrelated change noted. In this case, more significant duration of treatment dependent decreases in NOELs occurred. The most sensitive treatment-related effects observed in animals ingesting or inhaling 1,3-dichloropropene were quite similar between studies for a given species and method of administration. Affected tissues often represented portal-ofentry tissues, for example gastric mucosa for ingested and nasal mucosa for inhaled 1,3-dichloropropene, consistent with the irritant nature of this chemical. Exceptions to this were the occurrence of hyperplasia of the transitional epithelium of the urinary bladders of both sexes of mice inhaling vapors or dosed orally via gavage, and anemia in dogs ingesting 1,3-dichloropropene. In most bioassays, tumorigenic responses, when present, have involved portal-of-entry tissues. Ingestion or inhalation of relatively recent formulations of 1,3-dichloropropene resulted in tumors in the livers, often regarded as a portal-of-entry tissue of the enteric tract, of rats and the lungs of male mice, respectively. However, chronic gavage of an older and highly mutagenic, epoxide stabilized formulation of 1,3-dichloropropene, while causing forestomach and liver tumors in rats and mice, also caused urinary bladder and lung tumors in mice. The effect of 1,3-dichloropropene upon portal-of-entry tissues is consistent with a direct toxicity of the molecule and its removal by saturable metabolic pathways. As noted, a major pathway for 1,3-dichloropropene metabolism is via glutathione-S-transferase dependent conjugation with glutathione. As reviewed by Watson et al. (1987) and in the IPCS (1993) review, this metabolism provides for practical thresholds in the dose-response of toxicity, and even mutagenicity, of 1,3-dichloropropene in test organisms. The saturation of this pathway may result in a nonlinear elevation in concentrations of 1,3-dichloropropene in cells of an in vitro mutagenicity or clastogenicity assay or tissues of an exposed animal, especially
in portal-of-entry tissues, and subsequent toxicological consequences.
85.5 TOXICOLOGY IN HUMANS 85.5.1 EXPERIMENTAL EXPOSURE
Seven out of 10 volunteers detected 1,3-dichloropropene at an air concentration of 3 ppm; some reported fatigue of the sense of smell after a few minutes. The same proportion of volunteers detected 1 ppm, but the odor was noticeably fainter (Torkelson and Oyen, 1977). In a population of 22 persons, the concentration at which odor was detected was 4.4 ± 3.1 ppm (mean ± S.D.) (Rick and McCarty, 1987). 85.5.2 ACCIDENTAL POISONING
Forty-six people were treated for exposure to 1,3-dichloropropene fumes following a traffic accident in 1975 involving spillage of 4500 liters of a formulated product (Flessel et aI., 1978). Twenty-four of these, 3 of whom had lost consciousness, were hospitalized overnight with symptoms including headache, irritation of mucous membranes, and chest discomfort. All patients took showers and were given intravenous fluids and three received oxygen and corticosteroids because of chest pain and cough. Eleven of 41 persons tested had slightly higher than average serum SGOT and/or SGPT values which reverted to normal within 48-72 hours, except for 5 which still had slightly higher than average SGOT values. Follow-up interviews with patients 1-2 weeks later revealed symptoms including headache, abdominal, and chest discomfort and malaise. One was diagnosed as having had pneumonia. Symptoms were reported more frequently in those most heavily exposed to the fumes. Patient interviews conducted approximately two years after the accident revealed complaints of headache, chest pain or discomfort, and "personality changes" (fatigue, irritability, difficulty in concentrating, or decreased libido). Two had undergone cardiac catheterizations but their arteriograms were normal. There was no correlation of these long-persisting symptoms with intensity of exposure. Two fatalities involving 1,3-dichloropropene have been confirmed. Accidental ingestion of D-D® (admixture with 1,2dichloropropane) resulted in abdominal pain and vomiting, muscular twitching, pulmonary edema, and death (Gosselin et aI., 1976). Accidental ingestion of TELONE n® by a farm worker in Spain resulted in abdominal pain and vomiting, adult respiratory distress syndrome, hematologic changes, hepatorenal impairment, muscular twitching followed by coma, and death (Hernandez et aI., 1994). A possible association between overexposure to 1,3-dichloropropene and the development of hematologic malignancy has also been suggested by Markovitz and Crosby (1984). The latter was based upon the development of histiocytic lymphoma in a farm worker and two firemen accidentally exposed (acute) to high levels TELONE n® and D-D®,
85.5 Toxicology in Humans
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Table 85.1 Summary of Lowest Effective Levels, No-Observed-Effect Levels, and Most Sensitive Treatment-Related Effects for Toxicity Studies of Mixed and eis-Isomer 1,3-Dichloropropene Formulations Target Route species
tissue at
Study Sex
duration
LEU
NOELb
LEL
Reference
Males
2 Weeks
25 mkdc
10 rnkd
Bodywt.,
Haut et al. (1992a)
Oral (diet) Rats
gastric mucosa Females
50 rnkd
25 rnkd
Bodywt., gastric mucosa
Males
13 Weeks
15 mkd
5* rnkd
Body wt., gastric mucosa
15 rnkd
Females
5 rnkd
Haut et al. (1996)
Bodywt., gastric mucosa
Males
2 Years
12.5 rnkd
2.5 rnkd
Bodywt., gastric mucosa, liver ADd
Females
12.5 rnkd
2.5 rnkd
Bodywt., gastric
Stot! et al. (1996) Stott et al. (1996)
mucosa Mice
Males
2 Weeks
100 mkd
50 rnkd
Bodywt.
Haut et al. (1992b)
Females Males and
13 Weeks
175 rnkd
100 rnkd
Bodywt.
50 rnkd
15* rnkd
Body wt.
females Males and
2 Years
25 rnkd
2.5 rnkd
Body wt.
females Dogs
Males
Haut et al. (1996) Haut et al. (1996)
13 Weeks
15 rnkd
5* rnkd
Anemia
Stebbins et al. (1999)
Females
16 rnkd
5* mkd
BW,
IS rnkd
2.5 rnkd
Anemia
anemia Males and
I Year
females
Stebbins et al. (1999)
Oral (gavage) Rats
Males
2 Years
25mkd
NDe
Gastric mucosa,
NTP (1985)
liver AD Females
25 rnkd
ND
Gastric mucosa,
stomach AD Mice
Females
2 Years
50mkd
ND
Gastric and u. bladder
NTP (1985)
(cA!)
mucosa, lung AD and CA
(continues)
1856
CHAPTER 85
1,3-Dichloropropene
Table 85.1 (continued) Target Route species
tissue at
Study Sex
duration
LEL
NOEL
LEL
Males and
4 Weeks
ND
30ppm
ND
Reference
Inhalation Rats
females Males
Coate (1979a)
13 Weeks
30ppm
90ppm
Body wt.,
Coate
nasal
(1979b);
mucosa
Stott et al. (1988)
Females
30ppm
IOppm
Nasal
30ppm
lOppm
Body wt.,
Stott et a!.
(nasal
(1988)
mucosa Males and
13 Weeks
(90 ppm)g
females
mucosa) Males and
2 Years
60ppm
20ppm
females Mice
Males and
Body wt.,
(1988)
4 Weeks
ND
30ppm
ND
Coate
13 Weeks
90ppm
30ppm
Body wt.,
Coate
females Males and
(1979a)
females
Males and
Lomax eta!'
nasal
13 Weeks
30ppm
IOppm
females
nasal
(1979b);
mucosa
Stott et a!.
(females)
(1988)
Body wt.,
Stott et al.
nasal & u.
(1988)
bladder mucosa Males and
2 Years
20ppm
5 ppm
females
Nasal
Lomax eta!'
mucosa
(1988)
*Signifies no-observed-adverse-effect level. a Lowest effect level.
bNo-observed-effect level. C mkd = mg/kg/day. dAD = adenoma (bening tumors). eND = not determined. f CA = carcinomas (malignant tumors). gExposure level in parenthesis reflects exposure level at which effects on nasal mucosa were observed.
respectively; however, it was subsequently established that the farm worker had leukemia prior to the incident (public records, State of California, Court of Appeal, Case No. 28344). 85.5.3 USE EXPERIENCE
1,3-Dichloropropene causes edema, redness, and necrosis of the skin (Torkelson and Oyen, 1977) and in one documented case was believed to have caused a contact hypersensitivity in a repeatedly exposed farmer (Nater and Gooskens, 1976). A fertility study of 64 employees engaged in the production of chlorinated 3-carbon compounds, including 1,3-dichloropropene, revealed no effects upon hormone levels (LH, FSH, testosterone), sperm count, sperm motility, and % normal and ab-
normal sperm regardless of duration or magnitude of exposure (Venable et aI., 1980). Several studies have also been undertaken to evaluate potential biological effects of 1,3-dichloropropene vapors in fumigation workers. In a California study, urinary parameters were measured over time in the same workers following single or repeated occupational exposure(s) to a range of 0.39.4 mg/m 3 mixed isomers of 1,3-dichloropropene during soil fumigation operations (Osterloh and Feldman, 1993). Exposure was reportedly associated with elevated urinary excretion of N-acetylglucosamidase (NAG) and retinol binding protein (RBP). No changes were observed in albumin (ALB) excretion. These findings were interpreted to suggest a "subclinical" renal toxicity. In a Dutch study, several urinary and serum parameters were measured in the same workers occupation-
85.6 Summary Risk Characterization
ally exposed to 1.9-18.9 mgim3 cis-l,3-dichloropropene products once before and once after the tulip bulb field fumigation season (about three months duration) (Brouwer et al., 1991). A number of slight changes in several parameters were reported after relative to before the "season:" excretion of urinary NAG and ALB, decreased serum creatinine (CREAT-S), and decreased total bilirubin (TBILI) levels in combination with increased serum y-glutamyltranspeptidase (GGT). No differences were reported in serum (,B2-microglobulin (fhM), alanine aminopeptidase (AAP), ,B-galactosidase, alkaline phosphatase (ALP), aspartate aminotransferase (AST), alanine aminotransferase (ALT), or lactate dehydrogenase. These data were interpreted by Brouwer et al. (1991) to reflect a slight degree of liver and kidney toxicity. Both of these studies have been strongly criticized for perceived study design and data interpretation flaws (Stott et al., 1990; van Sittert et al., 1991). A subsequent comprehensive study in Dutch potato field fumigation workers conducted over the whole of the fumigation season failed to observe treatment-related toxicity (Verplanke et al., 1995). Workers were exposed to a range of 0.1-9.5 mgim3 cis-1,3-dichlororporpene and a number of parameters were measured before, during, and following the fumigation season. Unlike previous studies, this study employed a matched control group of individuals. Parameters measured included urinary AAP, NAG, RBP, and ALB, and serum ,B2M, CREAT-S, ALT, AST, GGT, ALP, and TBILL The only change observed was a slightly lower urinary ratio of 6-,B-hydroxycortisol to free cortisol ratio which was not considered to be related to 1,3dichloropropene exposure. It was concluded that no adverse effects on liver or kidney function were suggested by the data. Additional reviews of the human toxicity data for 1,3dichloropropene have been published by Yang (1986) and IPCS (1993).
85.6 SUMMARY RISK CHARACTERIZATION 1,3-Dichloropropene has found use for over 45 years and remains one of the few remaining compounds available to agriculture for fumigating soils to eliminate parasitic nematodes. This compound has been extensively evaluated in a number of test organisms for acute, subchronic, and chronic toxicity, reproductive and developmental toxicity, carcinogenicity, and genotoxicity. Its metabolism in animals, including humans, has also been extensively studied and is relatively well understood. 1,3-Dichloropropene is moderately to highly acutely toxic to animals. It is irritating to skin and if occluded can cause a chemical bum and death in rabbits; however, its relatively high vapor pressure results in much lower toxicity if left on skin unoccluded. Orally administered 1,3-dichloropropene, either neat or in an aqueous vehicle, may be lethal to rodents at roughly 100 mg/kg or greater. Two human fatalities from accidental imbibition of 1,3-dichloropropene formulations, a relatively purified 1,3-dichloropropene and an admixture with
1857
1,2-dichloropropane, have been reported. Toxic effects observed in humans exposed to high levels of vapors or having extended dermal contact with the liquid appear to reflect the irritant nature of 1,3-dichloropropene. In animal models, the results of subchronic and chronic toxicity studies of this chemical reflect both its irritant properties, as evidenced by effects on portal-of-entry tissues, and potential toxicity of a metabolite(s), as evidenced by effects upon distal tissues (e.g., urinary bladder mucosa). Despite this, studies have demonstrated a lack of reproductive or developmental effects, even at toxic dosages. 1,3-Dich1oropropene is rapidly and extensively metabolized upon absorption by animals, including humans. Elimination from the blood of rats occurs in a biphasic manner, with half-lives for both isomers in the prominent a-phase of approximately 4-7 minutes and the ,B-phase of approximately 25-45 minutes. No appreciable excretion of parent chemical occurs and metabolites are primarily eliminated via the urine as products of a glutathione conjugation metabolic pathway or via exhalation of C02, product of a hydrolytic pathway. Evidence of the former pathway has been dose-related decreases in tissue glutathione levels of rats and mice administered 1,3dichloropropene via oral or inhalation routes. The major urinary metabolite, the mercapturate conjugate of 1,3-dichloropropene, has represented a useful biomarker by which to estimate the exposure of workers to this molecule during soil fumigation operations. Genotoxicity tests of 1,3-dichloropropene have often provided contradictory results. Many short-term assays of genotoxicity have been confounded by the presence of a known mutagen, epichlorohydrin, in the formulated material tested which was historically added as a stabilizing agent. The potential of 1,3-dichloropropene to undergo autooxidation to generate a mutagenic epoxide has further complicated interpretation of the in vitro genotoxicity data. Epoxide may be formed upon prolonged exposure to oxygen or during gas-chromatographic "purification" proceedures carried out prior to testing. In vivo assays of mutagenic or clastogenic activity, with their intact compliment of metabolizing enzymes, have almost uniformly been negative, especially at nontoxic dosages or at dosages which do not deplete tissue glutathione levels. It has been proposed that the genotoxic potential of 1,3-dichloropropene is directly related to the extensive depletion of glutathione in target organisms and tissues (IPCS, 1993). Based upon a weightof-the-evidence analysis inclusive of in vivo assay data, 1,3dichloropropene lacks significant genotoxic activity. Bioassay data have provided an equally complicated assessment of the potential of 1,3-dichloropropene to cause tumors in animals. Inhalation of vapor, the primary route of occupational exposure to this chemical, has been shown to cause an increase in the incidence of benign lung tumors in male mice. Ingestion of this chemical via the diet as a stabilized microencapsulated product has been shown to cause a low incidence of benign liver tumors in rats. These results contrast with those of a previous oral oncogenicity study conducted by repeated bolus dosing (gavage) of an older, epichlorohydrin stabilized, and highly mutagenic formulation of 1,3-dichloropropene which re-
1858
CHAPTER 85
1,3-Dichloropropene
suIted in numerous benign and maligant tumors in both rats and mice.
REFERENCES Bartels, M. J., Waechter, J. M., Kat!, P. E., Dietz, E K., and Hansen, S. C. (1999). Phannacokinetics and metabolism of 1,3-dichloropropene in the rat and mouse. Unpublished. Breslin, W. J., Kirk, H. D., Streeter, C. M., Quast, J. E, and Szabo, J. R. (1989). 1,3-Dichloropropene: Two-generation inhalation reproduction study in Fischer 344 rats. Fund. Appl. Pharmacol. 12, 129-143. Brouwer, E. J., Evelo, C. TA., Verplanke, A. J. w., van Welie, R. T H., and de Wolff, E A. (1991). Biological effect monitoring of occupational exposure to 1,3-dichloropropene: Effects on liver and renal function and on glutathione conjugation. Brit. 1. Ind. Med. 48,167-172. Bruckner, J. v., MacKenzie, W. E, Ramanathan, R., Muralidhara, S., Kim, H. J., and Dallas, C. E. (1989). Oral toxicity of 1,2-dichloropropane: Acute, shortterm and long-term studies in rats. Fund. Appl. Toxicol. 12,713-730. Climie, I., Hutson, D., Morrison, B., and Stoydin, G. (1979). Glutahione conjugation in the detoxication of (Z)-1,3-dichloropene (a component of the nematocide D-D) in the rat. Xenobiotica 9, 149-156. Coate, W. B. (1979a). "Subacute Inhalation Study in Rats and Mice: TELONE H." Report of The Dow Chemical Company. Coate, W. B. (I 979b). "90-Day Inhalation Study in Rats and Mice: TELONE H." Report of The Dow Chemical Company. Creedy, C., Brooks, T, Dean, B., Hutson, D., and Wright, A. (1984). The protective action of glutathione on the microbial mutagenicity of the Z- and E-isomers of 1,3-dichloropropene. Chem.-Biol. Interact. 50, 39-48. Dean, B. J., Brooks, T M., Hodson-Walker, G., and Hutson, D. H. (1985). Genetic toxicology testing of 41 industrial chemicals. Mut. Res. 153, 57-77. De Lorenzo, E, Degl'Innocenti, S., Ruocco, A., Silengo, L., and Cotese, R. (1977). Mutagenicity of pesticides containing 1,3-dichloropropene. Cancer Res. 37, 1915-1917. Dietz, E, Hermann, E., and Ramsey, J. (l984a). The phannacokinetics of 14C_ 1,3-dichloropropene in rats and mice following oral administration. Toxicologist 4, Abst. No. 585. Dietz, E, Dittenber, D., Kirk, H., and Ramsey, J. (l984b). Non-protein sulfhydryl content and macromolecular binding in rats and mice following administration of 1,3-dichloropropene. Toxicologist 4, Abs!. No. 586. Eder, E., Neudecker, T., Lutz, D., and Henschler, D. (1982). Correlation of alkylating and mutagenic activities of allyl and allylic compounds: standard alkylation test vs. kinetic investigation. Chem.-Biol. Interact. 38, 303-315. Fisher, G. D., and Kilgore, W. W. (1988a). Mercapturic acid excretion by rats following inhalation exposure to 1,3-dichloropropene. Fund. Appl. Toxicol. 11,300--307. Fisher, G. D., and Kilgore, W. W. (l988b). Tissue levels of glutathione following acute inhalation of 1,3-dichloropropene. 1. Toxicol. Environ. Hlth. 23, 171-182. Flessel, P., Goldsmith, J., Kahn, E., Wesolwski, J., Maddy, K., and Peoples, S. (1978). Acute and possible long-term effects of 1,3-dichloropropeneCalifornia. Morbidity Motality Weekly Rep. 27, 50--55. Gardner, J. R. (1989). "Cis-l,3-Dichloropropene: Acute Oral and Dermal Toxicity, Skin and Eye Irritancy and Skin Sensitisation Potential." Report of Sittingbourne Research Center. Ghia, M., Robbiano, L., Allavena, A., Martelli, A., and Brambilla, G. (1993). Genotoxic activity of 1,3-dichloropropene in a battery of in vivo short-term tests. Toxicol. Appl. Pharmacol. 120, 120--125. Gollapudi, B. B., Cieszlak, E S., Day, S. J., and Camey, E. W. (1998). Dominant lethal test with rats exposed to 1,3-dichloropropene. Environ. Mol. Mut. 32, 351-359. Gollapudi, B. B., Mendrala, A. M., Linscombe, A., and Stott, W. T (1999). Lack of genotoxicity of 1,3-dichloropropene in in vitro and in vivo genotoxicity tests. Unpublished.
Gosselin, R., Hodge, H., Smith, R., and Gleason, M. (1976). "Clinical Toxicology of Commercial Products," 4th ed., pp. 119-121. Wilkins-Williams, Baltimore. Hanley, T R., John-Greene, Young, J. T, Calhoun, L. L., and Rao, K. S. (1987). Evaluation of the effects of inhalation exposure to 1,3-dichloropropene on fetal development in rats and rabbits. Fund. Appl. Toxicol. 8, 562-570. Haut, K. T, Stebbins, K. E., Kropscott, B. E., and Stott, W. T (I 992a). "TELONE@H Soil Fumigant: Palatability and Two-Week Dietary Probe Studies in Fischer 344 Rats." Report of The Dow Chemical Company, Midland,MI. Haut, K. T, Stebbins, K. E., Kropscott, B. E., and Stott, W. T (1992b). "TELONE@H Soil Fumigant: Palatability and Two-Week Dietary Probe Studies in B6C3FI Mice." Report of The Dow Chemical Company, Midland, MI. Haut, K. T., Johnson, K. A., Shabrang, S. N., and Stott, W. T (1996). Subchronic toxicity of ingested 1,3-dichloropropene in rats and mice. Fund. Appl. Toxicol. 32, 224-232. Hernandez, A. E, Martin-Rubi, J. c., Ballesteros, J. L., Oliver, M., Pia, A., and Villanueva, E. (1994). Clinical and pathological findings in fatal 1,3dichloropropene intoxication. Hum. Exptl. Toxicol. 13, 303-306. Hine, C. H., Anderson, H. H., Moon, H. D., Kodama, J. K., Morse, M., and Jacobson, N. W. (1953). Toxicology and safe handling of CBP-SS (technical l-chloro-3-bromopeopene-I). Arch. Ind. Hyg. Occ. Med. 7, 118-136. Hutson, D. H., Moss, J. A., and Pickering, B. A. (1971). Components of the soil fumigant D-D* and their metabolites in the rat. Fd. Cosmet. Toxicol.9, 677-680. IPCS (International Programme on Chemical Safety) (1993). "1,3Dichloropropene, 1,2-Dichloropropane and Mixtures." Environmental Health Criteria No. 146, World Health Organization, Geneva. Jones, J. R. (1988a). "1,3-Dichloropropene cis-Isomer: Modified NineInduction Buehler Contact Sensitisation Study in the Guinea Pig." Report of Safephann Laboratories Limited, Derby, UK. Jones, 1. R. (l988b). "1,3-Dichloropropene eis-Isomer: Acute Oral Toxicity Test in the Rat." Report of Safephann Laboratories Limited, Derby, UK. Jones, J. R. (l988c). "1,3-Dichloropropene cis-Isomer: Acute Dermal Toxicity Test in the Rat." Report of Safephann Laboratories Limited, Derby, UK. Jones, J. R., and Collier, T A. (1986a). "TELONE H: Acute Oral Toxicity in the Rat." Report of The Dow Chemical Company. Jones, J. R., and Collier, T A. (1986b). "TELONE H: Acute Dermal Toxicity Test in the Rat." Report of The Dow Chemical Company. Kevekordes, S., Gebel, T, Pav, K., Edenharder, R., and Dunkelberg, H. (1996). Genotoxicity of selected pesticides in the mouse bone marrow micronucleus test and in the sister-chromatid exchange test with human Iymphocytes in vitro. Toxicol. Letl. 89, 35-42. Kezic, S., Monster, A. C., Verplanke, J. w., and de Wolff, E A. (1996). Dermal absorption of cis-l ,3-dichloropropene vapour: human experimental exposure. Hum. Exp. Toxicol. 15,396--399. Kitchin, K. T, and Brown, J. L. (1994). Dose-response relationship for rat liver DNA damage caused by 49 rodent carcinogens. Toxicology 88, 31-49. Linnett, S. L., Clark, D. G., Blair, D., and Cassidy, S. L. (1988). Effects of subchronic inhalation of D-D (I ,3-dichloropropene/l ,2-dichloropropane) on reproduction in male and female rats. Fund. Appl. Toxicol. 10,214-223. Lomax, L. G., Stott, W. T., Johnson, K. A., Calhoun, L. L., Yano, B. L., and QuasI, J. E (1988). The chronic toxicity and oncogenicity of inhaled technical grade 1,3-dichloropropene in rats and mice. Fund. Appl. Toxicol. 12, 418-431. Loveday, K. S., Lugo, M. H., Resnick, M. A., Anderson, B. E., and Zeiger, E. (1989). Chromosome aberration and sister chromatid exchange tests in Chinese hamster ovary cells in vitro: H. Results with 20 chemicals. Environ. Mol. Mutag. 13,60--94. Martelli, A., Allavena, A., Ghia, M., Robbiano, L., and Brambilla, G. (1993). Cytotoxic and genotoxic activity of 1,3-dichlororporpene in cultured mammalian cells. Toxicol. Appl. Pharmacol. 120, 114-119. Markovitz, A., and Crosby, W. H. (1984). Chemical carcinogenesis. A soil fumigant, 1,3-dichloropropene, as possible cause of hematologic malignancies. Arch. Intern. Med. 144, 1409-1411.
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Myhr, B. C., and Caspary, W. J. (1991). Chemical mutagenesis at the thymidine kinase locus in L5178Y mouse lymphoma cells: Results for 31 coded compounds in the National Toxicology Program. Environ. Mol. Mutag. 18, 51-83. Nater, J. P., and Gooskens, V. H. J. (1976). Occupational dermatosis due to a soil fumigant. Contact Dermat. 2, 227-229. National Toxicology Program (NTP) (1985). "Toxicology and Carcinogenesis Studies of TELONE H® in F3441N Rats and B6C3Fl Mice (Gavage Studies)." NTP Tech. Rep. 269, Government Printing Office, Washington, DC. Neudecker, T., Stefani, A., and Henschler, D. (1977). In vitro mutagenicity of soil nematocide 1,3-dichloropropene. Experientia 33, 1084-1085. Nitschke, K. D., and Lomax, L. G. (1990). "Cis-l,3-Dichloropropene: 2-Week Vapor Inhalation Toxicity Study in Fischer 344 Rats." Report of The Dow Chemical Company, Midland, MI. Nitschke, K. D., Crissman, J. w., and Schuetz, D. J. (1990a). "Cis-l,3Dichloropropene: Acute Inhalation Toxicity Study with Fischer 344 Rats." Report of The Dow Chemical Company, Midland, MI. Nitschke, K. D., Lomax, L. G., and Sanderson, T. G. (1990b). "Cis-I,3Dichloropropene: 13-Week Vapor Inhalation Toxicity Study in Fischer 344 Rats." Report of The Dow Chemical Company, Midland, MI. Onkenhout, w., Mulder, P. P. J., Boogaard, P. J., Buijs, w., and Vermeulen, N. P. E. (1986). Identification and quantitative determination of mercapturic acids formed from Z- and E-l,3-dichloropropene by the rat, using gas chromatography with three different detection techniques. Arch. Toxicol. 59, 235-241. Osterloh, J. D., and Feldman, B. J. (1993). Urinary protein markers in pesticide applicators during a chlorinated hydrocarbon exposure. Environ. Res. 63, 171-181. Osterloh, J., Letz, G., Pond, S., and Becker, C. (1983). An assessment of the potential testicular toxicity of 10 pesticides using the mouse-sperm morphology assay. Mut. Res. 116,407-415. Osterloh, J. D., Cohen, B. S., Popendorf, w., and Pond, S. M. (1984). Urinary excretion of the N -acetyl cysteine conjugate of cis-l,3-dichloropropene by exposed individuals. Arch. Environ. Hlth. 39, 271-275. Parker, c., Coate, w., and Voelker, R. (1982). Subchronic inhalation toxicity of 1,3-dichloropropene/l,2-dichloropropane (D-D®) in mice and rats. 1. Toxicol. Environ. Hlth. 9, 899-910. Rick, D. L., and McCarty, L. P. (1987). "The Determination of the Odor Threshold of Vapors and Gases." Report of The Dow Chemical Company. Schiffman, D., Eder, E., Neudecker, T., and Henschler, D. (1983). Induction of unscheduled DNA synthesis in HeLa cells by allylic compounds. Can. Let. 20,263-269. Schneider, M., Quistad, G. B., and Casida, J. E. (1998). 1,3-Dichloropropene epoxides: Intermediates in bioactivation of the promutagen 1,3dichloropropene. Chem. Res. Toxicol. 11, 1137-1144. Shelby, M. D., Erexson, G. L., Hook, G. J., and Tice, R. R. (1993). Evaluation of a three-exposure mouse bone marrow micronucleus protocol: Results with 49 chemicals. Environ. Mol. Mut. 21, 160--179. Shirasu, Y., Moriya, M., Tequka, H., Teramoto, S., Ohata, T., and Inoue, T. (1981). Mutagenicity screening studies on pesticides. In "Environmental Mutagens and Carcinogens: Proceedings of the Third International Conference on Environmental Mutagens," Tokyo, Mishima, and Kyoto, September 21-27,pp.331-335. Stebbins, K. E., Stott, W. T., Haut, K. T., Quast, J. E, and Shabrang, S. N. (1999). Subchronic and chronic toxicity of ingested 1,3-dichloropropene in beagle dogs. Unpublished. Stolzenberg, S., and Hine, C. (1980). Mutagenicity of 2- and 3-carbon halogenated compounds in the Salmonella/mammalian-microsome test. Environ. Mut. 2, 59-66. Stott, W. T., and KastI, P. L. (1986). Inhalation pharmacokinetics of technical grade 1,3-dichloropropene in rats. Toxicol. Appl. Pharmacal. 85,332-341. Stott, w., Young, J., Calhoun, L., and Battjes, J. (1988). Subchronic toxicity of inhaled technical grade 1,3-dichloropropene in rats and mice. Fundam. Appl. Toxicol. 11,207-220. Stott, W. T., Waechter, J. M., and Quast, J. T. (1990). Letter to the editor. Arch. Environ. Hlth. 45, 250--253.
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Stott, W. T., Mendrala, A. M., Redmond, J. M., Nwosu, A. E, and Lomax, L. G. (1992). Mechanism of 1,3-dichloropropene (1,3-D) induced toxicity in urinary bladder epithelium of mice. Toxicologist 12, Abstr. No. 415. Stott, W. T., Johnson, K. A., Stebbins, K. E, Redmond, J. M., and Jeffries, T. K. (1996). Dietary chronic toxicity/oncogenicity study of microencapsulated 1,3-dichloropropene (1,3-D) in rats and mice. Toxicologist 30, Abstr. No. 276. Stott, W. T., Gilbert, J. R., McGuirk, R. J., Brzak, K. A., Alexander, L. M., Dryzga, M. D., Mendrala, A. L., and Bartels, M. J. (1998). Bioavailability and pharmacokinetics of microencapsulated 1,3-dichloropropene in rats. Toxicol. Sci. 41, 21-28. Sudo, S., Kimura, Y., Yamamoto, K., and Ichihara, S. (1979). "The Mutagenicity Test on 1,3-Dichloropropene in Bacteria Test System." Report of The Dow Chemical Company, Midland, MI. Talcott, R., and King, J. (1984). Mutagenic impurities in 1,3-dichloropropene preparations. 1. Natl. Cancer Inst. 72, 1113-1116. Til, H. P.,Spanjers, M. T., Feron, V. J., and Reuzel, P. J. C. (1973). "Sub-chronic (90-Day) Toxicity Study with TELONE* in Albino Rats." Report of The Dow Chemical Company, Horgen, Switzerland. Torkelson, T. R. (1994). Halogenated aliphatic hydrocarbons containing chlorine, bromine, and iodine. In "Patty's Industrial Hygeine and Toxicology" (G. D. Clayton and E E. Clayton, eds.), 4th ed., pp. 4007-4251. Wiley, New York. Torkelson, R., and Oyen, E (1977). The toxicity of 1,3-dichloropropene as determined by repeated exposure of laboratory animals. Am. Ind. Hyg. Assoc. 1. 38, 217-223. Valencia, R., Mason, J. M., Woodruff, R. c., and Zimmering, S. (1985). Chemical mutagenesis testing in Drosophila. HI. Results of 48 coded compounds tested for the national toxicology program. Environ. Mut. 7, 325-348. Van Duuren, B. L., Goldschmidt, B. M., Loewengart, G., Smith, A. C., Melchionne, S., Seidman, I., and Roth, D. (1979). Carcinogenicity of halogenated olefinic and aliphatic hydrocarbons in mice. 1. Natl. Cancer Inst. 63, 1433-1439. van Sittert, N. J., Veenstra, G. E., Dumas, E. P., and Tordoir, E. E (1991). Letter to the editor. Br. 1. Med. 48, 646-648. van Welie, R. T. H., van Duyn, P., Brouwer, D. H., van Hemmen, J. J., Brouwer, E. J., and Vermeulen, N. P. E. (1991). Inhalation exposure to 1,3-dichloropropene in the Dutch flower-bulb culture. Part H. Biological monitoring by measurement of urinary excretion of two mercapturic acid metabolites. Arch. Environ. Contam. Toxicol. 20, 6-12. Venable, J. R., McClimans, C. D., Flake, R. E., and Dimick, D. B. (1980). A fertility study of male employees engaged in the manufacture of glycerine.l. Occ. Med. 22, 87-91. Verplanke, A. J. W., Bloemen, L. J., Brouwer, E. J., Van Sittert, N. J., Boogaard, P. J., Herber, R. E M., and De Wolff, F. A. (1998). Monitoring of occupational exposure to cis-l,3-dichloropropene and effects on liver and kidney. Part 2. Effects on liver and kidney. Unpublished. von der Hude, W., Scheutwinkel, M., Gramlich, U., Fibler, B., and Basler, A. (1987). Genotoxicity of three-carbon compounds evaluated in the SCE test in vitro. Environ. Mutag. 9,401-410. Watson, W. P., Brooks, T. M., Huckle, K. R., Hutson, D. H., Land, K. L., Smith, R. J., and Wright, A. S. (1987). Microbial mutagenicity studies with (Z)-1,3-dichloropropene. Chem.-Biol. Interact. 61, 17-30. Waechter, J. M., Brzak, K. A., McCarty, L. P., LaPack, M. A., and Brownson, P. J. (1992). Cisltrans 1,3-dichloropropene (1,3-dichloropropene): Inhalation pharmacokinetics and metabolism in human volunteers. Toxicologist 13, Abstr. No. 1090. Yang, R. S. H. (1986). 1,3-dichloropropene. Residue Rev. 97,19-35. Yang, R. S. H., Huff, J. E., Boorman, G. A., Haseman, J. K., Kornreich, M., and Stookey, J. L. (1986). Chronic toxicology and carcinogenesis studies of TELONE H by gavage in Fischer 344 rats and B6C3Fl mice. 1. Toxicol. Environ. Hlth. 18,377-392.
CHAPTER
86 Phosphine V. F. Garry and A. V. Lyubimov University of Minnesota
86.1 IDENTITY, PROPERTIES, AND USES
are hypophosphorous and phosphoric acids (Van Wazer, 1958; WHO, 1988).
Chemical Name: Hydrogen Phosphide Structure:
86.1.2 CHEMISTRY
PH3.
Synonyms: Phosphoretted hydrogen, phosphorus hydride, Phosphorus trihydride. The CAS Registry No.: 7803-51-2.
Conversion factor: 1 ppm = 1.39 mg/m3. The most common commercial fumigants generating phosphine are alurninum phosphide and magnesium phosphide. Aluminum Phosphide (AlP) is sold under the following trade names: Phostoxin, Furnitoxin, Agtoxin, Weevilcide, Detia, Gastoxin, MaxKill, Phosfume, Fastphos. Common trade names for Magnesium Phosphide (Mg3P2) are Fumi-Cel, Fumi Strip, Magtoxin, Magnaphos, Magphos.
Phosphine is a nucleophile and acts as a strong reducing agent (Lam et aI., 1991). Under standard conditions of temperature, pressure, and humidity PH3 is stable and does not undergo autoxidation. Very early work suggests that under conditions of increased atmospheric pressure and oxygen content autoxidation can occur (Van Wazer, 1958). Further, in the presence of trace levels of diphosphine and perhaps other higher phosphines in air, PH3 will undergo a branched chain oxidation reaction (Green et aI., 1984; Osadchenko and Tomilov, 1969), a form of autooxidation. Similarly, under experimental conditions the reaction can be induced photolytically by ultraviolet (UV) light or ammonia (Buchanan and Hanrahan, 1970; Woller, 1965). The branched chain reaction when it occurs is a generator and a good source of free radicals (see below) (Green et aI., 1984): Propagation
86.1.1 PHYSICAL PROPERTIES Pure phosphine is an odorless and colorless gas with a molecular weight of 34.00 and density of 1.17 at 25°C. Commercial grade phosphine derived from aluminum or magnesium phosphide can contain to a variable degree higher molecular weight phosphines including diphosphines. These higher phosphines give commercial grade fumigants containing aluminum or magnesium phosphide odor characteristics described as decaying fish or "garlic-like." Commercial grade phosphine containing diphosphines can ignite and form explosive mixtures at concentrations exceeding 1.8% phosphine in air. The rate of conversion of the phosphide to phosphine is temperature and humidity dependent. Similarly, metal phosphides readily hydrolyze in water to yield phosphine, which is poorly soluble in water. Major products resulting from the oxidation of phosphine in water Handbook of Pesticide Toxicology Volume 2. Agents
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02 + PH2'
=}
HPO+OH·
OH.+ PH 3
=}
PH2·+H20
02 + PH2'
=}
PH + H02'
02+ PH
=}
HPO+O
Branching 0+ PH3
=}
Termination 0+02+ M
=}
03 +M
Radical + Wall
=}
Compound
Secondary Reactions HPO + 02
=}
HP03
Copyright © 2001 by Academic Press. All rights of reproduction in any form reserved.
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CHAPTER 86
Phosphine
86.2 SOURCES, USES, AND FORMULATIONS
86.4 TOXICITY AND MODE OF ACTION 86.4.1 ACUTE TOXICITY
86.2.1 NATURAL SOURCES 86.4.1.1 Symptoms
Phosphine can be generated in decaying organic matter in open air sewage treatment plants (D€vai et aI., 1998) and other sources of decaying organic material (Glindemann et aI., 1996) including landfills, compost processing, and river sediments. Maximum concentrations detected were approximately 20 ppb. 86.2.2 COMMERCIAL SOURCES
Metal phosphides, notably aluminum, magnesium, and zinc phosphide, are the most common commercial sources of phosphine. Aluminum and magnesium phosphides are commonly used fumigants supplied as pellets, tablets, sachets, ropes, or strips (Meister, 1999) for insect control in stored grains and other products. Zinc phosphide baits are commonly used for rodent control. Ammonia from ammonium carbamate is sometimes used as a warning odorant in some fumigant formulations. Phosphine gas is also used in the synthesis of flame retardants, as a dopant in the semiconductor industry, and as a polymerization initiator and catalyst (U.S. DHHS, 1993).
86.3 TOXICOLOGY 86.3.1 OVERVIEW
The modern history of our understanding of the biologic effects of the toxicant phosphine begins with the works of O. R. Klimmer (1969, 1970). In these works the investigator established the dose related lethal effects of phosphine in multiple species, determined the dose threshold for lethality, and explored possible mechanisms for lethality including effects on hemoglobin. From these works, there is ample evidence that the acute lethal effects of phosphine can occur at levels less than 8 mg/m 3 . Since that time, work by others has gone forward to explore the avenues for the lethal effects of this toxic ant gas in insects, mammals, and humans in vivo and in vitro. Genotoxicity and reproductive effects have also been considered. Because phosphine is an explosive hazard, many of the laboratory-based studies have been conducted under exposure conditions to eliminate or reduce the possibility of the branched chain oxidation reaction in air. Thus, these studies reflect the effects of phosphine in the unoxidized state. Human case and field population studies and some in vitro studies may reflect to a greater or lesser degree the toxicant effects of phosphine and its autooxidation products induced by the contaminant diphosphine in the commercial product, and uncontrolled environmental conditions including UV light, humidity, temperature, and/or ammonia as well. The studies reviewed below emerge as a complex picture of the toxicant effects of phosphine.
Early on, Klimmer found that animals exposed to high concentrations of phosphine quickly develop lassitude, ataxia, apnea, and cardiovascular collapse resulting in death within one halfhour (Klimmer, 1969). At lower concentrations (range studied 7.5 to 564 mg/m 3 ) time to death varied with dose (Fig. 86.1). Concentrations as low as 7 mg/m3 are lethal over a period of 820 hours. In humans, case studies involving suicide and suicide attempts by ingestion of pellets of aluminum phosphide are instructive. Rapid onset of epigastric distress, hypotension, cardiovascular collapse, and death are a recurrent pattern. In those who reach a hospital, altered sensoria, vomiting, severe acidosis, hypotension, cardiac arrhythmia, jaundice, and pulmonary crepitation were common occurrences (Banjaj and Wasir, 1988; Misra et aI., 1988a, b; Singh et aI., 1996). In a review of 195 intentional intoxication cases, Singh et al. (1985) concluded that ingestion of 1.5 g aluminum phosphide can be lethal in adults. Autopsy findings from published accidental death investigations (Garry et aI., 1993; Heyndrickx et aI., 1976; Wilson et aI., 1980) show microscopic pulmonary congestion with edema and alveolar cell necrosis, individual myocardial cell and liver cell necrosis, and anoxic changes in the brain. Klimmer (1970) noted earlier in autopsied animals a peculiar crimson color to the blood. These findings were variably recorded in the human autopsy and in clinical case studies. These human case studies and early animal acute toxicity studies provide some insights for formal mechanistic studies.
1000 SOO
,.....
...
i
8
200
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'§
! 2 I 0.1
10
100
1000
Time (hrs) Figure 86.1 Comparison of dose and time to death in different species. Response of rats (0), rabbit (6), guinea pigs (e) and cats (D) to phosphine. Each point indicates the concentration of phosphine to which a group of animals was exposed and the average time to death. From data of Klimmer (1969) and Hayes and Laws (1991). Reproduced with permission.
86.8 Peroxidases, Lipid Peroxidation, Catalase, and Cholinesterase
86.5 ANIMAL DOSEIRESPONSE 86.5.1 THRESHOLD FOR LETHALITY The early studies of Klimmer (1969) as illustrated previously show that for rats, rabbits, cats, and guinea pigs, a threshold for acute lethality by the inhalation route occurs at about 7 mg/m 3 . Similarly, Newton et al. (1993), demonstrated in pregnant Fischer 344 female rats, exposed six hours daily, four days was the median lethal time at a concentration of 9.7 mg/m 3 . Concentrations below 7 ppm showed no lethality. In mice (both sexes) the Median Lethal Dose after two weeks exposure is 9 mg/m 3 (Barbosa et aI., 1994). Concentrations below this level were not lethal. As indicated before, there are only minor differences in the mortality data from earlier to more recent studies regarding duration time-dose threshold for acute lethality.
86.5.2 ACUTE AND SUBACUTE DOSEIRESPONSE Given the time-duration effects noted above and other factors, the LCso for inhaled phosphine is somewhat variable. Early studies by Waritz and Brown (1975) showed a four hour LCso of 11 ppm in male rats. Using highly purified phosphine, Omae et a!. (1996) reported a four hour LCso between 26.5 and 33.4 ppm in male mice. Newton et a!. (1993) reported no lethality in male and female rats acutely exposed to 10 ppm phosphine for six hours. In subacute studies in Fischer 344 female rats these authors indicate that three day exposure to 10 ppm phosphine was lethal. They further demonstrated that female rats were more sensitive to the lethal effects of the inhaled gas.
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endpoints including the cytochromes and cytochrome oxidase system, hemoglobin, peroxidases and lipid peroxidation, catalase, cholinesterase, and DNA have been studied in some detail. The reported phosphine effects in each of these systems will be discussed below. As part of the discussion, the importance of oxygen as a modifier enhancing toxicity and the reduction of toxicity in reduced oxygen atmospheres will be considered.
86.7.2 CYTOCHROMES AND CYTOCHROME OXIDASE Early on, the requirement for oxygen to mediate the toxicity of phosphine was identified in insects (Bond et a!., 1969) indicating that the gas may be an aerobic mitochondrial respiratory poison. Since that time, in vitro studies, both animal and insect, have shown that the respiratory enzyme, cytochrome c oxidase, may be the specific site of action (Bolter and Chefurka, 1990; Chaudhry, 1997; Kashi and Chefurka, 1976; Price, 1980). On the other hand, in vivo treatment of insects with lethal dose levels of phosphine (Nakakita, 1987) showed no more than 50% inhibition of the enzyme. Further work showed that this level of respiratory enzyme inhibition was sufficient to generate superoxide anions (Bolter and Chefurka, 1990) and these authors suggested that the toxicity of phosphine was due to free radical damage.
86.7.3 HEMOGLOBIN
Aside from empirical observations regarding ingestion and respiratory exposure, there is little toxicokinetic data regarding absorption, distribution, and excretion of phosphine and its reaction products. In one study, 32p labeled phosphine as reaction product residues (hypophosphite and phosphite) in flour were fed to mice. Labeled material in excreta was found to persist for periods up to three weeks (Robinson and Bond, 1970).
In seminal efforts Trimborn and Klimmer (1962) described phosphine-induced hemoglobin denaturation, oxidation to methemoglobin, and formation of a peculiar pigmented form of hemoglobin "Verdichromogen." Studies of purified hemoglobin by Potter et al. (1991) and Chin et a!. (1992) showed that with increasing duration of exposure, phosphine in concentrations as low as 0.11 f.lM gradually resulted in the formation of hemichrome pigment. In intact red blood cells, Potter et a!. (1991) noted formation of Heinz bodies (hemoglobin protein aggregates) at PH3 concentrations as low as 2 f.lg/ml. The toxicant effects both in intact cells and in purified hemoglobin were abolished by incubation in a reduced oxygen atmosphere, indicating an oxygen requirement for phosphine hemoglobin toxicity.
86.7 CELLULAR AND MOLECULAR STUDIES
86.8 PEROXIDASES, LIPID PEROXIDATION, CATALASE, AND CHOLINESTERASE
86.6 ABSORPTION, DISTRIBUTION, METABOLISM, AND EXCRETION
86.7.1 GENERAL Much of the work regarding phosphine as a metabolic poison centers on the concept that reactivity of phosphine as a nucleophile, and/or the electrophilic character of the intermediates arising from oxidation, could lead to derivatization of critical biomolecules (Lam et a!., 1991). Certain critical biologic
Because phosphine is a strong reducing agent, peroxidation and formation of peroxides and their reduction are concerns mechanistically and therapeutically. Studies by Pazynich et a!. (1984) in animals showed that the gas inhibited myeloperoxidase enzyme at concentrations of 8 mg/m3 . More recently in the occupational setting, Garry et a!. (1990) noted histochemically a 50% reduction in myeloperoxidase activity in neutrophils
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CHAPTER 86
Phosphine
from exposed workers compared to control subjects. Ambient air monitoring data obtained at the time varied from OA to 5.8 mg/m 3. The permissible exposure limit (PEL) for phosphine at the time in the US was OA mg/m 3. In other human studies by Chugh et al. (1996) in 45 patients recovering from phosphine poisoning, serial studies of serum levels of superoxide dismutase (SOD), malondialdehyde (MDA), and catalase were performed. Increased levels of SOD and MDA were found in non survivors while catalase was inhibited. Remarkably similar findings (i.e., decreased peroxidase and catalase and increased superoxide dismutase) were reported by Bolter and Chefurka (1990) and Chaudhry and Price (1990) in insects. Taken together, the works cited above indicate that phosphine intoxication can lead to accumulation of cellular peroxides. Further, the oxidation of phosphine (Lam et aI., 1991) can lead to formation of reactive phosphorylating species. As such effects on cholinesterase are also possible. Significant inhibition of cholinesterase was detected in animals (Pazynich et aI., 1984). Occupational studies of grain fumigant applicators (Potter et aI., 1993) and in vitro studies in human red blood cells (Potter et aI., 1991) demonstrate that significant phosphineinduced inhibition of red cell cholinesterase occurs at concentrations exceeding 10 J.!g/ml.
86.9 GENOTOXICITY, CANCER AND REPRODUCTIVE EFFECTS Studies in the occupational setting (Garry et aI., 1989, 1990, 1992) suggest that in enclosed space applications where PH3 ambient air concentrations exceed the permissible exposure limit of OA mg/m3 (range OA-5.8 mg/m 3) for a duration of more than 20 minutes, increased chromosome aberrations are detectable in human lymphocytes from exposed workers. Studies by Barbosa and Bonin (1994) using micronucleus assay found no increase in micronucleus frequency in exposed workers where ambient exposures were less than the PEL. Later follow up studies by Garry et al. (1996) of the same worker population did not show increased chromosome aberrations. During the interim, changes in application practice from manual probe application to more automated methods and non use of phosphine in pesticide applications were noted (unpublished). In subacute tightly controlled animal studies (Kligerman et aI., 1994b) using purified PH3 mixed with nitrogen, no increased numbers of micronuclei or chromosome aberrations were found in spleen cells cultured from animals exposed to phosphine for six hours per day for 9 days at concentrations as high as 7 mg/m 3 in ambient air. A single 6 hours 20 mg/m 3 study by this investigator showed similar negative results (Kligerman et aI., 1994a). In similarly constructed subchronic studies, Barbosa et al. (1994) found significantly increased numbers of micro nuclei at the highest concentration tested (6.3 )!g/m3). Cast in the light of these in vivo studies (both animal and human) one can conclude that in regard to genotoxicity, phosphine may be a genotoxin. From a mechanistic view, in vitro studies of genotoxicity offer some additional insights. In these studies (Garry
et aI., 1989; Hsu et aI., 1998) aluminum or magnesium phosphide was used as a phosphine generating system. Exposure of human lymphocytes (Garry et aI., 1989) to concentrations of phosphine (1.4-4.5 J.!g/l) derived from AlP for 20 minutes yielded increased chromosome aberrations after 96 hours of lymphocyte culture, indicating that the expression of genotoxicity of phosphine is delayed. In a much more detailed mechanistic examination of the genotoxicity of phosphine derived from AlP or Mg3P2 in Hepa cells at a nominal concentration of 1 mM PH3, Hsu et al. (1998) found that reactive oxygen species were maximally generated between 0.5 to 1.5 hr, while damage to DNA expressed as 8-hydroyguanine adducts occurred between 4 and 6 hours. Both of these in vitro studies demonstrate that phosphine or its reaction products derived from AlP can generate DNA damage and that expression of these effects is delayed, probably indirectly and dependent on generation of hydrogen peroxides (Hsu et aI., 1998). No completed long term animal studies were noted in this review regarding carcinogenicity. One preliminary report (U.S. EPA, 1998) noted no carcinogenic effects in rats chronically exposed to an inhaled dose of 3 ppm phosphine after one year. One human epidemiologic study of grain worker mortality (Alavanja et aI., 1987a, b) shows an excess of cancers of the lymphatic and hematopoietic system in this occupational setting where exposure to phosphine and other chemicals and biologic agents occurs. One study of animal teratogenicity (Newton et aI., 1993) with exposure concentrations as high as 4.9 ppm during days 6-15 of gestation in rats showed neither maternal toxicity nor developmental toxicity. No other reproductive endpoint studies were available for review.
86.10 TREATMENT OF POISONING There is no current medical standard of treatment for acute phosphine intoxication. In general, support of vital functions, prevention and/or treatment of shock, and early gastric lavage for ingested poison are suggested (Singh et aI., 1985). Few clinical research efforts have been devoted to evaluation of antiperoxidants use (Chugh et al., 1996; Gupta and Ahlawat, 1995) such as magnesium sulfate for treatment of acute intoxication. Finally, there is clear need to fully evaluate use of antioxidants as potential therapeutic agents in light of the current toxicologic findings.
86.11 REGULATORY NOTES (EXPOSURE GUIDELINES) NIOSH REL: TWA 0.3 ppm (0.4 mg/m3), STEL 1 ppm (1.4 mg/m3) OHSHA PEL: TWA 0.3 ppm (0.4 mg/m 3) 1993-1994 AGGIH TLV: 0.3 ppm (0.42 mg/m3) TWA, 1 ppm (lA mg/m 3) STEL Revised IDHL (immediately dangerous to health or life): 50 ppm (NIOSH, 1996)
References
Acute reference dose (RID) was established as 0.018 mg/kg/ day (U.S. EPA, 1998). Chronic reference dose was found to be 0.0113 mg/kg/day. Earlier EPA established RID for phosphine was 0.0003 mg/kg/d based on body weight and clinical parameters (U.S. EPA, 1995).
86.12 SUMMARY AND COMMENTS Phosphine is a toxic ant gas with strong reducing properties capable of chemical and biologic oxidant effects. The signature threshold for lethality over a narrow dose range and slow evolution of mortality at lower doses indicates that the chemical induces a cumulative biologic oxidant cascade involving progressive alteration of a number of critical biologic endpoints. The critical threshold for these effects may be moderated by environmental-chemical interactions affecting conditions of exposure. As O. R. Klimmer (1969) said "It is a most peculiar poison."
REFERENCES Alavanja, M. c., Malker, H., and Hayes, R. B. (l987a). Occupational cancer risk associated with the storage and bulk handling of agricultural foodstuff. J. Toxicol. Environ. Health 22, 247-254. Alavanja, M. C., Rush, G. A., Stewart, P., and Blair, A. (l987b). Proportionate mortality study of workers in the grain industry. J. Natl. Cancer I. 78, 247252. Banjaj, R., and Wasir, H.S. (1988). Epidemic caluminium phosphide poisoning in northern India. Lancet I, 820-821. Barbosa, A., and Bonin, A. M. (1994). Evaluation of phosphine genotoxicity at occupational levels of exposure in New South Wales, Australia. Occup. Environ. Med. 51,700-705. Barbosa, A., Rosinova, E., Dempsey, l., and Bonin, A. M. (1994). Determination of genotoxic and other effects in mice following short term repeateddose and subchronic inhalation exposure to phosphine. Environ. Mol. Mutagen. 24,81-88. Bolter, C. l., and Chefurka, W. (1990). Extramitochondrial release of hydrogen peroxide from insect and mouse liver mitochondria using the respiratory inhibitors phosphine, myxothiazol, and antimycin and spectral analysis of inhibited Cytochromes. Arch. Biochem. Biophys. 278, 65-72. Bond, E. J., Robinson, J. R., and Buckland, C. T. (1969). The toxic action of phosphine: Absorption and symptoms of poisoning in insects. J. Stored Prod. Res. 5, 289-298. Buchanan, J. W., and Hanrahan, R. J. (1970). The radiation chemistry of phosphine-ammonia mixtures in the gas phase. Mutat. Res. 44, 206-304. Chaudhry, M. Q. (1997). A review of the mechanisms involved in the action of phosphine as an insecticide and phosphine resistance in stored-product insects. Pestic. Sci. 49,213-228. Chaudhry, M. Q., and Price, N. R. (1990). A spectral study of the biochemical reactions of phosphine with various haemproteins. Pestic. Biochem. Physiol. 36, 14-21. Chin, K. L., Meaklim, M. J., Scollary, G. R., and Leaver, D. D. (1992). The interaction of phosphine with haemoglobin and erythrocytes. Xenobiotica 22,599-607. Chugh, S. N., Arora, v., Sharma, A., and Chugh, K. (1996). Free radical scavengers and lipid peroxidation in acute aluminum phosphide poisoning. Indian J. Med. Res. 104, 190-193. D6vai, I., FelfOldy, L., Wittner I., and Plosz, S. (1998). Detection of phospine: New aspects of the phosphorous cycle in the hydrosphere. Nature 333, 343345.
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Garry, V. F., Danzl, T. J., Nelson, R. L., Cervenka, J., Krueger, L. A., Griffith, J., and Whorton, E. (1989). Human genotoxicity: Pesticide applicators and phospine. Science 246, 251-255. Garry, V. F., Nelson, R. L., Danzl, T. J., Cervenka, J., Krueger, L. A., Griffith, J., and Whorton, E. (1990). Human genotoxicity in phosphine-exposed fumigant applicators. Prog. Clin. BioI. Res. 340C, 367-376. Garry, V. F., Danzl, T. J., Tarone, R., and Griffith, J. (1992). Chromosome rearrangements in fumigant appliers: possible relationship to non-Hodgkin'S lymphoma risk. Cane. Epi. Biomark. Prey. 1, 287-291. Garry, V. F., Good, P. F., Manivel, c., and Perl, D. (1993). Investigation of a fatality from nonoccupational aluminum phosphide exposure: Measurement of aluminum in tissue and body fluids as a marker of exposure. J. Lab. Clin. Med 122,739-747. Garry, V. F., Tarone R. E., Long, L., Griffith, J., Kelly, J. T., and Burroughs, B. (1996). Pesticide appliers with mixed pesticide exposure: G-banded analysis and possible relatonship to non-Hodgkin's lymphoma. Cane. Epi. Biomark. Prey. 5, 11-16. Glindemann, D., Stottmeister, U., and Bergmann, A. (1996). Free phosphine from the anaerobic biosphere. Environ. Sci. Pollut Res. Intern. 3,17-19. Green, A. R., Sheldon, S., and Banks, H. J. (1984). The flammability limit of pure phosphine-air mixtures at atmospheric pressure. In "Controlled Atmosphere and Fumigation in Grain Storage" (B. E. Ripp, ed.), Vol. 5, pp. 433451. Elsevier, Amsterdam. Gupta, S., and Ahlawat, S. K. (1995). Aluminum phosphide poisoning: A review. J. Toxicol. Clin. Toxicol. 33, 19-24. Hayes, W. J., and Laws, E. R., ed. (1991). "Handbook of Pesticide Toxicology," Vol. 2, p. 657. Heyndrickx, A., Van Peteghem, C., Van Den Heede, M., and Lauwaert, R. (1976). A double fatality with children due to fumigated wheat. Eur. J. Toxicol. 9, 113-118. Hsu, C.-H., Quistad, G. B., and Casida, J. E. (1998). Phosphine induced oxidative stress in Hepa lc1c7 cells. Toxicol. Sci. 46, 204-210. Kashi, K. P., and Chefurka, W. (1976). The effect of phosphine on the absorption and circular dichroic spectra of cytochrome c and cytochrome oxidase. Pestic. Biochem. Physiol. 6, 350-362. Kligerman, A. D., Bryant, M. F., Doerr, C. L., Erexson, G. L., Kwanyuen, P., and McGee, J. K. (1994a). Cytogenic effects of phosphine inhalation by rodents: I. Acute 6 hour exposure of mice. Environ. Mol. Mutagen. 23, 186189. Kligerman, A. D., Bishop, J. B., Erexson, G. L., Price, H. c., O'Connor, R. w., Morgan, D. L., and Zeiger. E. (1994b). Cytogenic and germ cell effects of phosphine inhalation by rodents. n. Subacute exposures to rats and mice. Environ. Mol. Mutagen. 24,301-306. Klimmer, O. R. (1969). Beitrag zur Wirkung des Phosphorwasserstoffes (PH 3). Zur Frage der sog chronischen Phosphorwasserstoffvergiftung. Arch. Toxikol. 24, 164-187 [in German]. Klimmer, O. R. (1970). Akute Vergiftungen durch Insektizide und Herbizide. Z. Allgemeinmedizin 46,1731-1734 [in German]. Lam, W. W., Toia, R. F., and Casida, l. E. (1991). Oxidatively initiated phosphorylation reactions of phosphine. J. Agric. Food Chem. 39,2274-2278. Meister, R. T. (1999). "Farm Chemicals Handbook." Meister, Wikkoughby. Misra, U. K., Bhargave, S. K., Nag, D., Kidwai, M. M., and Lal, M. M. (1988a). Occupational phosphine exposure in Indian workers. Toxicol. Lett. 42,257263. Misra, U. K., Tripathi, A. K., Pandey, R., and Bhargwa, B. (l988b). Acute phosphine poisoning following ingestion of aluminum phosphide. Human Toxicol. 7,343-345. Nakakita, H. (1987). The mode of action of phosphine. J. Pestic. Sci. 12, 299309. Newton, P. E., Schroeder, R. E., Sullivan, J. B., Busey, W. M., and Banas, D. A. (1993). Inhalation toxicity of phosphine in the rat: Acute, subchronic, and developmental. Inhal. Toxicol. 5,223-239. NIOSH Pocket Guide to Chemical Hazards (1996). "Documentations for Immediately Dangerous to Life or Health Concentrations (IDLH): Phosphine."
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Omae, K., Ishizuka, c., Nakashima, H., Sakurai, H., Yamazaki, K., Mori, K., Shibata, T., Kanoh, H., Kudo, M., and Tati, M. (1996). Acute and subacute inhalation toxicity of highly purified phosphine (PH3) in male ICR mice. l. Oeeup. Health 38, 36-42. Osadchenko, I. M., and Tomilov, A. P. (1969). Phosphorous hydrides. Russian Chem. Rev. 33,495-504. Pazynich, V. M., Mazur, I. A., Podlozny, A. v., Chinchevich, V. I., and Mandrichenko, B. E. (1984). Experimental substantiation and prediction of time related maximum permissible concentration of phosphine in the air. Gig. Sanit. 1, 13-15 [in Russian]. Potter, W. T., Rong, S., Griffith, J., White, J., and Garry, V. E (1991). Phosphine mediated Heinz Body formation and hemoglobin oxidation in human erythrocytes. Taxiea!. Lelt. 57,37-45. Potter, W. T., Garry, V. E, Kelly, J. T., Taronef, R., Griffith, J., and Nelson, R. L. (1993). Radiometric assay of red cell and plasma cholinesterase in pesticide appliers from Minnesota. Taxiea!. Appl. Phannaea!. 119, 150--155. Price, N. R. (1980). Some aspects of the inhibition of cytochrome e oxidase by phosphine in susceptible and resistant strains of Rhyzopertha Dominicia. Insect Biaehem. 10, 147-150. Robinson, J. R., and Bond, E. J. (1970). The toxic action of phosphine: Studies with 32PH3; terminal residues in biological materials. 1. Stared Prod. Res. 6, 133-146. Singh, S., Dilawari, J. B., Vashist, R., Malhotra, H. S., and Sharma, B. K. (1985). Aluminum phosphide ingestion. Br. Med. 1. (Clin. Res.) 290, 1110-III 1. Singh, S., Singh, D., Wig, N., Jit, I., and Sharma, B. K. (1996). Aluminum phosphide ingestion-A cIinico-pathologic study. 1. Taxieol. Clin. Taxiea!. 34, 703-706.
Trimbom, H., and Klimmer, O. R. (1962). Experimentelle untersuchungen iiber chemische veranderungen des blutfarbstoffs in vitro durch phosphorwasserstoff. Arch. Int. Phannaeadyn. CSSSVII, 331-347. U.S. Department of Health and Human Services (1993). "Hazardous Substances Data Bank" (HSDB, online database). National Toxicilogy Information Program, National Library of medicine, Bethesda. U.S. EPA (1995). "Integrated Risk Information System (IRIS) on Phosphine." Environmental Criteria and Assessment Office, Office of Health and Environmental Assessment, Office of Research and Development, Cincinnati, OH. U.S. EPA (1998). "Prevention, Pesticides And Toxic Substances (7508C). Reregistration Eligibility Decision (RED) Aluminum and Magnesium Phosphide." EPA 738-R-98-017. Van Wazer, J. R. (1958). "Phosphorus and Its Compounds," VoI. I. Chemistry. Interscience Publishers, New York. Waritz, R. S, and Brown, R. M. (1975). Acute and subacute inhalation toxicities of phosphine, phenylphosphine and triphenylphosphine. Am. Ind. Hyg. Assae. l. 36,452-458. Wilson, R., Lovejoy, E R., Jaeger, R. J., and Landrigan, P. L. (1980). Acute phosphine poisoning aboard a grain freighter. lAMA, l. Am. Med. Assae. 244, 148-150. Woller, C. R. (1965). Aliphatic Compounds of Some Elements. In "Chemistry of Organic Compounds" (w. B. Saunders, ed.), pp. 317-323. World Health Organization (1988). Phosphine and selected metal phosphides. Environmental Health Criteria 73,17-19.
CHAPTER
87 Metam-Sodium Linda L. Carlock Toxicology and Regulatory Consulting
Timothy A. Dotson UCB Chemicals Corporation
87.1 INTRODUCTION Metam-sodium (C2H4NNaS2, CAS no. 137-42-8), also known as metham sodium, sodium metam, sodium-N-methyldithiocarbamate, methylcarbamodithioic acid sodium salt, methyldithiocarbamic acid sodium salt, carbam, and SMDC, is a white crystalline powder in the pure form but is normally found as a clear yellow liquid with a strong sulfurlike odor (Merck, 1989). Metam-sodium is prepared from methylamine, carbon disulfide, and sodium hydroxide in an aqueous solution. Metamsodium has a molecular weight of 129.18. Metam-sodium is stable in its dry, crystalline state, and in concentrated aqueous solution. In solution, metam-sodium has a vapor pressure of 21 mg Hg at 25°C (U.S. EPA, 1994a). Metam-sodium is very stable at a pH greater than 8.8, but at pH 7 and below it readily hydrolyzes. In soil or when diluted with water, metam-sodium is converted to methyl isothiocyanate (MITC). Other degradates of metam-sodium include carbon disulfide (CS2) and hydrogen sulfide (H2S). Metam-sodium is an agricultural general use pesticide used primarily as a broad spectrum preplant soil fumigant to control weeds, weed seeds, fungi, nematodes, and soil insects. End use products are formulated as 18-42% aqueous solutions sold under the trade names of Metam CLR, Vapam, and Sectagon. Metam-sodium has been registered since 1954. Registered uses of metam-sodium include agricultural soil fumigation, wood preservative, slimicide, tree-root killer, and aquatic weed control. Approximately 10 million pounds of metam-sodium were used in 1990, with 40-45% used for agricultural purposes (U.S. EPA,1994a). As a soil fumigant, metam-sodium is applied after harvest and/or 14 to 21 days prior to planting by shank injection, disc, rotary tiller, drip irrigation, solid set sprinkler, or center pivot chemigation. In some parts of North America, fall applications are preferred because metam-sodium volatilizes over the winter and clears the soil, allowing planting to begin as soon as favorable springtime conditions arrive. By treating the soil with metam-sodium, fruit and vegetable growers can control Handbook of Pesticide Toxicology Volume 2. Agents
weeds, reduce nematode populations, and control soil-borne pests. Metam-sodium may be used on all crops but is particularly important in the production of melons, peppers, tomatoes, potatoes, strawberries, citrus, grapes, almonds, artichokes, asparagus, carrots, lettuce, spinach, squash, forest tree seedlings, ornamentals, and cut flowers. By reducing competition from soil pests, metam-sodium promotes healthier plants and increased yields. The U.S. EPA (1997) considers metam-sodium to be a commercially viable alternative to methyl bromide fumigation for fruit and vegetable production due to its low cost, wide range of control, and long record of safe use. It can be used to control weeds (e.g., bluegrass, Bermuda grass, chickweed, dandelion, ragweed, henbit, nutsedge, and wild morning glory), nematodes, and soil diseases caused by species of Rhizoctonia, Fusarium, Pythium, Phytophthora, Verticillium, and Sclerotinia (U.S. EPA, 1997). Metam-sodium has also been shown to be useful in integrated pest management systems as it can be used in conjunction with other treatment methods such as biological controls and soil pasteurization. Metam-sodium is a slightly to moderately toxic compound that when used according to label directions has been shown to be a safe and versatile product for over 45 years. For agricultural use, metam-sodium must be applied in a manner where there is no contact with workers or other persons, either directly or through drift. Only handlers equipped with the proper personal protection equipment may be in the area during application. In California, application must also be in compliance with the Technical Information Bulletin "Guidelines for All Application Methods for Metam-sodium in California." The potential routes of human chemical exposures are oral (ingestion), dermal (direct skin contact), and inhalation, however, the chance for nonoccupational exposure to metam-sodium is minimal. Approved agricultural uses of metam-sodium do not leave residues on crops, thus eliminating diet as a source of exposure. The primary means of exposure to metam-sodium is through dermal occupational exposure. Most of the potential for exposure to metam-sodium itself
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CHAPTER 87 Metam-Sodium
comes from transloading and handling the liquid when preparing for application. The use of required protective gloves, boots, and clothing minimizes or eliminates dermal exposure to metam-sodium. The V.S. EPA's Occupational and Residential Exposure Branch assumes that dermal exposure is minimal for handlers and nonexistent for nearby residents and bystanders (V.S. EPA, 1994a). There is little potential for inhalation exposure to metam-sodium, which has been proven through extensive monitoring and in a number of worker exposure studies. Proper protective equipment such as in-cab filtering systems and NIOSH-approved respirators that are used by workers provide protection in the unlikely situation that the liquid compound becomes aerosolized. The toxicology of "technical grade" or formulated metamsodium (approximately 42% a.i.) is well established. Metamsodium is synthesized in aqueous solution and then diluted with water, as necessary, to achieve the desired concentration and meet the label guarantee for the formulated product. Thus, "technical grade" is synonymous with the formulated material. Little toxicity information is available regarding pure or analytical-grade metam-sodium. The following discussion of technical grade/formulated metam-sodium toxicity briefly covers a number of published studies and the results of toxicity studies submitted to governmental agencies in support of metam-sodium registration.
87.2 ACUTE TOXICITY Metam-sodium is slightly to moderately acutely toxic depending on the route of exposure. The following toxicity values pertain to technical grade metam-sodium (V.S. EPA, 1994a). All of the following values were obtained with standard acute toxicity studies designed to determine the dose or concentration that causes death to 50% of the test animals (LD50 or LC50): • The acute LD50 for technical grade metam-sodium (43.7% a.i.) is reported as 870 mg/kg for male rats and 924 mg/kg for female rats. The combined (male and female) LD50 is 896 mg/kg (placing the compound into Toxicity Category III (V.S. EPA, 1994a) or similarly classified as slightly toxic (LD50 = 5-15 g/kg; Klaassen, 1986). • The acute dermal LD50 of technical grade metam-sodium (43.7%) applied to male and female rabbits is 368 mg/kg (Toxicity Category Ill). • The acute inhalation LC50 of aerosolized technical grade metam-sodium (42%) in rats is 2.275 mg/l (Toxicity Category Ill). • Technical grade metam-sodium (42%) was found to be slightly irritating to the eyes of New Zealand White rabbits (Toxicity Category Ill). • Technical grade metam-sodium (42%) is irritating to the shaved skin of male and female rabbits (Liggett and McRae, 1991) and is classified as a moderate to severe dermal irritant (Toxicity Category 11).
• Metam-sodium (42%) was also found to be a skin sensitizer to guinea pigs using the delayed contact hypersensitivity test (Parcell and Denton, 1991). Acute studies conducted with 32.7% metam-sodium showed similar but milder results than the above cited data for the 42% formulated compound. Jowa (1998) reported the following values for multiple studies conducted with metam-sodium: • The acute oral LD50 for 32.7% metam-sodium varied from 1294 to 1415 mg/kg for male rats and 1350 to 1428 mg/kg for female rats. • The acute dermal LD50 for 32.7% metam-sodium varied from 1012 to 3500 mg/kg in rabbits. • The acute inhalation LC50 varied from >4.7 to >5.4 mg/l for male rats exposed to 32.7% metam-sodium for four hours. • In one eye irritation study with rabbits, 32.7% metam-sodium was found to be a mild irritant, but in another study it was found to be nonirritating. • Dermal irritation studies with rabbits exposed to 32.7% metam-sodium showed that the compound was a severe irritant in one study and was corrosive in another study. • Testing guinea pigs with 32.5% metam-sodium in the Buehler test resulted in sensitization. Standardized acute toxicity studies provide limited information regarding subtle toxic effects and are not designed to establish a no observed effect level (NOEL). To further understand the sublethal effects of a compound, lower dose levels or concentrations are required.
87.3 SUB CHRONIC TOXICITY Effects of metam-sodium exposure over longer periods vary with the species tested and route of administration. A variety of toxicity studies have shown that there is a definite doseresponse effect to metam-sodium (V.S. EPA, 1992, 1993). At very low doses levels there is no evidence of toxicity, but as the dose level increases the prevalence and severity of toxic effects increases. In a 90-day study (Whiles, 1991), male and female mice were administered metam-sodium in drinking water at dose levels of 0,0.018,0.088,0.35, or 0.62 mg/ml (2.7, 11.7,52.4, or 78.7 mg/kg/day for males; 3.6, 15.2, 55.4, or 83.8 mg/kg/day for females). No treatment-related mortality, morbundity, or clinical signs of toxicity were observed during the 90-day study period. Treatment-related statistically significant decreases in mean body weight were observed in both males and females at dose levels of 0.35 and 0.62 mg/ml. Treatment-related changes in hematology parameters were noted at doses as low as 0.088 mg/ml for females and 0.62 mg/ml for males. The lowest effect level was determined to be 0.088 mg/ml (11.7 mg/kg/day for males, 15.2 mg/kg/day for females) based on urinary bladder lesions observed in both males and females and in statisti-
87.4 Genetic Toxicity
cally significant decreases in hemoglobin, red blood cell, and hematocrit in females. The NOEL for systemic toxicity was 0.018 mg/ml (2.7 and 3.6 mg/kg/day for males and females, respectively). In another 90-day metam-sodium study (AlIen, 1991), male and female rats received metam-sodium in the drinking water at nominal dose levels of 0,0.018,0.089, and 0.443 mg/ml (1.7, 8.1 and 26.9 mg/kg/day for males; 2.5, 9.3, and 30.6 mg/kg/day for females). Systemic toxicity was evident by significant decreases in food and water consumption, decreased body weight gain, and histological changes in the nasal cavity olfactory epithelium in both males and females receiving metam-sodium at 0.443 mg/ml. Renal tubular dilation and basophilia along with increases in blood and protein in the urine were also observed in 0.443 mg/ml rats. In both males and females receiving 0.089 mg/ml there were significant decreases in red blood cell count and hematocrit. Females at the 0.089 mg/ml dose level also had a significant decrease in group mean body weight and decreased body weight gain (11 %) when compared to controls. Based on the results of this study the NOEL was 0.018 mg/ml (1.7 mg/kg/day for males; 2.5 mg/kg/day for females). In a subchronic dog study (Brammer, 1992), metam-sodium (43.15% purity) was administered by gelatin capsule to male and female beagles at nominal dose levels of 0, 1, 5, or 10 mg/kg/day once daily for 13 weeks. Toxic effects were observed at all dose levels tested but were primarily evident at the 5 and 10 mg/kg/day dose levels. Decreased body weight and body weight gain were observed in males and females receiving metam-sodium at 10 mg/kg/day. There were no significant clinical effects at 1 or 5 mg/kg/day and no ophthalmoscopic abnormalities in any animals. Regurgitation within 30-60 minutes of dosing occurred throughout the study in the 10 mg/kg/day group and on isolated occasions in the 5 mg/kg/day dogs. There was no regurgitation in the 1 mg/kg/day dosing group. In dogs receiving 5 and 10 mg/kg/day there were changes in hematologic parameters (increases in cell volume, cell hemoglobin, neutrophils, and monocytes; decreases in mean corpuscular hemoglobin concentration); significant increases in plasma alanine aminotransferase (ALT), aspartate aminotransferase (AST), alkaline phosphatase (ALP), and gammaglutamyltransferase; increased blood, urobilinogen, bilirubin, and protein in the urine; and microscopic evidence of hepatitis). One female receiving 1 mg/kg/day showed increased plasma ALT. Biliary duct proliferation with inflammatory cell infiltration (less severe than hepatitis) was observed in one male and one female at the 5 mg/kg/day dose level and in one female at the 1 mg/kg/day dose level. No evidence of tumors were found in this study. Toxic effects appeared to be dose- and timerelated. For female dogs, no systemic NOEL was established (NOEL < 1 mg/kg/day) due to increases in plasma ALT and biliary duct proliferation with inflammatory cell infiltration observed in a single female from the 1 mg/kg/day dose group. For male dogs, the systemic NOEL is 1 mg/kg/day. The lowest observed effect level (LOEL) of 5 mg/kg/day is based on statistically significant increases in plasma ALT, AST, and alkaline
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phosphatase, and the increased incidence of hepatitis and bile duct proliferation. In order to further study the effects of metam-sodium on the liver of dogs, a study was conducted at the dose level that caused moderate to marked hepatitis in all dogs during the 90day study described above (Brammer, 1993). One male and one female beagle dog received metam-sodium (43.14% purity) in a gelatin capsule daily at a dose level of 10 mg/kg. Dosing of each dog continued until there were elevations in plasma enzyme activities (or other clinical signs) indicative of liver toxicity. Following cessation of dosing, each dog was monitored until the enzyme activities returned to normal or prestudy levels. Dosing ceased after 12 weeks of dosing for the female and after 13 weeks for the male. Recovery was monitored for 8 weeks. After Week 6 the plasma ALT levels in the female began to increase and by Week 10 they were over 200 lUlL. In the male, elevated plasma ALT was noted at Week 9 and exceeded 200 UIIL by Week 11. In both dogs, plasma ALP levels gradually increased until dosing ceased. Following cessation of dosing, ALT levels increased during the first recovery week then gradually declined to normal levels after 8 weeks. Plasma ALP decreased in the female dog immediately after cessation of dosing and by Recovery Week 4 was less than prestudy values. In the male, ALP continued to rise during the first recovery week then gradually decreased so that by Recovery Week 5, ALP values were less than prestudy values. In both dogs, ALP levels continued to fall until study termination. At study termination, there were no macroscopic abnormalities in either dog and liver weights were normal. Microscopic evaluations revealed that there was a minimal or slight increase in the number of pigmented macropahges/Kupffer cells in the liver, but this is a common finding in beagle dogs of this strain (Alderley Park). The significant elevations in plasma ALT and ALP levels found in this study are consistent with the findings of the previous 90-day dog study at the same dose level (10 mg/kg/day) and are indicative of liver injury. However, after cessation of exposure, enzyme levels returned to normal, with full recovery eight weeks after the last exposure to metam-sodium. There was no evidence of liver injury at the end of the study. These findings confirm the reversible nature of induced liver effects from subchronic exposure to relatively high levels of metamsodium.
87.4 GENETIC TOXICITY Metam-sodium is not mutagenic but has been shown to be directly cytotoxic to bacteria, fungi, and mammalian cells. Metam-sodium has been tested and found to be negative in both in vitro and in vivo genetic toxicology assays covering a range of genetic toxicology endpoints including mutations, cytogenetics, and DNA repair. There is evidence that at high enough dose levels, exposure to metam-sodium can be immunotoxic, with response evident in a dose-dependent manner. A review of metam-sodium genetic toxicity studies (Mackay, 1996) concluded that "[M]etam sodium shows no in vitro or in
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Metam-Sodium
vivo genotoxic activity in a series of assays conducted up to concentrations/dose levels inducing significant toxicity in the target cells/animals."
both male and female mice. There were no statistically or biologically significant increases in the incidence of micronucleated polychromatic erythrocytes.
• In a bacterial gene mutation assay using Salmonella typhimurium strains TA92, TA98, TA 100, TA1535, TA1537, and TA1538 in the presence and absence of metabolic activation (AROCHLOR 1254-induced rat liver S9 mix) there were no significant increases in the number of revertant colonies in any of the strains or S9 combinations tested. • In a Chinese hamster ovary mammalian cell gene mutation (HGRPT locus) assay metam-sodium was tested in the presence and absence of metabolic activation. There was no evidence of any reproducible dose-related effects of metam-sodium on mutation frequency or evidence of in vitro mutagenic activity. • In two in vitro cytogenetic assays using human lymphocytes there was no evidence of clastogenic activity from metam-sodium treatment when tested at concentrations up to those limited by toxicity and/or cytotoxic effects on chromosomal morphology.
V.S. EPA Tox One liners report on the results of genetic studies submitted to and reviewed by the V.S. EPA. Jowa (1998) reported on the same genetic studies submitted to and reviewed by the California Environmental Protection Agency (Cal EPA). In some cases, the results presented by Jowa did not agree with conclusions of the V.S. EPA.
• The first study found an increase in aberrant cells at concentration levels that caused severe cytotoxicity (20 I-Lg/ml without S9 mix; 40 and 20 I-Lg/ml with S9 mix) and therefore were unsuitable to be included in the evaluation of clastogenic potential. At concentration levels of 1, 5, and 10 I-Lg/ml with and without S9 mix, there were no increases in the percentage of aberrant cells. • In the second in vitro human lymphocyte clastogenic study, metam-sodium at concentrations of 2.5, 20, and 30 I-Lg/ml in the absence of S9 mix and 5, 20, and 40 I-Lg/ml with S9 were tested. No statistically or biologically significant increases in the percentage of aberrant cells were observed at any of the metam-sodium concentrations tested in the absence of the S 9 mix. There was a small statistical increase in the number of aberrations observed in the 40 I-Lg/ml test concentration with the S9 mix, but the values observed were well within the historical solvent control range and do not indicate clastogenic activity. • In an in vitro unscheduled DNA synthesis assay using primary rat hepatocytes treated with metam-sodium there was no evidence of induction of DNA repair, even in cultures treated with toxic concentrations of metam-sodium. • In an in vivo Chinese hamster bone marrow chromosomal aberration assay there was no evidence of any polyploidy inducing effect of metam-sodium nor was there evidence of any clastogenic activity. • When metam-sodium was administered to CD-l mice in an in vivo mouse bone marrow micronucleus test, there was no evidence of clastogenic activity in the mouse bone marrow when tested up to the maximum tolerated dose level for
• In two separate Ames studies using multiple strains of Salmonella typhimurium (TA 1535, 1537, 1538,92,98, and 100) up to 2500 I-Lg/plate with and without activation (S9 mix), metam-sodium did not induce mutations and the results were negative. • In a study with yeast (Sacchromyces cerevisiae strain D4) with and without S9 mix, metam-sodium did not induce mutations and the results were negative. • In an in vitro study with Bacillus subtilis, metam-sodium did not cause DNA damage. • According to the Cal EPA review, equivocal results were obtained for a REC assay in Bacillus subtilis H17 and M45 (+/- S9). • According to the V.S. EPA review, metam-sodium (42.2%) is not a recombinogenic agent (i.e., causes DNA damage) to Bacillus subtilis strains H17 and M45 at concentrations up to 150 1-L1Iwell. • In an in vitro study using cultured lymphocytes procured from a single male human donor, there was evidence of possible aberrant chromosomes. However, according to the Metam-sodium Task Force, the scientific validity of this study is under question since the aberrant chromosomes were observed only at concentration levels that were clearly cytotoxic to the cells. When the cells from noncytotoxic concentration levels were evaluated, there was no indication of any clastogenic activity. • In a mammalian cytogenetic study with Chinese hamsters, metam-sodium did not induce cytogenic effects. • According to the Cal EPA review, there was evidence of polyploidy in Chinese hamster ovary cells at dose levels of 150 and 300 mg/kg. • According to the VS EPA review, metam-sodium (42.2%) had a negative response (no effect) in the Chinese hamster bone marrow cytogenetic assay at concentrations of 150, 300, and 600 mg/kg. • Metam-sodium was found to be negative in an unscheduled DNA synthesis study with primary rat hepatocyte culture. A study was conducted to assess the immunotoxicological and selected general toxicological effects of metam-sodium
87.5 Developmental and Reproductive Toxicity
(Pruett et aI., 1992). Metam-sodium was administered to female B6C3Fl mice at 200 mg/kg/day for 3, 5, 10, or 14 days. Selected organ weights were measured, hematological and bone parameters were examined, changes in thymus and spleen lymphocyte sUbpopulations were evaluated, and production of antibody-forming cells in vitro was measured. Major effects of metam-sodium administration included decreased thymus weight at all time points; increased spleen weight and bone marrow cellularity after 10 or 14 days of exposure; significant decreases in mature lymphocytes in the thymus and spleen; decrease in thymocytes; and decreased body weight. According to Pruett and co-workers (1992), overall patterns of change indicate that metam-sodium rapidly depletes most CD4 + CD8 + thymocytes, more slowly depletes a smaller number of mature lymphocytes in the thymus and spleen, and induces compensatory and/or detoxication mechanisms after 10-14 days of exposure. Pruett and co-workers (1992) conducted subsequent experiments to assess selected immune function parameters after exposure to metam-sodium. Metam-sodium was administered for seven days (either orally or dermally) and immunological assays were conducted on Day 8. Mice receiving metamsodium orally at dose levels of 50 to 300 mg/kg showed substantial, dose-dependent suppression of NK cell activity. Evaluation of humoral responses indicated that the cellular and molecular components required for humoral immune responses are not major targets for the acute effects of metamsodium. There was no suppression of antibody production in vivo or splenocyte responses to mitogens or allogeneic lymphocytes in vitro, which indicates that the lymphocytes which survive metam-sodium exposure are still able to proliferate and differentiate and are not significantly impaired with regard to function. The authors also noted that the pattern of thymic subpopulation changes is consistent with direct or indirect induction of apoptosis. These studies showed that immunological parameters could be significantly suppressed in the absence of a significant decrease in body weight, suggesting that most of the effects of metam-sodium on the immune system are not secondary to generalized toxicity. In response to reports that metam-sodium is immunotoxic, a series of in vivo and in vitro studies was conducted with metamsodium and other dithiocarbamates (Padgett et aI., 1992). Metam-sodium in distilled water was administered orally via daily gavage to female mice for seven days at dose levels of 0, 150, 225, or 300 mg/kg. Body weight was not significantly decreased at any dose level, but thymus weight was significantly decreased in mice receiving metam-sodium at dose levels of 225 and 300 mg/kg. In tests of splenic NK cell activity, metam-sodium at dose levels of 225 and 300 mg/kg was found to significantly inhibit NK activity. This study also demonstrated that metam-sodium was directly cytotoxic to lymphoid cells in vitro, but that cytotoxic potency in vitro does not correlate well with immunological changes in vivo.
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87.5 DEVELOPMENTAL AND REPRODUCTIVE TOXICITY Developmental studies in two different species found evidence of increased fetal loss, increased skeletal variations, and developmental delays from oral administration of metam-sodium to pregnant animals at dose levels that also caused overt maternal toxicity. Visceral or skeletal abnormalities were not present at low dose levels but increased in incidence and severity with increasing dose (U.S. EPA, 1991). In a multigeneration reproductive study, metam-sodium did not affect reproductive performance, even at toxic dose levels (U.S. EPA, 1994a). A developmental study with rats receiving metam-sodium (Hellwig and Hildebrand, 1987) indicated that there were significant maternal and fetal effects at higher dose levels and that these effects were dose-related. An aqueous solution of metamsodium (42.2%) was administered at 0, 10,40, or 120 mg/kg by gavage to pregnant Wistar rats on Days 6-15 of gestation. Body weight gains were significantly decreased in dams receiving metam-sodium at dose levels of 40 and 120 mg/kg during the dosing period. Cesarean section observations revealed that there was a statistically significant increase in the percentage of postimplantation loss and a significant decrease in the percentage of live fetuses per dam at the 10 and 120 mg/kg dose levels, but not at the 40 mg/kg dose level. It is possible that the effects observed at the 10 mg/kg dose level were statistical anomalies, but this remains unconfirmed in the absence of a review ofthe individual data, which were not available. All other parameters in the 10 mg/kg group were comparable to controls, including the total number of live fetuses and live fetuses per dam. Since there were no statistically significant changes in Cesarean section observations in the 40 mg/kg group, it is likely that the statistically significant changes in percentage of live fetuses per dam and the percentage of postimplantation loss in the 10 mg/kg group are not treatment-related. The only abnormal finding observed during the macroscopic examination of the fetuses was meningocele (hernial protrusion of the meniges through a bony defect) in two fetuses from one litter in the 120 mg/kg dose group. Since this is a rare finding that was not present in historical controls, this anomaly was considered to be treatment-related. Skeletal evaluations of the fetuses revealed an increased incidence of variations and a delay in the development of fetuses in the 40 and 120 mg/kg dose groups. Fetal weights were significantly reduced in the 120 mg/kg group. The NOEL for fetal and maternal effects was 10 mg/kg. In another rat developmental toxicity study (Tinston, 1993) groups of pregnant rats were administered metam-sodium at dose levels of 0, 5, 20, or 60 mg/kg/day on Days 7-16 (inclusive) of gestation. Maternal toxicity evidenced by reduced body weight gain, reduced food consumption, and the presence of clinical signs (piloerection, salivation, and urinary incontinence) occurred at the 20 and 60 mg/kg/day dose levels. Body weight gain and food consumption were marginally reduced at the 5 mg/kg/day dose level but there were no treatment-related clinical signs. In both the 20 and 60 mg/kg/day dose groups
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Table 87.1 Summary of Metam-Sodium No Observed Effect Levels (NOELs) and Lowest Observed Effect Levels (LOELs) Dosing Study
Species
duration
Dose levels
NOEL
LOEL effects
90-day drinking water
Mouse
90 days
0,0.018,0.088,0.35, and 0.62 mg/ml
0.018 mg/ml
0.088 mg/ml
0,2.7, 11.7,52.4, and 78.7 mg/kg/day
2.7 mg/kg/day
• urinary bladder lesions (d)
0,3.6, 15.2,55.4, and 83.8 mg/kg/day
3.6 mg/kg/day
• decreases in hemoglobin, RBC, and hematocrit
('i')
90-day drinking water
Rat
90 days
0,0.018,0.089, and 0.443 mg/ml
0.018 mg/ml
0,1.7,8.1, and 26.9 mg/kg/day (d)
1. 7 mglkg/day
and body weight gain
0,2.5,9.3, and 30.6 mg/kg/day ('i')
2.5 mg/kg/day
• decreases in RBC and
0, 1,5, and 10 mg/kg/day
1 mg/kg/day 0'
0' 5 mg/kg/day
0.089 mg/ml • decreased body weight
hematocrit 90-dayora1
Dog
90 days
• increased plasma ALT, AST, and ALP • hepatitis and bile duct proliferation <1 mglkg/day 'i'
'i' 1 mglkg/day
• increased plasma ALT • bile duct proliferation Developmental toxicity
Rat
10 days
0, 10,40, and 120 mglkg/day
10 mg/kg/day
40 mg/kg/day • decreased maternal weight gain • increased fetal skeletal variations • delay in development
Developmental toxicity
Rat
10 days
0, 5, 20, and 60 mg/kg/day
5 mg/kg/day
20 mg/kg/day • decreased maternal weight gain • reduced food consumption • maternal clinical signs • increased fetal skeletal variations • reduced ossification of manus and pes • reduced fetal weights
Developmental toxicity
Rabbit
13 days
0, 10, 30, and 100 mg/kg/day
10 mg/kg/day-fetal
30 mg/kg/day-fetal • decreases in live fetuses • increased resorptions
30 mg/kg/day-maternal
100 mg/kg/day-maternal • decreased body weight gain
Developmental toxicity
Rabbit
13 days
0, 5, 20, and 60 mglkg/day
5 mg/kg/day
20 mg/kg/day • reduced maternal body weights • change in fetal ossification pattern
(continues)
87.5 Developmental and Reproductive Toxicity
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TableS7.1 continued Dosing Study
Species
duration
Dose levels
NOEL
LOEL effects
MuItigeneration
Rat
Chronic
0,0.01,0.03, and 0.1 mg/ml
0.03 mg/ml-systemic
0.1 mg/ml-systemic toxicity • changes in Bowman's
0, 1.2,3.2, and 11.5 mg/kg/day (d')
gland and olfactory
0, 1.8, 3.9, and 13.5 mg/kg/day (I')
epithelium (adults) • decreased mean pup weight
Carcinogenicity: two-year
Rat
Chronic
0,0.019,0.056, and 0.19 mg/ml
0.1 mg/ml-reproductive
>0.1 mg/ml
0.056mg/ml
0.19mg/ml
drinking
• decreased body weight gain
0, 1.3, 3.9, and 12.0 mg/kg/day (d') 0,2.3,6.2, and 16.2 mg/kg/day (I')
• decreased food consumption, food efficiency and water consumption • changes in hematology and clinical chemistry • abnormalities in nasal cavity, voluntary muscle and sciatic nerve
Carcinogenicity: two-year
Mouse
Chronic
0,0.019,0.074, and 0.23 mg/ml
0.019 mg/ml
drinking
• increased liver weight 0, 1.6,6.5, and 27.7 mg/kg/day (0')
• changes in kidney and epididymis weights
0,2.3, 8.7, and 29.9 mg/kg/day (I') I-year oral
0.074mg/ml
Dog
Chronic
0,0.05,0.1, and 1.0 mg/kg/day
0.1 mg/kg/day
1.0 mg/kg/day • increase in hepatocyte and liver macrophage/ Kupffer cells • increased plasma ALT
Acute neurotoxicity
Rat
Single dose
0,22, 324, and 647 mg/kg
<22 mg/kg
22 mg/kg • reduced ambulatory and total motor activity
Subchronic neurotoxicty
Rat
13 weeks
0, 0.02, 0.06, and 0.2 mg/ml
0.06 mg/ml (0')
0, 1.4,5.0, and 12.8 mg/kg/day (d')
0.02 mg/ml (I')
0.2 mg/ml (d') 0.06 mg/ml (I')
0,2.3,7.0, and 15.5 mg/kg/day (I')
there was an increase in fetal effects (reduced fetal weights, reduced ossification of manus and pes, and increased incidences of minor skeletal defects and/or variants). The no observed adverse effect level (NOAEL) for maternal toxicity or fetal effects in this study was 5 mg/kg/day. In a teratology/developmental study (Hellwig, 1987), pregnant Himalayan rabbits were administered a 42.2% aqueous solution of metam-sodium at dose levels of 0, 10, 30, or 100 mg/kg by gavage from gestation Days 6 through 18. Evaluation of body weight data revealed a treatment -related decrease in body weight gain in the 100 mg/kg dams. There were no statistically significant treatment-related effects noted in food consumption or food efficiency. Cesarean section observations
• decreased body weight gain
revealed statistically significant decreases in the total number of live fetuses and statistically significant increases in total resorptions in the 30 and 100 mg/kg/day groups. Macroscopic fetal examinations revealed meningocele and spina bifida in one rabbit in one litter in the 100 mg/kg/day group (it is not clear from the data if the findings were in the same rabbit or in two separate rabbits). Due to the rarity of this event and that it was also present in the rat developmental study, this abnormality is considered to be treatment-related. There were no treatment-related effects noted from the visceral examinations. Skeletal examinations revealed no treatment-related effects. However, these examinations were done using acceptable European methods that have not been validated by EPA and are
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not considered to be comparable with D.S. EPA-accepted methods. In another rabbit developmental toxicity study (Hodge, 1993) groups of pregnant rabbits were administered metamsodium at dose levels of 0, 5, 20, or 60 mg/kg/day on Days 8-20 (inclusive) of gestation. At the 60 mg/kg/day dose level, dams showed marked weight loss and reduced food consumption. At 20 mg/kg/day, body weight of dams was slightly reduced. There were no observable effects noted on dams at the 5 mg/kg/day dose level. Fetal examinations revealed a marked increase in embryonic lethality at the 60 mg/kg/day maternal dose level and changes in ossification pattern at the 20 and 60 mg/kg/day maternal dose levels. The NOAEL for maternal and developmental toxicity was 5 mg/kg/day. In a multigeneration reproduction study (Milburn, 1993), Alpk:ApfSD rats received metam-sodium in drinking water at the following concentrations: 0, 0.01, 0.03, or 0.1 mg/ml. These concentrations corresponded to dose levels of 0, 1.2, 3.2, or 11.5 mg/kg/day for males and 0, 1.8, 3.9, or 13.5 mg/kg/day for females. After the first 10 weeks of treatment, animals were mated on a one-to-one ratio. Males were then removed from their cages and females were allowed to give birth and raise pups. At 21 days of age, pups from the parental (FO) generation were selected as parents for the Fl generation. In parents, body weights were marginally reduced in rats receiving 0.10 mg/ml (the highest concentration tested) during the premating period and markedly reduced during pregnancy and lactation. Water consumption was reduced in the 0.10 mg/ml rats throughout the study and to a lesser extent in the 0.03 mg/ml group. In offspring, there was a marginal reduction in food consumption during the premating period in the FO and Fl rats in the 0.10 mg/ml group, but there were no effects on food consumption in the 0.01 and 0.03 mg/ml treatment groups. Offspring body weights and total litter weights were reduced in the 0.10 mg/ml group in both generations. There were no effects on any of the reproductive parameters at any treatment level for parents or offspring. Histopathological evaluations indicated increased changes in the epithelium of the nasal passages of the FO and Fl adult females in the 0.10 mg/ml groups. This effect was not observed in 0.10 mg/ml adult males or in male or female offspring of either generation. No treatmentrelated histopathological changes were observed in rats receiving metam-sodium at concentrations of 0.01 or 0.03 mg/ml. Metam-sodium did not affect reproductive performance at any dose level tested. Evidence of toxicity was observed only at the highest concentration level tested, i.e., 0.1 mg/ml. In adult female rats receiving metam-sodium at the 0.1 mg/ml concentration level (13.5 mg/kg/day), evidence of systemic toxicity consisted of (1) duct hypertrophy of Bowman's gland with loss of alveolar cells, (2) degeneration, disorganization, and/or atrophy of the olfactory epithelium, and (3) dilation of the Bowman's gland ducts. Changes in Bowman's glands were accompanied in all affected animals by degeneration, disorganization, and/or atrophy of the olfactory epithelium. In pups in the 0.1 mg/ml group, evidence of toxicity consisted of a 14% decrease in mean pup weight on Day 22 for the Fl generation,
a 16% decrease in mean body weight gain for F21itters, and decreases of 8-9% in testes and epididymis weight in male pups in the Fla and F2a litters. The NOEL for systemic toxicity (adults and pups) was 0.03 mg/ml. The NOEL for reproductive effects was 0.1 mg/ml (11.5 mg/kg/day for males, 13.5 mg/kg/day for females).
87.6 CHRONIC/ONCOGENICITY TOXICITY A two-year combined chronic toxicity/carcinogenicity study demonstrated that metam-sodium shows no carcinogenic potential in rats (Thomassen, 1998; D.S. EPA, 1994b). However, a two-year carcinogenicity study in mice revealed an increased incidence of angiosarcoma in mice at higher dose levels (D.S. EPA, 1994c). A one-year study with dogs showed no evidence of carcinogenicity but evidence of liver damage similar to but less severe than the effects (that were shown to be reversible) observed in previous subchronic metam-sodium dog studies. In a two-year combined chronic toxicity/carcinogenicity study with Wistar rats (Rattray, 1994), metam-sodium (43.14% a.i.) was administered in drinking water at concentration levels of 0,0.019,0.056, or 0.19 mg/ml (achieved dosages of 0, 1.3, 3.9, or 12.0 mg/kg/day for males and 0, 2.3, 6.2, or 16.2 mg/kg/day for females). There was no evidence of an adverse effect of metam-sodium on the survival or rats. There were no ophthalmological changes associated with metam-sodium treatment. Evidence of toxicity was present in both males and females at the highest concentration level tested, i.e., 0.19 mg/ml. At 0.19 mg/ml, male and female rats had decreased mean body weight gain for Weeks 1-13 (12% for males, 16% for females) and for Weeks 1-105 (18% for males, 20% for females). Food consumption, food efficiency, and water consumption were significantly decreased for both males and females receiving 0.19 mg/ml metam-sodium. Effects were also observed in 0.19 mg/ml male and female hematology (decreased red blood cells, hemoglobin, and hematocrit) and clinical chemistry (decreased cholesterol and triglycerides). Nasal passages were identified as the target organ. Microscopic abnormalities of the nasal cavity were mainly confined to 0.19 mg/ml animals. These changes included (1) an increased incidence of rhinitis, (2) hypertrophy of Bowman's ducts/glands, (3) atrophy and adenitis of Steno's gland, and (4) hyperplasia and degeneration of olfactory epithelium. The incidence of degenerative myopathy of voluntary muscle was similar in all groups, including controls. However, there was an increase in the severity of myopathy in animals in the 0.019 mg/ml group. There was no indication of an increased incidence of neoplasia or early onset of tumors from treatment with metam-sodium. Evaluation of the tumor incidence demonstrated that metam-sodium shows no carcinogenic potential in rats (Rattray, 1994; U.S. EPA, 1994b). The NOEL for both male and female Wistar rats was 0.056 mg/ml. The Cal EPA, Department of Pesticide Regulation evaluated the tumor d{ita from the two-year metam-sodium drinkingwater study in nits and concluded that there was a possible
87.6 Chronic/Oncogenicity Toxicity
tumorigenic effect at the 0.056 mg/ml concentration level (D.S. EPA, 1995). According to the Cal EPA review, the incidence of hemangiosarcoma (8/64) was increased at this dose, in relation to the control incidence (0/64) and the high dose (0.19 mg/ml) incidence (3/64). The hypothesis that this could be a positive response was based on the positive findings in the twoyear mouse study and that this increased incidence could be based on decreased body weight observed at the high dose in relation to other doses. When the D.S. EPA (1995) re-evaluated the tumor data for their Carcinogenicity Classification, they did not find the effect that the Cal EPA found (presumably because the Cal EPA analysis did not exclude animals that died before observation of the first tumor). However, there was a significant pairwise comparison in the incidence of hemangiosarcoma in male rats at the 0.019 and 0.056 mg/ml (1.3 and 3.9 mg/kg/day) levels when compared to controls. The D.S. EPA also considered debatable the hypothesis of increased incidence of hemangiosarcoma at the mid-dose level based on decreased body weight in male rats at the high dose level. Rats in this study were not fed a calorierestricted diet, nor was their access to food controlled. In addition, the decreases in body weight gain were observed for both male and female rats, although the preponderance of hemangiomasihemangiosarcomas was observed only in male rats. In addition, the time to tumor formation was observed at approximately the same time in all dose levels. In calorie-restricted studies, the numbers of tumors are often reduced in conjunction with a delay in the time to tumor formation. In response to the position taken by Cal EPA, the two-year drinking water study with Wistar rats was reviewed and compared to an expanded historical control data base for Wistar rats that at the time of their original review was not available (Thomassen, 1998). Hemangiomatous tumors (hemangioma and/or hemangiosarcoma) were observed only in rats sacrificed at termination of the study (Study Week 105) or in rats that were found dead or were euthanized due to their clinical condition (moribund or to prevent suffering). No hemangiomatous tumors were observed in rats euthanized during the interim sacrifice at Study Week 53. Therefore, tumor analysis (as was done by the D.S. EPA) should exclude animals that were sacrificed or died prior to the first observance of a hemangioma or hemangiosarcoma. Statistically significant increased incidence of hemangiosarcomas occurred only in the 0.019 and 0.056 mg/ml males and not in the 0.19 mg/ml males, although the actual numbers were very similar (3/49 at 0.019 mg/ml and 3/51 at 0.19 mg/ml). Three possible explanations for reduced tumor incidence with an increase in treatment are: (1) the high dose of metam-sodium exceeded the maximum tolerated dose and had a negative impact on the tumor response in the high dose males; (2) reduced body weight associated with reduced tumor incidence accounted for the difference (as suggested by the Cal EPA reviewer); or (3) biological variability was responsible. Neither the U.S. EPA nor the Cal EPA thought the maximum tolerated dose had been exceeded. The D.S. EPA carcinogenicity peer review panel did not believe that reduced body weight accounted for the reduced tumor incidence (D.S.
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EPA, 1995). However, the possibility that biological variability could account for the effect was hampered by the lack of historical control data for Wistar rats. Hemangiomatous tumors (variously diagnosed as angiomas and angiosarcomas, hemangiomas and hemangiosarcomas, and lymphangiomas and lymphangiosarcomas) are common in some but not all strains of Wistar rats (Bomhard, 1992; Bomhard et aI., 1986; Crain, 1958; Deerberg et al., 1980; Kroes et aI., 1981; Rehm et aI., 1984). The reported incidences of these tumors in Wistar-derived rats used in European laboratories vary considerably, but there are reports of up to a 74% incidence for males and 44% for females (Rehm et aI., 1984). These reports also indicate that there is a definite propensity for development of tumors in the lymph nodes, particularly the mesenteric lymph nodes of male Wistar rats. Although Zeneca Central Toxicology Laboratory did not have an historical control data base for Wistar rats used in this study, there was a large historical control tumor data base compiled by several European laboratories utilizing Wistar-derived rats (49 studies ranging in duration from 24 to 31 months). This data base was published as the RITA Wistar Rat Control Tumor Data Base (Thomassen, 1998). Information presented in the RITA control data base is consistent with the types and numbers of tumors observed in the metam-sodium two-year rat study. Based on a thorough review of the original study and comparisons with the RITA control data base for Wistar rats, it was concluded that: • Metam-sodium is not a carcinogen in the rat. • The reduced number of hemangiosarcomas in the high dose male rats in the metam-sodium study is not due to reduced caloric intake. • The natural distribution and incidence of spontaneously occurring hemangiosarcomas in untreated male Wistar rats can account for the distribution and incidence of hemangiosarcomas observed in male rats treated with metam-sodium. In a two-year carcinogenicity study in mice (Homer, 1994), metam-sodium (43.15%) was administered in the drinking water to C57BL/I0JfCD-lfAlpk mice for 104 weeks at nominal concentration levels of 0,0.019,0.074, or 0.23 mg/ml (actual achieved doses of 0, 1.6,6.5, or 27.7 mg/kg/day for males and 0,2.3,8.7, or 29.9 mg/kg/day for females). Metam-sodium did not adversely affect survival of mice at any dose level. Clinical signs of toxicity were considered to be unremarkable. Male and female mice receiving metam-sodium at 0.074 and 0.23 mg/ml had dose-related and statistically significant increases in absolute liver weight when compared to controls (Ill % and 119% for 0.074 mg/ml males and females, respectively; 135% and 122% for 0.23 mg/ml mice). At 0.23 mg/ml, male mice had decreased body weight gain of 14% for Weeks 1-13 and 20% for Weeks 1-104. Food consumption was unaffected during the early part of the study, but during Weeks 24 to 52 there were statistically significant decreases in food consumption for the 0.074 and 0.23 mg/ml male mice. Decreases in food consumption were not observed for female mice. Water consumption
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was significantly decreased for both males and females in the 0.23 mg/ml group during the study's first week, but by Week 9, males in the 0.23 mg/ml group had significantly increased water consumption. By Week 11, water consumption was significantly increased for both 0.074 and 0.23 mg/ml males. By Week 48, water consumption for all groups (male and female) were approximately equal to controls. Hematological investigations showed no significant treatment-related effects at any dose level. Macroscopic observations revealed several changes in 0.23 mg/ml mice including liver appearance (accentuated lobular pattern, pale), subcutaneous tissue masses, urinary bladder wall thickening, and reduced incidence of enlarged seminal vesicles. Several changes were noted in liver, kidney, and epididymis weights at 0.074 and 0.23 mg/ml treatment levels. Microscopic evaluations revealed several non-neoplastic effects in 0.23 mg/ml mice but also revealed evidence of neoplastic changes at this same dose level. There was evidence of dose-dependent metam-sodium induced carcinogenicity in mice. In both males and females at the 0.23 mg/ml treatment level, there was an increased incidence of hepatic adenoma and angiosarcoma, splenic angiosarcoma, subcutaneous tissue angiosarcoma, and a single incidence of a urinary bladder transitional cell papilloma in one high dose male and a single incidence of urinary bladder transitional cell carcinoma in one high dose female. The overall incidence of angiosarcoma, regardless of site, increased for both males and females in the 0.23 mg/ml treatment group when compared to concurrent as well as historical controls. The no observed effect level for neoplastic changes is 0.074 mg/ml. According to the U.S. EPA (1994c), there was equivocal evidence of a possible increase in splenic angiosarcoma at 0.074 mg/ml [something that the study author and registrants believe is related to the difficulty in determining the primary site(s) of angiosarcoma]. There was no evidence of increased tumors at the lowest dose level. In the U.S. EPA's (1994c) evaluation of the two-year mouse study, the reviewers suggested that the dosing levels in this study were adequate due to the degree of toxicity (increased liver weights, non-neoplastic changes in bladder, and tumors) observed in both males and females at the 0.23 mg/ml treatment level. According to the U.S. EPA (1994c) based on the significant increase observed in liver weight in male and female mice, the LOEL is considered to be 0.074 mg/m!, which is the NOAEL for neoplastic changes. In a one-year toxicity study (Brammer, 1994), metamsodium was administered orally to beagle dogs at dose levels of 0,0.05,0.1, or 1.0 mg/kg/day. Animals were observed daily for food consumption, evidence of gastro-intestinal upset, and changes in clinical condition. Animals also received detailed clinical evaluations weekly and complete veterinary examinations (including ophthalmoscopy) every three months. Blood chemistry, hematology, urine chemistry, and cytology evaluations were conducted at regular intervals throughout the study. At study termination, each animal received a full necropsy and histopathological evaluation of selected tissues. Throughout the study, there were no overt signs of toxicity at any dose level and all dogs remained in good health. There were no toxico-
logically significant effects on body weight, food consumption, clinical condition, or on the incidence of gastro-intestinal effects (i.e., vomiting, loose stools, etc.). There were no ophthalmoscopic abnormalities nor were there significant changes in hematology or urinalysis or in organ weights. There were no macroscopic findings that could be attributed to treatment with metam-sodium. Microscopic evaluations revealed a slight increase in hepatocyte and macrophage/Kupffer cells in the liver of one female dog dosed at 1.0 mg/kg/day. This same female also had significant elevations in plasma alanine transaminase activity. These changes were similar to but less severe than those observed in previous subchronic dog studies with metamsodium and are considered to be treatment-related. Therefore, the NOEL for this study was 0.1 mg/kg/day.
87.7 NUROTOXICITY Metam-sodium is not neurotoxic based on evidence from neurotoxicity studies. In an acute neurotoxicity study (Lamb, 1993), male and female Sprague-Dawley Crl: CD®BR rats received metamsodium (43.15%) orally at doses of 0, 50, 750, or 1500 mg formulated metam-sodiumlkg body weight or 0, 22, 324, or 647 mga.i./kg. Mortality was observed at the1500 mg/kg dose level (males 31 %, females 19%). Signs of systemic toxicity were observed at the 750 and 1500 mg/kg dose levels and included changes in posture, palpebral closure, respiratory rate, arousal, rearing activity, time to first step, olfactory and pupil responses, tail pinch response, hindlimb strength, body temperature, and body weight. Lacrimation and salivation were also noted among some animals at both the 750 and 1500 mg/kg dose levels. Reductions in ambulatory and motor activity were observed at the 50 mg/kg dose level and above on Day 0 (day of dosing) yet there were no treatment-related effects on the functional observational battery in the 50 mg/kg dose group. No signs indicative of neurotoxicity were observed at any dose level. There was no significant change in brain cholinesterase (ChE) activity at any dose level and there were no signs of cholinergic effects at any dose level. There were no treatment related differences in brain weight or dimensions in any treatment group. Histopathological evaluations of brain and nervous system tissues showed no evidence of neurotoxicity. According to the U.S. EPA (1994d), plasma and RBC ChE activity levels were reduced in 1500 mg/kg male and female rats 24 hours postdose (6% and 12% for male plasma and RBC ChE, respectively; 24% and 14% for females). [Although statistically significant, none of these decreases in ChE activity are considered to be biologically relevant as all decreases are well within the range of normal variation and are below the thresholds set by the World Health Organization (JMPR, 1995; WHO, 1990) and other regulatory agencies (Carlock et aI., 1999).] Based on the results of this study, the 1500 mg/kg dose level was considered the NOAEL in males and females for acute neurotoxicity while the 50 mg/kg/day dose level was considered
87.9 Metabolism
the LOEL for acute systemic toxicity (based on reduced motor activity). In a subchronic neurotoxicity study (AlIen, 1991), male and female Sprague-Dawley rats were given metam-sodium (43.15%) in drinking water at concentration levels of 0,0.02, 0.06, or 0.2 mg/ml for 13 weeks (achieved dosages of 0, lA, 5.0, or 12.8 mg/kg/day for males and 0, 2.3, 7.0, and 15.5 mg/kg/day for females). Male and female rats administered 0.2 mg metam-sodiurnlml drinking water showed reductions in body weight, food consumption, and water consumption. Similar effects were observed in females at the 0.06 mg/ml concentration level. Body weight gain was reduced 14% for the 0.2 mg/ml males and the 0.06 mg/ml females, and 18-21 % for the 0.2 mg/ml females. Food utilization was slightly reduced in males at the 0.2 mg/mllevel. Reduced water consumption was also observed in males at the 0.06 mg/mllevel and in females at the 0.02 mg/mllevel. All of these effects were considered to be a consequence of poor potability of the drinking water rather than toxicity of metam-sodium. A functional observational battery and comprehensive neuropathological examination of the peripheral and central nervous systems revealed no evidence of any effects attributable to treatment with metam-sodium. Since there was no evidence of a neurotoxic effect from metam-sodium, the NOAEL for neurotoxicity is 0.2 mg/ml.
87.8 OTHER STUDIES (MAMMALIAN) In vitro percutaneous absorption of metam-sodium through rat and human skin was evaluated (Clowes, 1993). Metam-sodium was applied at dose levels of 940 and 94.0 J.tg/cm 2. Ten hours after dermal application, the skins were washed to determine how much of the dose could be removed from the skin surface, receptor fluid was analyzed, and the proportion of the dose remaining associated with the skin after washing and the amount absorbed were quantified. The absorption of metam-sodium was found to be dose and time dependent through both rat and human skin. The highest amount of metam-sodium absorbed was through rat skin from the 940 J.tg/cm 2 application (mean 200 J.tg/cm2; 21.3% of the applied dose at 10 hours). A correspondingly smaller amount was absorbed from the 94.0 J.tg/cm2 application through rat skin (mean 18.2 J.tg/cm2; 1904% at 10 hours). Absorption through cadaver human skin was 2.19% for the 940 J.tg!cm2 dose (mean 20.6 J.tg/cm 2 ) and 12.2% (mean 11.5 J.tg/cm2) of the applied dose. Absorption of metam-sodium through both rat and human skin increased with time but at a decreasing rate over the 10 hour period. The percentage of the dose remaining in the skin increased with decreasing dose. There was less metam-sodium absorbed through human skin than rat skin, and at the highest dose level there was approximatelya 1O-fold decrease in absorption of metam-sodium by human skin when compared to rat skin. An in vivo percutaneous (dermal) absorption study in the rat (Stewart, 1992) showed that metam-sodium and/or its radiolabeled degradation products are only poorly absorbed following
1877
a single dermal application to the rat. Radiolabeled 14C metamsodium was applied to male rats in aqueous solution at nominal dose levels of 0.1, 1, and 10 mg/animals. A glass saddle containing an activated charcoal filter to adsorb any volatile radioactivity evaporating from the skin surface protected the application site. Four animals from each group were evaluated at 1, 2, 10, and 24 hours after treatment for radioactivity in the excrement, in and on the skin, and in the body. Another four animals per group had the treatment area washed 10 hours after administration. Radioactivity in the excrement was monitored over a total of 72 hours prior to evaluations of the skin and body. Overall mean recoveries of radioactivity were in the range of 83.5 to 95.7% of the applied dose. The extent of absorption was similar at each dose level with an overall mean of approximately 3%. In general, absorption increased with time. Levels of absorbed material 24 hours postapplication for the 0.1, 1, and 10 mg/animal dose levels were approximately 7.5,50, and 231 J.tg equivalents of 14C metam-sodium, respectively. Substantial quantities of the nonabsorbed dose were recovered from the charcoal, suggesting that metam-sodium or its degradation products are highly volatile. At the 0.1, 1, and 10 mg/animal dose levels, the amounts of metam-sodium absorbed over a 10 hour exposure period were 204, 3.7, and 1.5% of the applied dose, respectively. Absorbed radioactivity was either eliminated in urine or exhaled and subsequently trapped in expired air traps. Less than 0.7% of the applied dose was recovered in feces and the recovery of radioactivity from the carcass ranged from below the limit of detection to 1.2%. Following applications of metam-sodium at 1 and 10 mg/animal, concentrations of radioactivity in blood and plasma peaked at one hour postdose. Levels of radioactivity in blood and plasma in the 0.1 mg/animal group were below the limit of detection. This study showed that (1) metam-sodium is poorly absorbed following a single dermal application; (2) absorbed radioactivity is rapidly excreted, primarily via the urine and expired air; and (3) washing the application site with soap and water effectively removes the majority of the applied dose. A further dermal absorption study with metam-sodium showed that absorption for rats was only 2.5% of the applied dose (U.S. EPA, 1994a). Radiolabeled 14C metam-sodium was applied to shaved dorso-lumbar skin sites of rats at concentrations of 8.6, 86.2, or 862 J.tg!cm2. Animals were exposed to metam-sodium for 1, 2, 10,24, or 72 hours. At the end of the study, total dermal absorption after 72 hours was determined to be 2.5% of total applied dose. Since metam-sodium is poorly absorbed dermally, human skin surfaces are acidic, and sweat is approximately pH5, metam-sodium that may come in contact with skin is expected to rapidly degrade prior to absorption.
87.9 METABOLISM After oral ingestion, metam-sodium is rapidly absorbed, metabolized, and excreted from the body. Exhalation and excretion in the urine are the major elimination pathways after oral
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Metam-Sodium
exposure. Metam-sodium is poorly absorbed following dermal application but the metam-sodium that is absorbed is rapidly excreted, primarily through the urine and expired air. In study of biokinetics and metabolism, radiolabeled metamsodium [14C] (purity> 99%) was administered to SpragueDawley rats at dose levels of 10 or 100 mg/kg (Hawkins et aI., 1987). Blood, urine, and feces were tested for radioactivity for up to seven days postdosing while expired air was collected up to 72 hours postdose. The results of this study showed that metam-sodium was rapidly and completely absorbed after oral ingestion. Radioactivity in plasma reached a maximum level in 1 hour and decreased to near background levels by 24 hours. Animals receiving 4C] metam-sodium at 10 mg/kg eliminated approximately 25% of the total radioactivity through the urine during the first 8 hours. By 168 hours, 55% of the total activity had been eliminated through the urine and 3-4% through feces. At the 100 mg/kg dose, 18% of the total activity had been eliminated in the urine by 8 hours and 40% by 168 hours. Within 24 hours, expired air from 10 mg/kg rats contained approximately 32% of the radioactivity with 1% MITC, 15% carbon disulfide (CS2)/carbonyl sulfide (COS), and 17% carbon dioxide. At 24 hours for the 100 mg/kg dose level, expired air contained approximately 48% of the total radioactivity with 24% MITC, 18% CS2/COS, and 6% C02. Negligible amounts of radiolabeled material were expired from 24 to 72 hours at either dose level. Approximately 98% of the radioactivity had been eliminated by the seventh day, with only 2% of the activity remaining in the tissues. The highest concentration of radioactivity was found in the thyroid, but significant concentrations were also found in the liver, kidneys, and lungs. Analysis of the urinary metabolites found that glutathione conjugation with MITC is the source of the major urinary metabolite, N -acetyl-S-(N -methylthiocarbamoyl)-l-cysteine, which accounted for 21 % of the excreted dose. No evidence for glucuronide or sulfate conjugates of the metabolites of metamsodium was found. Based on the results of this study, it appears that metam-sodium degrades to either CS2 or MITC in the stomach (accelerated by the stomach pH). MITC is eliminated either through exhalation or in the urine after glutathione conjugation in the liver. CS2 is eliminated by exhalation or further metabolized in the liver to C02 prior to elimination. Therefore, two different metabolic pathways, CS2 metabolism and MITC conjugation, are involved in urinary elimination. At higher dose levels, saturation of the metabolic processes results in greater exhalation of unmetabolized products. The in vivo dermal absorption study conducted by Stewart (1992), which was described previously, further demonstrated that (1) metam-sodium is poorly absorbed following a single dermal application; (2) absorbed radioactivity is rapidly excreted, primarily via the urine and expired air; and (3) washing the application site with soap and water effectively removes the majority of the applied dose. In both this study and the Hawkins et al. (1987) metabolism study, peak blood and plasma radioactivity levels occurred one hour after dosing.
e
REFERENCES AlIen (1991). Bomhard, E. (1992). Frequency of spontaneous tumors in Wistar rats in 3-month studies. Exp. Toxic. Pathol. 44, 381-392. Bomhard, E., Karbe, E., and Loesser, E. (1986). Spontaneous tumors of 2000 Wistar TNOIW.70 rats in two-year carcinogenicity studies. J. Environ. Path. Toxieol. Oneol. 7, 35-52. Brammer, A. (1992). "Metam-Sodium: 90-Day Oral Dosing Study in Dogs." Unpublished study (Rep. CTL1P13679) conducted by Zeneca Central Toxicology Laboratory, Alderley Park, Macclesfield, Cheshire, UK. Submitted by Metam-sodium Task Force. Brammer, A. (1993). "Metam-Sodium: Assessment of Recovery in Dogs." Unpublished study (Rep. CTL/L/5204) conducted by Zeneca Central Toxicology Laboratory, Alderley Park, Macclesfield, Cheshire, UK. Submitted by Metam-sodium Task Force. Brammer, A. (1994). "Metam-Sodium: I-Year Oral Toxicity Study in Dogs." Unpublished study (Rep. CTL1P14196) conducted by Zeneca Central Toxicology Laboratory, Alderley Park, Macclesfield, Cheshire, UK. Submitted by Metam-sodium Task Force. Carlock, L. L., Chen, W. L., Gordon, E. B., KilIeen, J. C., Manley, A., Meyer, L. S., MulIin, L. S., Pendino, K. J., Percy, A., Sargent, D. E., Seaman, L. R., Svanborg, N. K., Stanton, R. H., TelIone, C. 1., and Van Goethem, D. L. (1999). Regulating and assessing risks of cholinesteraseinhibiting pesticides: Divergent approaches and interpretations. J. Toxieol. Environ. Health. B 2, 105-160. Clowes, H. M. (1993). "Metam Sodium: In Vitro Absorption through Rat and Human Skin." Unpublished study (Rep. CTL1P14118) conducted by Zeneca Central Toxicology Laboratory, Alderley Park, Macclesfield, Cheshire, UK. Submitted by Metam-sodium Task Force. Crain, R. C. (1958). Spontaneous tumors in the Rochester strain of the Wistar rat. Amer. J. Pathol. 34,311-335. Deerberg, F., Rapp, K., Rehm, S., and Pitterman, W. (1980). Genetic and environmental influences on lifespan and diseases in Han : Wistar rats. Meeh. Ageing Devel. 14,333-343. Hawkins, D. B., Elsom, L. F., and Girkin, G. (1987). "The Biokinetics and Metabolism of 14C-Metam in the Rat." Unpublished study conducted by Huntingdon Research Centre, UK. Submitted by BASF Corporation, Research Triangle Park, Ne. HelIwig, J. (1987). "Report on the Study of the Prenatal Toxicity of Metam-Sodium (Aqueous Solution) in Rabbits after Oral Administration (Gavage)." Unpublished study (Project 38R0232/8579) conducted by BASF AktiengeselIschaft, Federal Republic of Germany. Submitted by BASF Corporation Chemicals Division, Parsippany, NJ. HelIwig, J., and Hildebrand, B. (1987). "Report on the Study of the Prenatal Toxicity of Metam-Sodium in Rats after Oral Administration (Gavage)." Unpublished study (Rep. 87/0128) conducted by BASF AktiengeselIschaft, West Germany. Submitted by BASF Corporation, Research Triangle Park, Ne. Hodge, M. C. E. (1993). "Metam Sodium: Developmental Toxicity Study in the Rabbit." Unpublished study (Rep. CTLIP14035) conducted by Zeneca Central Toxicology Laboratory, Alderley Park, Macclesfield, Cheshire, UK. Submitted by Metam-sodium Task Force. Homer, S. A. (1994). "Metam-Sodium: Two Year Drinking Study in Mice." Unpublished study (Rep. CTL1P14095) conducted by Zeneca Central Toxicology Laboratory, Cheshire, UK. Submitted by Metam-sodium Task Force. JMPR (Joint Meeting on Pesticide Registrations, World Health Organization) (1995). "Pesticide Residues in Food-1995." FAO Plant Production and Protection Paper 133, p. 4. Jowa, L. (1998). Metam: Animal toxicology and human risk assessment. In "Toxicology and Risk Assessment: Principles, Methods, and Applications" (A. M. Fan and L. W. Chang, eds.), p. 619. Dekker, New York. Klaassen, e. D. (1986). Chapter 2: Principles of toxicology. In "Cassarett and DoulI's Toxicology: The Basic Science of Poisons" (e. D. Klaassen, M. O. Amdur, and J. DoulI, eds.), pp. 11-32. McGraw-Hill, New York.
References
Kroes, R., Garbis-Berkvens, J. M., de Vries, T., and van Nesselrooy, H. J. (1981). Histopathological profile of a Wistar rat stock including a survey of the literature. J. Gerontal. 36, 259-279. Lamb, 1. C. (1993). "An Acute Neurotoxicity Study of Metam-Sodium in Rats (Definitive)." Unpublished study (Study WIL-188009) conducted by WIL Research Laboratories, Inc., Ashland, OH. Submitted by Metam-sodium Task Force, Los Angeles, CA. Liggett, M. P., and McRae, L. A. (1991). "Skin Irritation to Rabbits with Metam-Sodium." Unpublished study (Study 90997DlUCB 368/SE) conducted by Huntingdon Research Centre, Ltd., UK. Submitted by UCB Chemicals Corporation, Norfolk, VA. Mackay (1996). Merck (1989). Metham sodium. In "The Merck Index" (S. Budavari, M. L. O'Neil, A. Smith, and P. E. Heckelman, eds.), I I th ed., p. 937. Merck, Rahway, NJ. Milburn, G. M. (1993). "Metam Sodium: Multigeneration Study in the Rat." Unpublished study (Rep. CTL1P/3788) conducted by Zeneca Central Toxicology Laboratory, AlderIey Park, MaccJesfield, Cheshire, UK. Padgett, E. L., Bames, D. R, and Prnett, S. B. (1992). Disparate effects of representative dithiocarbamates on selected immunological parameters in viva and ceIl survival in vitro in female B6C3Fl mice. J. Taxical. Environ. Health 37,559-571. ParceIl, R 1., and Denton, S. M. (1991). "Delayed Contact Hypersensitivity in the Albino Guinea Pig." Unpublished report (Rep. 901002DIUCB370/SS) conducted by Huntingdon Research Centre, Ltd., UK. Submitted by UCB Chemicals Corporation, Norfolk, VA. Pruett, S. B., Bames, D. B., Han, Y. c., and Munson, A. E. (1992). Immunotoxicological characteristics of sodium methyldithiocarbamate. Fund. Appl. Taxical. 18,40-47. Rattray, N. J. (1994). "Metam-Sodium: Two Year Drinking Study in Rats." Unpublished study (Project PR0838) conducted by Zeneca Central Toxicology Laboratory, Cheshire, UK. Submitted by Metam-sodium Task Force. Rehm, S., Deerberg, E, and Rapp, K. G. (1984). A comparison of life-span and spontaneous tumor incidence of male and female Han: WIST virgin and retired breeder rats. Lab. Anim. Sci. 34, 458-464. Stewart, E P. (1992). "Metam-Sodium: In VIva Percutaneous Absorption Study in the Rat." Unpublished study (Report 7268-38/142) conducted by Hazleton UK, Harrogate, North Yorkshire, UK. Thomassen, R. W. (1998). "Review of the 2-Year Drinking Water Study with Metam-Sodium in the Wistar Rat." Unpublished report submitted to the California Office of Environmental Health Hazard Assessment. Reviewed with California Office of Environmental Health Hazard Assessment at a conference on July 29, 1998. Tinston, D. J. (1993). "Metam Sodium: Developmental Toxicity Study in the Rat." Unpublished study (Report CTL1P/4052) conducted by Zeneca Central Toxicology Laboratory, AlderIey Park, MaccJesfield, Cheshire, UK.
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United States Environmental Protection Agency (U.S. EPA) (1991). "MetamSodium-Review of Two Developmental Toxicity Studies in Rats and Rabbits Submitted by the Registrant." Memorandum from Y. M. Ioannou to S. Lewis. United States Environmental Protection Agency (U.S. EPA) (1992). "MetamSodium-Review of a 90-Day Study in Mice." Memorandum from Y. M. Ioannou to S. Lewis. United States Environmental Protection Agency (U.S. EPA) (1993). "Sodium N-Methyldithiocarbamate (Metam-Sodium)." Memorandum from T. E McMahon to A. Mehta. United States Environmental Protection Agency (U.S. EPA) (1994a). "Worker and ResidentiallBystander Risk Assessment of Metam-sodium During Soil Applications." Memorandum from A. Mehta to J. EIIenberger and J. Housenger. United States Environmental Protection Agency (U.S. EPA) (1994b). "MetamSodium: Review of a Chronic Toxicity/Carcinogenicity Study in Rats and Chronic Toxicity Study in Dogs Submitted by the Registrant." Memorandum from T. E McMahon to T. Myers. United States Environmental Protection Agency (U.S. EPA) (1994c). "MetamSodium: Review of a Mouse Carcinogenicity Study Submitted under FIFRA Section 6(a)(2) by the Registrant." Memorandum from T. E McMahon to T. Myers. United States Environmental Protection Agency (U.S. EPA) (1994d). "MetamSodium: Review of an Acute Neurotoxicity Study Submitted by the Registrant." Memorandum from T. E McMahon to T. Myers. United States Environmental Protection Agency (U.S. EPA) (1994e). "MetamSodium: Review of a Subchronic Neurotoxicity Study in Rats." Memorandum from T. E McMahon to T. Myers. United States Environmental Protection Agency (U.S. EPA) (1995). "Carcinogenicity Peer Review of Metam-Sodium." Memorandum from T. E McMahon and E. Rinde to L. Cole and T. Myers. United States Environmental Protection Agency (U.S. EPA) (1997). "Case Study-Methyl Bromide Alternative: Metam-Sodium as an Alternative to Methyl Bromide for Fruit and Vegetable Production." US EPA document posted at http://earthl.epa.gov/ozone/mbr/metams.htrn. Whiles, A. J. (1991). "Metam-Sodium: 90-Day Drinking Water Study in Mice with a 28-Day Interim Kill." Unpublished Report (Rep. CTL1P/3185) conducted at ICI Central Toxicology Laboratory, AlderIey Park, Macclesfield, Cheshire, UK. World Health Organization (WHO) (1990). "Environmental Health Criteria 104. Principles for the Toxicological Assessment of Pesticide Residues in Food." International Program on Chemical Safety, World Health Organization, Geneva.
CHAPTER
88 Sulfuryl Fluoride Kenneth D. Nitschke The Dow Chemical Company
David L. Eisenbrandt Dow AgroSciences LLC
88.1 CHEMISTRY AND FORMULATIONS Sulfuryl fluoride, S02F2, is manufactured and sold under the trade name Vikane* Gas Fumigant. The CAS registry no. is 2699-79-S. The molecular weight of sulfuryl fluoride is 102.07 and it is a colorless, odorless gas with a melting point of -135.S2°C and a boiling point of -55.3S°C. The vapor pressure is 13 x 103 Torr at 25°C. The solubility in water is 0.75 g/kg at 25°C. Sulfuryl fluoride is of low solubility in most organic solvents but is miscible with methyl bromide. It is stable and noncorrosive by DOT definitions. Sulfuryl fluoride is not hydrolyzed by water but is hydrolyzed by NaOH solution.
88.3 HAZARD IDENTIFICATION-TOXICITY TO LABORATORY ANIMALS (PRE-1980) 88.3.1 ACUTE TOXICITY
88.2 USES Since first marketed in 1961 as Vikane gas fumigant, sulfuryl fluoride has been used to fumigate over one million buildings, including houses, museums, historical landmarks, rare book libraries, government archives, and scientific and medical research laboratories (Dow AgroSciences, 1997). Initial concentrations in fumigated structures are typically 2000-4000 ppm although other concentrations may be used depending upon the target pest to be controlled, temperature, and the length of the exposure period. Su1furyl fluoride can be used to control a wide variety of household pests, including cockroaches, rodents, clothes moths, bedbugs, and carpet beetles. The activity of sulfury1 fluoride is dependent on the concentration reaching the target pest and the duration of exposure. Insect eggs require a higher dosage of sulfuryl fluoride compared to postembryonic life stages. Since the immature stages of some insects, such as termites and ants, cannot survive without adult care, dosages substantially less than required for egg control are effective. Prior to fumigation with sulfuryl fluoride, a small amount of chloropicrin is introduced into the structure to warn people and animals that the structure is being fumigated (Dow AgroSciences, 1997). Chloropicrin has a noticeable disagreeable pungent odor at less than 1 ppm and causes irritation Handbook of Pesticide Toxicology Volume 2. Agents
of the eyes, tears, and noticeable discomfort. Since chloropicrin diffuses from structures more slowly than sulfuryl fluoride, building occupants may experience eye irritation immediately after entering previously fumigated buildings. The ACGIH threshold limit value and OSHA permissible exposure level for sulfuryl fluoride are 5 ppm TWA, 10 ppm STEL (ACGIH, 1995).
Liquid sulfuryl fluoride directly contacting skin can result in frostbite (Torkelson et aI., 1966). Animals exposed to lethal doses of sulfuryl fluoride had tremors and convulsions. The convulsions were characterized by stiffening of the animal to a very rigid position and then toppling over backward. Excessive salivation, loss of bladder control, and chromodacryorrhea were observed also. A 1 hour LC50 in male and female rats was 3020 and 3730 ppm, respectively (Vernot et aI., 1977). 88.3.2 SUB CHRONIC TOXICITY Lung, kidney, and liver microscopic changes were observed within 3 weeks in rats, guinea pigs, and rabbits exposed to 400 ppm sulfuryl fluoride for 7 hours/day, 5 days/week (Torkelson, 1959). In studies at 20, 50, 100, 150, and 200 ppm for up to 6 months, lung pathology was the most significant effect observed. In addition, slight central lobular degeneration and vacuolation of the liver and slight cloudy swelling of the tubular epithelium of the kidney were observed. Increased levels of fluoride in the blood, lungs, bone, and teeth as well as fluorosis of the teeth were observed. Groups of male and female rats (lO/sex/dose) were maintained for 66 days on diets fumigated with 0 (control), 2, 10, 100, or 200 pounds/l 000 cu. ft. of sulfuryl fluoride (Lockwood,
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CHAPTER 88
Sulfuryl Fluoride
1958). The animals were weighed twice weekly for the first 28 days and then once a week thereafter and food consumption was recorded for the first month. They were observed frequently for gross changes in appearance or behavior and the teeth were examined for visual evidence of fluorosis. Samples of urine were obtained from all male rats for fluoride analysis. Terminal hematological values were obtained from five female rats in the 0, 2, and 10 pounds/lOOO cu. ft. levels and from two male rats at each dietary level. Animals were autopsied at study termination and the lungs, heart, liver, kidneys, spleen, and testes were weighed. Portions of these organs as well as pancreas and adrenals were preserved and prepared for histological examination. Two additional rats of each sex were included in each dose group for collection of samples at 30 days of blood, urine (male), kidney, lung, liver, and bones and subsequent analysis for fluorides. Control diets averaged 36 ppm of fluoride while diets fumigated with 2, 10, 100, or 200 pounds/lOOO cu. ft. sulfuryl fluoride contained 55, 89, 386, or 740 ppm fluoride, respectively. Male and female rats tolerated a diet fumigated with sulfuryl fluoride at the rate of 2 pounds/lOOO cu. ft. with no evidence of adverse effects, although fluoride content of bone was increased somewhat. Diets fumigated with higher levels of sulfuryl fluoride resulted in retardation of growth and evidence of fluorosis in the teeth. The severity of the effects was directly proportional to the fluoride content of the diet. Fluoride analysis of urine and bones showed increased amounts of fluorides in proportion to the amount of sulfuryl fluoride exposure to the diet (in males, concentration of fluoride in bone was 260, 408, 413, 1615, and 1920 ppm and in urine was 9.9, 1l.9, 13.3,85.6, and 174.4 ppm for diets containing 36,55,89,386, or 740 ppm fluoride, respectively); analysis of blood, kidney, lung, and liver for fluorides was inconsistent with dosage levels.
88.4 TOXICITY TO LABORATORY ANIMALS (POST-1980) 88.4.1 ACUTE TOXICITY
The 4 hour LC50 in male and female Fischer 344 rats was 1122 and 991 ppm, respectively (Miller et al., 1980). Gross pathologic changes were observed primarily in the upper and lower respiratory tract. Histopathologic examination of animals exposed to 1200-1250 ppm revealed liver and kidney and possibly lung, heart, and spleen effects. Animals that survived for two weeks following exposure to 1200-1250 ppm exhibited regenerative responses in the kidney. The highest concentration at which all animals survived was 450 ppm for males and 790 ppm for females. B6C3Fl mice were exposed to 400, 600, or 1000 ppm for 4 hours (Nitschke and Lomax, 1989). All animals died within 90 minutes after termination of the exposure to 1000 ppm and within 6 days following exposure to 600 ppm. Body tremors were observed in several female mice shortly after exposure to 600 ppm and animals surviving after the exposure period were
lethargic prior to death. There were no clinically visible effects noted in mice exposed to 400 ppm. In the B6C3Fl mouse, the 4-hour LC50 was between 400 and 600 ppm for both males and females. CD-l mice were exposed to 600, 700, or 800 ppm for 4 hours to determine the LC50 (Nitschke and Quast, 1990). Effects noted in the CD-l mouse were similar to those observed in the B6C3Fl mouse, but they occurred at slightly higher concentrations. The 4-hour LC50 was 660 and 642 ppm for males and females, respectively. Acute dermal exposure to sulfuryl fluoride was evaluated in rats exposed to 1000 or 9600 ppm for 4 hours (Bradley et al., 1990). These concentrations are 1 and 10 times greater than the 4-hour LC50 in rats identified by Miller et al. (1980). The exposures occurred in a Rochester-type chamber equipped with a door modified to allow the head of rats to protrude outside of the chamber. In the door, an elastic dental dam surrounded the neck of the rats and served as a barrier between the chamber air containing sulfuryl fluoride and the breathing air outside of the chamber. In addition, the backs of the rats were shaved prior to exposure to maximize dermal exposure. The only clinical effects during the exposure were chromodacryorrhea and fecal soiling. The incidence of chromodacryorrhea and fecal soiling was comparable between the two exposure groups and the effects were considered to be stress related, due to the method of restraint used during the exposure. There was no evidence of body tremors in these rats. Histopathologic examination of the skin and brain revealed no treatment-related lesions. Thus the dermal route does not appear to play a significant role in the toxicity of sulfuryl fluoride. 88.4.2 NEUROTOXICITY
Rats were exposed to 0, 100, or 300 ppm sulfuryl fluoride for 6 hours/day for 2 consecutive days (Albee et al., 1993). A functional observational battery, grip performance, landing foot splay, motor activity, and a battery of electrodiagnostic tests, including flash evoked potentials, somatosensory evoked potentials, and auditory brainstem responses, were conducted pre- and postexposure. Except for motor activity data, postexposure data were collected within 5 hours after the second exposure. Motor activity data were collected 18 hours after exposure. There were no exposure-related effects in any of the parameters. 88.4.3 TIME TO INCAPACITATION
In an effort to understand the mode of action of sulfuryl fluoride, rats were exposed to 4000, 10,000,20,000, or 40,000 ppm to determine the time to incapacitation (Nitschke et al., 1986). Rats were exposed to sulfuryl fluoride in a 14 liter cylindrical chamber equipped with a motor-driven activity wheel. Animals were forced to walk on the activity wheel for designated intervals during the exposure. Exposures were terminated when incapacitation or convulsions occurred. All rats either
88.4 Toxicity to Laboratory Animals (Post -1980)
died or were moribund within 3 hours following the end of the exposure. At the two highest concentrations, 20,000 ~nd 40,000 ppm, rats were incapacitated within 12 minutes and dIed within 10 minutes after terminating exposure. At 10,000 ppm, rats were incapacitated after 16 minutes and, at 4000 ppm, rats were incapacitated after 40 minutes. At the lowest concentration, the mean survival time was 2.5 hours after incapacitation occurred. Animals exposed to 10,000 ppm and higher appeared to be slightly cyanotic shortly after exposure occurred. The bluish skin discoloration disappeared within 10 minutes after purging the chamber with room air following exposure to 10,000 ppm. The skin discoloration did not appear to be reversible at higher concentrations. The cause of death at all concentrations appeared to be cardiovascular failure and acute death. Pulmonary congestion increased in severity as the concentration of the test chemical was increased in the atmosphere and at the two highest exposure concentrations, the pulmonary lesions appeared to contribute significantly to the death of the animals. At the two lowest concentrations, the pulmonary effects were not severe enough to be the major factor in the death of these animals (Nitschke et aI., 1986). The lungs of rats exposed to 4000 or 20,000 ppm sulfuryl fluoride until incapacitation occurred were examined by light and electron microscopy (Eisenbrandt et aI., 1987). Histopathologic examination of the lungs from rats exposed to 4000 ppm revealed minimal congestion, edema, and hemorrhage. Rats exposed to 20,000 ppm had more severe changes. Swelling and focal disruptions in alveolar epithelial cells were observed by electron microscopic examination of the lung. In addition, multifocal destruction of the alveolar wall with associated thrombosis was present in the lungs of rats exposed to 20,000 ppm. Both concentrations increased permeability of the alveolar wall as evidenced by the presence of edema, red blood cells, and fibrin in alveoli and interstitial spaces. In another study to understand the mode of action of sulfuryl fluoride, rats were exposed to 4000 or 10,000ppm sulfuryl fluoride to determine the effect on respiration (Landry and Streeter, 1983). Respiratory frequency as well as tidal and minute volume were measured at 1 minute intervals for 10 minutes prior to exposure and during a 20 minute exposure. At 4000 ppm, there was a rapid initial increase in mean respiratory frequency and a decrease in mean tidal volume and mean minute volume relative to pre-exposure values. These respiratory parameters peaked after 2 minutes of exposure; there was a 39% increase in frequency, a 40% decrease in tidal volume, and a 23% decrease in minute volume. After approximately 10 minutes of exposure, frequency and tidal volume were near pre-exposure levels. These effects were considered to be insufficient to have resulted in mortality. Body temperature, heart rate, blood pressure, and electroencephalogram were monitored in rats exposed to 4000 or 20,000 ppm sulfuryl fluoride until rats expired (Gorzinski and Streeter, 1985). At 4000 and 20,000 ppm animals survived for 79 and 14 minutes, respectively. A decrease in heart rate, a gradual increase in blood pressure, decreased respiration, occasional spiking and high frequency loss in the EEG, and power loss in
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the EEG were observed at both levels. All physiological parameters ceased to function at about the same time when the animals died regardless of the concentration of sulfuryl fluoride and did not help explain the cause of death in these animals. 88.4.4 THERAPEUTIC/AMELIORATION
OF TOXICITY Therapeutic treatment with calcium gluconate was evaluated because sulfuryl fluoride is extensively dehalogenated in termites (Meikle et aI., 1963) and the similarity of effects in the mammalian incapacitation studies suggested fluoride toxicity (Nitschke et al., 1986). Rats were exposed to 4000 or 10,000 ppm sulfuryl fluoride for 45 or 16 minutes (sufficient to result in 100% lethality within 3 hours), respectively, or were treated i.p. with calcium gluconate prior to and after exposure to sulfuryl fluoride. Animals which were still alive 3 days after exposure to sulfuryl fluoride were considered to have survived. In separate groups of rats, serum fluoride levels were determined in rats exposed to sulfuryl fluoride alone or pretreated with calcium gluconate. Four of five rats pretreated with calcium gluconate prior to exposure to 4000 ppm sulfuryl fluoride survived 3 days after dosing. The other rat died 90 minutes after exposure. Treatment with calcium gluconate after exposure to sulfuryl fluoride was not effective. While survival was increased in rats pretreated with calcium gluconate, the surviving animals were extremely debilitated with no apparent protection from convulsions. Animals pretreated with calcium gluconate did not survive exposure to 10,000 ppm sulfuryl fluoride. Administration of calcium gluconate resulted in approximately 10% increase in serum calcium levels and did not appear to affect serum fluoride or magnesium levels (Nitschke et al., 1986). Since the cause of death in rats pre- or postexposure treated with calcium gluconate appeared to be due to convulsions, rats were pretreated with one of three anticonvulsants (phenobarbital, diazepam, or diphenylhydantoin) prior to exposure to 4000 ppm sulfuryl fluoride or were treated postexposure with phenobarbital and diazepam (Nitschke et at., 1986). These three anticonvulsants were selected due to their ready availability and different mechanisms of action. Phenobarbital was the most effective anticonvulsant followed by diazepam with five of five and four of five pretreated with phenobarbital and diazepam surviving exposure to 4000 ppm for 50 minutes. Diphenylhydantoin accentuated the adverse effects of sulfuryl fluoride. At higher concentrations of sulfuryl fluoride, 10,000 ppm phenobarbital was also ineffective. 88.4.5 REPEATED EXPOSURES
The study design of several repeated exposure studies followed the appropriate EPA Test Guidelines (FIFRA Guideline Nos. 82-1,82-4,83-1,83-2, and 83-5). For example, the study designs for several of these studies are detailed by Eisenbrandt and Nitschke (1989) and Hanley et al. (1989). Briefly, animals
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CHAPTER 88
Sulfuryl Fluoride
were exposed to 99.8% pure sulfuryl fluoride for 6 hours/day, 5 days/week for various time intervals. The analytical concentrations of sulfuryl fluoride in the air were measured by infrared spectrophotometry and analytical and nominal concentrations were in very close agreement. In general, animals were observed after exposure for changes in appearance or behavior. Animals were weighed periodically. Blood samples were obtained for hematological and clinical chemistry determinations. Urine was collected from rats and various parameters were measured. Animals were sacrificed the day following the last exposure to sulfuryl fluoride. Terminal body weights and selected organ weights were recorded. All animals were examined for gross pathological alterations by a veterinary pathologist. Animals that died or were moribund prior to the scheduled sacrifice were necropsied as soon as possible. An extensive set of tissues was collected and processed for light microscopy by conventional techniques and was stained with hematoxylin and eosin. In some cases special stains were used for selected tissues.
88.4.6 SUBCHRONIC TOXICITY-MICE Groups of 5 mice/sex were exposed to 0, 30, 100, or 300 ppm sulfuryl fluoride for 6 hours/day, 5 days/week for two weeks (Nitschke and Quast, 1995). All male mice and 4 of 5 female mice exposed to 300 ppm sulfuryl fluoride died during the second week of the study. These animals lost weight and many had tremors. Vacuoles were observed in the cerebrum and/or
Table 88.1 Serum Fluoride Levels in Mice Following I3-Week Exposure to Sulfuryl Fluoride Fluoride (ppm) Concentration S02F2 (ppm) 0
Males
Female
± 0.017 0.112 ± 0.027 0.156 ± 0.019 0.259 ± 0.073*
0.090 ± 0.015
0.107
10 30 100
± 0.019 ± 0.020* 0.233 ± 0.022* 0.088 0.132
*Statistically different from control mean by Dunnett's test, alpha
= 0.05.
N=4.
medulla of 8 of 10 mice exposed to 300 ppm and varied from very slight to moderate in severity. Four male and 2 female mice exposed to 100 ppm had very slight vacuoles in the cerebrum. The no-observed-effect-level (NOEL) was 30 ppm. Groups of 14 mice/sex were exposed to 10, 30, or 100 ppm for 6 hours/day, 5 days/week for 13 weeks (Nitschke and Quast, 1993). Four animals/sex/exposure level were used to measure serum fluoride levels. In addition, tissues of these 4 animals/sex/exposure level were perfused with glutaraldehyde/formaldehyde fixative and neural tissues were examined histopathologically. Standard parameters were evaluated in the main group of 10 mice/sex/concentration. At the highest concentration, 100 ppm, there was approximately a 10% body weight decrease in male and female mice from control values. Except for elevated serum fluoride levels (Table 88.1) which
Table 88.2 Histopathologic Observations in Brains of Males Exposed to Various Concentrations of Sulfuryl Fluoride for 13 Weeks Concentration Species
Control
Low
Middle
High
Mice 10
10
10
10
vacuolation caudate putamen very slight
0
0
0
7
vacuolation caudate putamen slight
0
0
0
2
vacuolation external capsule very slight
0
0
0
7
vacuolation external capsule slight
0
0
0
2
10
10
10
10
0
0
0
10
Brain--cerebrum-number of tissues examined
Rats Brain--cerebrum-number of tissues examined vacuolation cerebrum slight Rabbits 7
7
7
7
malacia severe
0
0
0
3
vacuolation very slight
0
0
0
3
4
4
4
4
0
0
0
Brain--cerebrum-number of tissues examined
Dogs Brain-midbrain-number of tissues examined vacuolation very slight
Animals were exposed to sulfuryl fluoride for 6 hrs/day, 5 days/week for 13 weeks. The low, middle, and high concentrations corresponded to targeted concentrations of 10, 30, and 100 ppm in mice, 30, 100, and 300 ppm in rats, 30, 100, and 300 ppm in rabbits, and 30, 100, and 200 ppm in dogs, respectively.
88.4 Toxicity to Laboratory Animals (Post-l980)
followed a dose-response relationship, there were no exposurerelated effects on clinical chemistry, hematology, urinalysis, organ weight, or gross pathology. Histopathologic examination of mice exposed to 100 ppm sulfuryl fluoride for 13 weeks revealed effects in the brain and thyroid gland (Tables 88.2 and 88.3). In the cerebrum, micro vacuolation in male and female mice was observed in the external capsule and the caudate putamen. The effect was very slight to slight in severity and was bilaterally symmetrical. Microvacuoles were also observed in the region of the thalamuslhypothalamus of these animals. In this region, the microvacuoles usually extended from the external capsule and involved the adjacent amygdaloid region. There were no recognizable inflammatory or degenerative changes associated with the microvacuoles. The single male mouse from the 100 ppm group without microvacuoles in the brain died during the course of the study due to accidental trauma. Microscopic changes in the thyroid gland were characterized by very slight hypertrophy of the follicular epithelial cells associated with a decrease in the amount of colloid present. The NOEL was 30 ppm in mice exposed to sulfuryl fluoride for 6 hours/day, 5 days/week for 13 weeks.
88.4.7 SUBCHRONIC TOXICITY-RATS Groups of five animals of each sex were exposed to 100, 300, or 600 ppm sulfuryl fluoride for 6 hr/day, 5 days/week for
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nine exposures in two weeks (Eisenbrandt and Nitschke, 1989). Nine of 10 rats in the 600 ppm exposure group were moribund and/or died between the second and sixth exposures. One female rat exposed to 600 ppm survived the nine exposures in the two week period. These animals had severe weight loss with body weights less than 70% of control values by the fifth exposure. Severe kidney lesions were observed in all rats exposed to 600 ppm. The papillary epithelium was necrotic over the tip of the papillae and the remainder of the epithelium was moderately hyperplastic. Subacute inflammation was associated with the necrosis and collecting ducts throughout the kidneys were dilated as a result of obstruction of the papillae. There was degeneration and necrosis of epithelial cells of the collecting ducts with regeneration of surviving epithelial cells. Several rats had mineralization at the junction of the inner and outer medulla which was accompanied by chronic interstitial inflammation in some animals. Degenerative and regenerative changes in the proximal tubules were also observed. Minimal kidney changes were observed in rats exposed to 300 ppm; there were no exposure-related kidney lesions observed in rats exposed to 100ppm. Severe effects were observed in the upper respiratory tract and, to a lesser extent, larynx, trachea, and lungs in the sole surviving female rat exposed to 600 ppm. Also, severe, diffuse inflammation was noted in the nasal mucosa and was accompanied by multifocal ulceration of the mucosa and slight bronchioalveolar inflammation was present in the lungs. The
Table 88.3 Histopathologic Observations in Brains of Females Exposed to Various Concentrations of Sulfuryl Fluoride for 13 Weeks Concentration Species
Control
Low
Middle
High
10
10
10
10
o o o o
o o o o
o
o o o
3 5 6 4
10
10
10
10
o
o
o
10
Mice Brain---cerebrum-number of tissues examined vacuolation caudate putamen very slight vacuolation caudate putamen slight vacuolation external capsule very slight vacuolation external capsule slight Rats Brain---cerebrum-number of tissues examined vacuolation cerebrum slight Rabbits Brain---cerebrum-number of tissues examined
7
7
7
7
malacia severe
o
o
vacuolation very slight
o o o
o
o o
3
o
o
vacuolation slight vacuolation moderate
o
I 2
o
Dogs Brain-midbrain-number of tissues examined vacuolation very slight Females were exposed to same concentrations as males in Table 88.2.
4
4
4
o
o
o
4
1886
CHAPTER 88
Sulfuryl Fluoride
NOEL was 100 ppm in rats exposed to sulfuryl fluoride for two weeks. Groups of 10 rats/sex were exposed to 0,30, 100, or 300 ppm for 6 hr/day, 5 days/week for 13 weeks (Eisenbrandt and Nitschke, 1989). Effects attributed to sulfuryl fluoride toxicity included mottled teeth in rats exposed to lOO or 300 ppm, decreased specific gravity of the urine, and histopathological changes in the respiratory tract, brain, and kidneys of rats exposed to 300 ppm. The respiratory tract effects consisted of very slight to severe inflammation in the nasal tissue with mucopurulent exudate in the nasal passages in the more severe cases. The more extensive inflammation was accompanied by degeneration and reactive changes in the mucosa. Slight subpleural histiocytosis also was observed in the lungs of rats exposed to 300 ppm sulfuryl fluoride. In the brain, minimal vacuolation in the area of the caudateputamen nuclei was observed in rats exposed to 300 ppm and was more prominent in the white fiber tracts of the internal capsule than in the adjacent neuropil (Tables 88.2 and 88.3). Special stains of the brain with LFB-PAS or Sevier Munger stain did not reveal any additional effects. Very slight hyperplasia of the renal collecting ducts was most apparent in the outer portion of the inner zone of the medulla of most female rats exposed to 300 ppm. As a separate part of the above mentioned 13 week study, groups of 7 rats/sex were exposed to 0, 30, 100, or 300 ppm for the same time period for the specific purpose of evaluating neurological function (Mattsson et aI., 1988). After 13 weeks of exposure to sulfuryl fluoride, hindlimb grip strength, observation battery, visual evoked response, cortical flicker fusion, auditory brainstem response to tone pips, auditory brain stem response to clicks, cerebellar evoked response, somatosensory evoked response, and caudal nerve action potential were measured approximately 12 hours after the last exposure to sulfuryl fluoride. All but two males and two females from the 0 and 300 ppm exposure groups were subjected to a gross pathologic examination at the end of the 13 week study. The 2 animals/sex from the 0 and 300 ppm exposure group were evaluated with an auditory brainstem response approximately 8 weeks after the last exposure to sulfuryl fluoride and then sacrificed. Hindlimb grip strength was normal for all rats, but evoked responses were clearly altered at 300 ppm and slightly altered at 100 ppm. The principal effect was decrease in flicker fusion and a slowing of all waveforms at 300 ppm, and a slowing of the visual evoked response and the somatosensory evoked response of female rats at 100 ppm. Histopathologic changes in the brain consisted of vacuoles in the white fiber tracts of the caudateputamen. No necrosis or neuronal destruction were noted. The two rats exposed to 300 ppm in the recovery group had normal auditory brainstem responses and normal brain histopathology. The fact that all of the evoked responses in rats exposed to 300 ppm for 13 weeks were affected suggested a widespread functional eNS effect. The electrophysiologic slowing of the evoked responses was felt to be due to some mechanism other than the minor vacuolization observed in the caudate-putamen. The NOEL was 30 ppm in rats exposed to sulfuryl fluo-
ride for 13 weeks in the standard subchronic and neurological studies.
88.4.8 SUBCHRONIC TOXICITY-RABBITS Groups of three rabbits of each sex were exposed to 100, 300, or 600 ppm sulfuryl fluoride for 6 hr/day, 5 days/week for nine exposures in two weeks (Eisenbrandt and Nitschke, 1989). All rabbits exposed to 600 ppm were hyperactive and one animal had a convulsion which resulted in a fractured tibia. A second rabbit had a fractured vertebra which may have been the result of a convulsion. Treatment-related malacia (necrosis) was present in the cerebrum of all rabbits exposed to 600 ppm and one male and one female rabbit exposed to 300 ppm. Reactive gliosis and demyelination accompanied the malacia. Rabbits exposed to 300 or 600 ppm also had vacuoles in the globus pallidus and putamen as well as the external and internal capsules of the brain. Moderate, subacute, to chronic inflammation of nasal tissues with mucopurulent exudate in the nasal cavities was observed in most rabbits exposed to 300 or 600 ppm sulfuryl fluoride. The inflammation was probably from irritation of the nasal mucosa due to the test material. At the higher concentration, acute inflammation was observed in the trachea, bronchi, and bronchioles of some rabbits and may have been treatment-related. The NOEL was 100 ppm in rabbits exposed to sulfuryl fluoride for two weeks. In a 13 week study, groups of 7 rabbits/sex were exposed initially to 30, 100, or 600 ppm for 6 hours/day, 5 days/week (Eisenbrandt and Nitschke, 1989). After 2 weeks exposure to 600 ppm, the target concentration was reduced to 300 ppm, which resulted in an average concentration over the 13 week period of 337 ppm. The exposure concentration was reduced from 600 to 300 ppm due to convulsions observed in one male and one female. A second female rabbit exposed to 600 ppm was euthanized after eight exposures due to a fractured vertebra. No clinically visible effects were noted in rabbits exposed to 300 ppm or lower concentrations. Except for elevated serum fluoride levels which followed a dose-response relationship, there were no clinical chemistry, hematology, urinalysis, organ weight, or gross pathologic changes observed. Histopathologic changes were noted in the nasal tissues and brain of rabbits exposed to 300 and in the nasal tissue of one male and in the brain of one female exposed to 100 ppm (Tables 88.2 and 88.3). In the nasal tissues, varying degrees of purulent nasal exudate, olfactory epithelial degeneration, and hyperplasia and hypertrophy of the respiratory epithelium in the nasal turbinates were observed. The brain changes consisted of vacuolation of the white matter at 100 ppm. In rabbits exposed to 300 ppm, malacia of the internal and external capsules, putamen, and globus pallidus were observed. Some animals exposed to 300 ppm had gliosis and/or hypertrophy of vascular endothelial cell in the same area. Special stains of the brain
88.4 Toxicity to Laboratory Animals (Post-1980)
with LFB-PAS or Sevier Munger stain were not remarkable. The NOEL was 30 ppm in rabbits exposed to sulfuryl fluoride for 13 weeks.
88.4.9 SUBCHRONIC TOXICITY-DOGS Groups of one male and one female dog were exposed to target concentrations of 0, 30, 100, or 300 ppm for 6 hr/day 5 day/week for nine exposures (Nitschke and Quast, 1991). At 300 ppm, infrequent intermittent episodes of tremors and tetany were observed in both dogs beginning with the fifth exposure. On test day 9, during the seventh exposure, the tremors and tetany were severe enough that the exposure was terminated after approximately 5.5 hours. Within 30 minutes after terminating the exposure, both dogs appeared to be normal. Similar clinical effects were noted during subsequent exposure periods and were rapidly reversible even during the exposure period. There were no exposure-related clinical effects in dogs exposed to 30 or 100 ppm. The female dog exposed to 300 ppm sulfuryl fluoride lost approximately 500 grams body weight over the nine exposures. Serum fluoride levels of dogs exposed to 100 or 300 ppm were approximately 2-4 fold higher than control values. Serum calcium levels measured shortly after exposure on test days 5 and 9 when the animals appeared to be clinically normal were comparable to control levels. There were no exposure-related hematological, organ weight, or gross pathological effects noted in dogs exposed to concentrations as high as 300 ppm. Minimal microscopic inflammatory changes were observed in the nasal turbinates of the male and female dog and trachea of the female dog exposed to 300 ppm. Although numerous microscopic sections were examined from the cerebral cortex, brainstem, cerebellum, and medulla oblongata, there were no histopathologic changes detected in dogs exposed to 300 ppm. The NOEL was 100 ppm in dogs exposed to sulfuryl fluoride for two weeks. In a 13-week study, groups of four male and four female beagle dogs were exposed to 0, 30, 100, or 200 ppm sulfuryl fluoride for 6 hours/day, 5 days/week for 13 weeks (Nitschke et aI., 1992). One male dog exposed to 200 ppm was laterally recumbent with tetany, tremors, salivation, and incoordination 75 minutes after exposure on test day 19. An hour later the activity of this animal was decreased relative to controls but was otherwise normal. Similar effects were not observed during the remainder of the study. After 13 weeks of exposure to sulfuryl fluoride, the mean body weight values of male and female dogs exposed to 200 ppm were 88 and 96%, respectively, of control values. Mean body weight values of male and female dogs exposed to lower concentrations of sulfuryl fluoride were comparable to control values. There were no exposure-related hematological, clinical chemistry, urinalysis, organ weight, and gross pathological effects observed. Histopathologically, a single small bilaterally symmetrical focal change was noted in the putamen of the midbrain of one male and one female dog exposed to 200 ppm (Tables 88.2 and 88.3). The minimal focal change was characterized by vacuolation, gliosis, perivascular
1887
cuffing, and hypertrophy of endothelial cells, and individual cells showed nuclear pyknosis and karyorrhexis. The focal reaction was slightly more prominent in the male compared to the female. All other microscopic observations were considered to be incidental findings unrelated to exposure to sulfuryl fluoride. The NOEL was 100 ppm in dogs exposed to sulfuryl fluoride for 6 hours/day, 5 days/week for 13 weeks.
88.4.10 CHRONIC TOXICITY-MICE Groups of 50 mice/sex were exposed to measured vapor concentrations of 0, 5, 20, or 80 ppm sulfuryl fluoride for 6 hours/day, 5 days/week for 18 months (Quast et aI., 1993a). Ten additional mice/sex/exposure level were randomly designated as a satellite group for necropsy after 12 months of exposure to evaluate chronic toxicity. During the first year of exposure a slightly earlier onset of mortality was observed in the 80 ppm males; however, mortality at the end of the 18 months of exposure was not statistically identified as increased in any of the sulfuryl fluoride exposed groups of male mice (Fig. 88.1a). The female mortality rate during the first year of exposure was comparable in all groups (Fig. 88.1b). In both males and females exposed to 80 ppm the incidence of exudative rhinitis and aspiration pneumonia accompanied by an impacted esophagus was increased from control values. Body weight of 80 ppm male mice was 10% lower than control values after 6 months and body weights of female mice exposed to 80 ppm were significantly decreased from control values after 1 month exposure (Fig. 88.2a and b, resp.). In general, the effect in the female mice was not as severe as in the male mice during the first 12 months. During the last several months of exposure, the body weight difference between control and high dose female mice ranged from 10-15%. The body weights of male and female mice exposed to 5 or 20 ppm were comparable to control values throughout the study. Clinical chemistry and hematology values of male and female mice exposed to the various concentrations of sulfuryl fluoride were comparable to control values at the 12 month and terminal sacrifice. Although terminal body weight effects were observed in male and female mice exposed to 80 ppm sulfuryl fluoride for 12 or 18 months, the organ weight effects observed were considered to be due to the marked body weight differences and not indicative of target organ toxicity. There were no treatment-related gross pathological effects noted in mice at 12 months. There was a decreased incidence of normally occurring spontaneous gross lesions observed in animals exposed to 80 ppm and included a marked decrease in the incidence of cystic ovaries and cystic endometrial hyperplasia of the uterus in females and a decreased incidence of dilated kidney pelvis. Based upon the necropsy findings, there were no target organs identified in the mice after 12 or 18 months. At the 12 month interim sacrifice, histopathologic examination of male and female mice exposed to various concentrations of sulfuryl fluoride revealed changes in the brain and thyroid gland of animals exposed to 80 ppm only. Essentially all
1888
CHAPTER 88
Sulfuryl Fluoride
100
o ppm --*-- 5ppm
90
- .. - 20ppm 80ppm
0
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70
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(a) 100
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270
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Cb) Figure 88.1
Mortality in (a) male and (b) female mice_
80 ppm exposed mice had very slight or slight microscopic vacuolation of the cerebrum in the region of the external capsule. The caudate-putamen was only affected in one male mouse in
the 80 ppm group. The amygdaloid adjacent to the external capsule was not affected. The vacuolation in the cerebrum was suggestive of edematous change and was not associated with an
88.4 Toxicity to Laboratory Animals (Post-1980)
1889
45 , ------------------------------------------------------------------,
40
~pm
30
··.··5ppm -+-- 20ppm --Et- 80 ppm 25+-~--~--~-r------------------------------------~----L__
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100
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300
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400
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500
55C
Test day
Ca) 40 .---------------------------------------------------------------------~
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25
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~~--~--~~--------~--T_~--~--
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50
100
150
200
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250
_+----------------------------~~
300
350
400
450
500
5E
Test days
Cb) Figure 88.2
Body weights of (a) male and (b) female mice.
inflammatory cell reaction. Very slight hypertrophy of the thyroid follicular epithelial cells was observed. More male mice were affected than female mice. Interestingly, mice exposed to
80 ppm for 12 or 18 months exhibited a lower incidence and severity of brain effects than mice exposed to 100 ppm for 13 weeks (Nitschke and Quast, 1993).
1890
CHAPTER 88
Sulfuryl Fluoride
In the oncogenicity group, target organs were limited to those previously defined after 12 months and consisted of the brain and thyroid gland of mice exposed to SO ppm only. Only a quarter of the mice exposed to SO ppm sulfuryl fluoride had histopathologic changes in the external capsule of the brain. There were no recognizable changes in the caudateputamen or amygdaloid regions of the brain. Thyroid changes in mice exposed to SO ppm were characterized by hypertrophy of follicular epithelial cells. These thyroid changes occurred at a much lower incidence in mice at IS months when compared to 12 months with males having a higher incidence than females. All other microscopic changes were considered to be unrelated to sulfuryl fluoride exposure. There was no increase in the incidence of any tumor in male or female mice exposed to concentrations as high as SO ppm sulfuryl fluoride.
88.4.11 CHRONIC TOXICITY-RATS Groups of 50 male and 50 female rats were exposed to 0, 5, 20, or SO ppm sulfuryl fluoride for 6 hr/day, 5 days/week for two years (Quast et aI., 1993c). Fifteen additional rats/sex/exposure level were randomly designated at the beginning of the study as a satellite group for assessment of general toxicity and neurotoxicity (functional observational battery, motor activity, and perfusion-fixed histopathology of nervous system tissues) following one year of exposure. Mortality of rats exposed to sulfuryl fluoride through the first 16 months of the study was similar to the control group in both male and female rats (Fig. SS.3a and b). After 16 months of exposure to SO ppm, mortality of both males and female rats was increased from control values. The mortality rate of female rats exposed to 5 or 20 ppm was lower than control values from 20 months until the end of the study. Slight body weight effects were observed in female rats exposed to SO ppm throughout the first year; body weights of male and female rats exposed to SO ppm became progressively more severe after one year (Fig. SS.4a and b). There were no consistent effects observed during the first year in urinary specific gravity values of males or females exposed to sulfuryl fluoride. However, urinary specific gravity values of male and female rats exposed to SO ppm for 19 or 21 months were significantly decreased from control values. Several clinical chemistry parameters commonly associated with kidney toxicity were affected in rats exposed to SO ppm for 19 and 21 months. These included increased urea nitrogen, cholesterol, triglycerides, creatinine, and phosphorus and decreased total protein, albumin, and chloride. In addition, albumin levels of male rats exposed to SO ppm for 12 months were also affected. At the 12 month sacrifice, increases in the relative kidney and liver weights of male rats exposed to SO ppm were the only organ weight differences noted in rats exposed to sulfuryl fluoride. Histopathologic changes were noted in the kidneys, lungs, and teeth of rats exposed to SO ppm sulfuryl fluoride for 12 months. A very slight to slight degree of chronic progressive
glomerulonephropathy was noted in males and females with the effects observed in females generally less severely affected (Table SS.2). In the lungs, very slight to slight aggregates of alveolar macrophages were noted in rats. These effects were very minimal and were not considered to significantly impair pulmonary function. Very slight to slight dental fluorosis involving the upper incisor teeth was observed also. The molars of these animals were unaffected. There were no exposure related histopathologic changes noted in rats exposed to 5 or 20ppm. As in rats exposed to sulfuryl fluoride for 12 months, gross and histopathologic changes were noted in the kidneys, lungs, and teeth of rats exposed to SO ppm sulfuryl fluoride for a lifetime. While there was no apparent difference in the gross and histopathologic changes noted in the lungs and teeth of rats exposed for 12 or 24 months, the kidney changes had progressed from very slight or slight to severe or very severe chronic progressive glomerulonephropathy (Table SS.2). These kidney changes have been commonly observed in control animals from lifetime studies, however, the incidence rate was much higher in the high exposure animals. Along with the kidney changes, secondary changes, such as hyperparathyroidism and mineralization of many tissues were observed. These changes have been described previously by Boorman et al. (1990) and Mohr et al. (1992). Except for a very slight fluorosis of the teeth of male rats exposed to 20 ppm, there were no effects observed in male or female rats exposed to 5 or 20 ppm. There was no increase in the incidence of any tumor in male or female rats exposed to concentrations as high as SO ppm sulfuryl fluoride. There was no evidence of a nervous system effect based on functional observational battery, motor activity or histopathological examination of perfusion-fixed nervous system tissues (Spencer et aI., 1994). The no-observed-adverseeffect level (NOAEL) for chronic toxicity was 20 ppm. The NOEL was 5 ppm in males and 20 ppm in females due to several rats with very slight fluorosis.
88.4.12 CHRONIC TOXICITY-DOGS Groups of 4 dogs/sex were exposed to 0, 20, SO, or 200 ppm for 6 hours/day, 5 days/week for one year (Quast et aI., 1993b). Body weight gains in the highest exposure group of males and females were less than controls within the first two weeks of exposure to sulfuryl fluoride and the differences became greater throughout the study until the dogs were removed due to morbidity or death. Although no clinical effects were noted in dogs exposed to 200 ppm for the first eight months, clinical effects were observed in these animals at approximately !line months into the study. Observations in these animals included labored breathing, shallow, rapid respiration, and pale or blue mucous membranes. The onset of these observations was relatively swift; in the first dog, effects were noted on test day 263 and the animal died on test day 267. Due to the relatively swift onset, the last dog was sacrificed on test day 2S2 when the exposure was stopped due to excessive toxicity.
88.4 Toxicity to Laboratory Animals (Post-1980)
1891
120 _
Oppm
. . * . '5ppm 100
-+-- 20 ppm --e- 80ppm
80 ~ ~
0 60
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90
180
270
360
450
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630
810
720
Test Day
(a) 120 ~-----------------------------------------------------------------------,
100
130 0
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100
200
300
400
500
600
700
800
Test Day
(b) Figure 88.3
Mortality in (a) male and (b) female rats.
There were no exposure-related effects noted in dogs exposed to 20 or 80 ppm sulfuryl fluoride for one year or dogs exposed to 200 ppm through six months. Effects were
noted in hematological or clinical chemistry values of dogs exposed to 200 ppm for approximately nine months. However, by this point, these dogs were starting to show signs of
1892
CHAPTER 88
Sulfuryl Fluoride
500 ~------------------------------------------------------------------~
450 400 350
Cl 300 .s=
Cl
.~
250
>-
"C
0
co 200 150
_Oppm
100
··.·-5ppm --+- 20ppm
50
--e- ao ppm 0 0
50
100
150
200
250
300
350
400
450
500
550
600
650
700
750
800
Test day
(a) 300 .---------------------------------------------------------------------~
250
200
150
50
o
50
100
150
200
250
300
350
400
450
500
550
600
650
700
750
80
Test days
(b) Figure 88.4
Body weights of (a) male and (b) female rats.
respiratory distress and the observed effects were considered to be minor secondary hematological and clinical chemistry changes.
Gross examination of dogs exposed to 200 ppm revealed dark colored lungs which appeared to be consolidated. There were no other tissues affected in dogs exposed to 200 ppm for
88.4 Toxicity to Laboratory Animals (Post-1980)
approximately nine months or dogs exposed to concentrations of 20 or 80 ppm for one year. Histopathologic changes were noted in the lungs, brain, thyroid gland, and canine teeth of dogs exposed to 200 ppm and in the lungs and canine teeth of dogs exposed to 80 ppm. The pulmonary changes appeared to be a chronic active inflammation which primarily involved the peripheral regions of the lung of animals exposed to 200 ppm without recognizable alterations in the major airways. An increased number of alveolar macrophages was observed in scattered alveoli. In the more advanced stages of the chronic active inflammation process these foci apparently increased in size and hypertrophied type 11 pneumocytes were observed. In addition, epithelial cells were hypertrophied and hyperplastic. In the more severe cases, a focal thickening of the pleura and interalveolar septae was observed also. Dogs exposed to 80 ppm had a very slight increase in the aggregates of alveolar macrophages with several dogs exhibiting a very slight degree of the chronic active inflammatory process. In the brain, a focus of malacia was observed in five of eight dogs inhaling 200 ppm. Although inflammatory cells were noted, they appeared to be an insignificant factor. Very slight hypertrophy of the follicular epithelium was observed in the thyroid gland of all male and three female dogs. Since the dogs were approximately five months of age when initially exposed to sulfuryl fluoride, the teeth of these animals were still growing. Consequently during the beginning of the study, concentric rings were observed for each exposure period. Thus five concentric rings were observed for each week of exposure. However, as these animals matured it was more difficult to recognize the concentric rings. There were no exposure-related effects noted in dogs exposed to 20 ppm sulfuryl fluoride.
88.4.13 TERATOLOGY STUDIES-RATS Groups of 35-36 bred Fischer 344 rats were exposed to 0, 25, 75, or 225 ppm sulfuryl fluoride for 6 hr/day on days 6-15 of gestation (Hanley et aI., 1989). There was no evidence of embryotoxicity, fetotoxicity, or teratogenicity noted in rats exposed to concentrations as high as 225 ppm sulfuryl fluoride.
88.4.14 TERATOLOGY STUDIES-RABBITS Groups of 28-29 inseminated New Zealand White rabbits were exposed to 0, 25, 75, or 225 ppm sulfuryl fluoride for 6 hr/day on days 6-18 of gestation (Hanley et aI., 1989). Pregnant rabbis exposed to 225 ppm lost weight during the exposure period and did not gain weight in the postexposure period (days 19-29 of gestation). Body weights of rabbits exposed to 25 or 75 ppm were unaffected. Body weights of the fetuses from dams exposed to 225 ppm were significantly lower (14% decrease) than in the control group. Fetal crown-rump length was also slightly decreased in this group. There was no evidence of embryotoxicity, fetotoxicity, or teratogenicity noted in rabbits exposed to concentrations as high as 225 ppm sulfuryl fluoride.
1893
88.4.15 REPRODUCTION TOXICITY-RATS Groups of 30 male and 30 female Sprague-Dawley rats were exposed to 0, 5, 20, or 150 ppm sulfuryl fluoride for 6 hours/day, 5 days/week for 10 weeks for the FO and 12 weeks for the Fl generation prior to mating and 6 hours/day, 7 days/week during mating, gestation, and lactation through two generations (Breslin et aI., 1993). Body weights ofFO male and female rats exposed to 150 ppm were significantly decreased from control values during most of the premating period. The body weight gain of these female rats during gestation was also decreased. However, during the lactation period, the body weight gain was increased possibly as a compensatory mechanism. There was no exposure-related effect on the FO male or female conception index, fertility indices, length of gestation, time to mating, pup survival indices, or pup sex ratio. Similarly, there was no exposure-related effect on the number of Fl pups born dead or alive, or on the litter size at any exposure level. However, the litter size of live pups at birth and on days 1 and 4 before culling were increased 2.0 pupsllitter in the 5 ppm exposure group and 1.5 pupsllitter in the 20 ppm exposure group when compared to the control group. While these increases in litter size were not considered exposure-related, as a dose-response was not observed and no effects on litter size were observed in animals exposed to 150 ppm, the increase in litter size above control values has biological significance in that average pup weights are known to decrease with increasing litter size (Tyle, 1988). Indeed, Fl female pup body weights were occasionally decreased in the 5 ppm exposure group. On the other hand, the body weights of Fl male and female pups from dams exposed to 150 ppm were statistically decreased throughout most of the lactation period. While the decreases in body weight of Fl pups from the 150 ppm exposure group were attributed to treatment, these body weight effects were considered secondary to the decreased maternal growth observed throughout the premating and gestation periods. Body weight effects in the Fl adults were similar to those observed in the FO generation. Exposure-related effects were noted in body weights of male and female Fl rats exposed to 150 ppm during premating and female Fl rats exposed to 150 ppm during gestation and lactation but not at lower exposure levels. No exposure-related effects were observed on the Fl male or female fertility indices, length of gestation, time to mating, pup survival indices, or pup sex ratio in any exposure group. Similarly, no exposure-related effects on the number of F2 pups born alive or dead, or on the litter size were observed at any exposure level. No exposure-related effects on the body weights of male or female F2 pups from dams exposed to 5 or 20 ppm were observed at any time during the lactation period. However, body weights of male and female pups from dams exposed to 150 ppm were significantly decreased on lactation Days 14 and 21. This was considered to be secondary to the decreased maternal growth observed during the premating and gestation periods. The decreased growth of F2 pups from dams exposed to 150 ppm sulfuryl fluoride was less severe than the
1894
CHAPTER 88
Sulfuryl Fluoride
decreased weights observed in the F1 pups at the same exposure level. The pathologic changes in the teeth, lungs, and brain in the adult FO and F1 Sprague-Dawley rats in this study were essentially identical to those observed in a subchronic study in Fischer 344 rats previously mentioned (Eisenbrandt and Nitschke, 1989). Minor changes in the lung and brain were present at lower exposure concentrations in this study (20 ppm in the lung and 150 ppm in the brain) than previously noted in Fischer 344 rats. Interestingly, the effects observed in the brain of the Fl adults occurred in fewer animals than in the FO adults even though the length of exposure to sulfuryl fluoride was increased. Female rats had a higher incidence of brain and lung lesions than did males at a given exposure level. The parental NOEL was 5 ppm, the NOEL for neonatal growth was 20 ppm, and the NOEL for reproductive toxicity and fertility was 150ppm.
88.5 GENETIC TOXICITY
as yet been determined in mammals. However, sulfuryl fluoride has been shown to be fairly stable in saline (Waechter, 1991). At concentrations of 1000 and 5000 ppm in saline, 93% sulfuryl fluoride remained after 240 minutes at 37°C. Sulfuryl fluoride was less stable in rat blood. At concentrations of 1000 and 5000 ppm, 50 and 84% sulfuryl fluoride, respectively, remained after 220 and 255 minutes at 37°C. Studies in termites have demonstrated extensive dehalogenation of 35S-labeled sulfuryl fluoride (Meikle et aI., 1963). The fluoride ion may play a role in the mechanism of action of sulfuryl fluoride in insects and possibly also in mammals. Serum fluoride levels have been elevated from control values in several species, including mice, rats, rabbits, and dogs, following acute or subchronic exposure to sulfuryl fluoride. Many of the observations in rodents overexposed to sulfuryl fluoride seem to be typical of acute fluoride poisoning (Drill, 1954; Goodman et aI., 1980; Greenwood, 1940). 88.5.5 TOXICOLOGY IN HUMANS
There have been a few case reports of individuals exposed to sulfuryl fluoride. A 30-year-old male was exposed to unknown concentrations of sulfuryl fluoride in air containing Sulfuryl fluoride was tested in strains TA98, TA100, TA1535, 1% chloropicrin for 4 hours (Taxay, 1966). Nausea, vomitand TA1537 with and without metabolic activation (Gollapudi ing, cramps, abdominal pain, and itching were observed while et aI., 1990a). Petri plates were exposed for 4 hours at 37°C exposed to sulfuryl fluoride. Vital signs were normal upon adto nominal concentrations of 300, 1000, 3000, 10,000, and mittance to the hospital, however, reddening of the conjunctival, 30,000 ppm sulfuryl fluoride. The plates were incubated for an pharyngeal, and nasal mucosae; diffuse rhonchi; and parestheadditional two days prior to determining the frequencies of musia of the lateral surface of the right leg were observed. Serum tants/plate. Sulfuryl fluoride was not mutagenic in any of the fluoride levels were elevated above normal values. The signs tester strains. and symptoms resolved quickly and the patient was discharged from the hospital after four days. In a second case report, an 88.5.2 UNSCHEDULED DNA SYNTHESIS elderly couple returned to their home approximately 5-8 hours after their house was ventilated to remove any remaining sulThe genotoxicity of sulfuryl fluoride was evaluated in the rat furyl fluoride following approximately 24 hour fumigation of hepatocyte unscheduled DNA synthesis (UDS) assay (Golla- their home (Nuckolls et aI., 1987). Within 24 hours of their pudi et aI., 1991). In two separate assays, sulfuryl fluoride did return, the wife experienced weakness, nausea, and repeated not elicit a positive UDS response at nominal concentrations vomiting and her husband complained of dyspnea and restlessranging from 204 to 1020 ppm. ness. Within 48 hours the husband had a generalized seizure followed by cardiopulmonary arrest. The wife died within seven days due to ventricular fibrillation. Serum fluoride level of the 88.5.3 MICRONUCLEUS TEST wife six days after the house was fumigated was 0.5 mg/liter (background levels are highly dependent upon fluoride levels Sulfuryl fluoride was evaluated in the mouse bone marrow miin drinking water and range from 0.010 to 0.2 mg/liter) (Burcronucleus test (Gollapudi et aI., 1990b). Groups of mice were tis and Ashwood, 1999). A couple of individuals have entered exposed to 0, 50, 175, or 520 ppm sulfuryl fluoride. Mice were structures under fumigation with sulfuryl fluoride (Scheuersacrificed at 24, 48, or 72 hours after exposure to sulfuryl fluman, 1986). These individuals were found dead or died shortly oride. There were no significant increases in the frequencies of after exposure. The cause of death appeared to be severe pulmicronucleated polychromatic erythrocytes in the bone marrow monary edema with congestion. Serum fluoride levels in two of mice. individuals were elevated above normal values. Structural fumigators using sulfuryl fluoride were evaluated in a neurobehavioral battery (Anger et aI., 1986). While there 88.5.4 METABOLISM were no significant differences from a reference group, a greater The metabolism, disposition, and the relationships, if any, of number of symptoms and slightly reduced performance on cogthe metabolites to the mechanism of action of S02F2 have not nitive tests was observed in the fumigator group. However,
88.5.1 AMES TEST
References
educational levels, race, and use of illegal drugs were different between the sulfuryl fluoride workers and the referent group. Thus, the slight difference in cognitive test results was very likely due to differences other than exposure to sulfuryl fluoride.
88.5.6 SUMMARY RISK CHARACTERIZATION In the toxicity studies reported here, the NOEL ranged from 600 to 20 ppm (Table 88.4). The highest value is the two-week rat neurotox NOEL. In general, neurotox NOELs were higher than general tox NOELs. Remarkably little difference was observed in all four species exposed in two-week to chronic studies. The label instructions approved by the EPA specify instructions for sealing structures to confine the gas during the fumigation with tarps and/or sealing (Dow AgroSciences, 1996). After the appropriate fumigation period, the building is aerated using one of two specified aeration procedures and is dependent upon the concentration used. Depending upon which aeration procedure is used, the building must be secured for either 6 or 8 hours. After this waiting period, the concentration of Vikane in the breathing zones must be determined. If the concentration of Vikane is greater than 5 ppm, the structure must be venti1ated until the concentration is less than 5 ppm, at which point the structure may be reoccupied. The fumigation site cannot be reoccupied until aeration is complete. Only an approved detection device of sufficient sensitivity, such as the INTERSCAN or MIRAN gas analyzer, can be used to confirm a concentraTable 88.4 Summary of NOAELs for Various Species Exposed to Sulfuryl Fluoride General Tox
Neuro Tox
Repeated exposure studies
NOEL (ppm)
NOEL (ppm)
2-week rat
100
600
2-week rabbit
100
lOO
2-week mouse
30
30
100
100 30
2-week dog 13-week rat
30
13-week rat (electrophysiology)
30
13-week rabbit
30
13-week mouse
30 100 (M), 30 (F)
30
30
13-week dog
100
100
12-month rat
5 (M)*, 20 (F)
80
12-month rat (Neurotox guideline)
80
80
12-month mouse
20
20
12-month dog
20
20
24-month rat 18-month mouse Two-generation rat reproduction
5 (M)*, 20 (F)
80
20
20
5
20
*Microscopic dental fluorosis was observed in several 20 ppm male rats.
1895
tion of sulfuryl fluoride is 5 ppm or less. Warning signs must remain posted until aeration is determined to be complete.
REFERENCES ACGIH (1998). "TLVs and other Occupational Exposure Values-1998." Albee, R. R., et al. (1983). "Sulfuryl Fluoride (Vikane*): Induced Incapacitation in Rats." Unpublished report of The Dow Chemical Company. Albee, R. R., et al. (1993). "Sulfuryl Fluoride: Electrodiagnostic, FOB, and Motor Activity Evaluation of Nervous System Effects from Short-Term Exposure." Unpublished report of The Dow Chemical Company. Anger, W. K., et al. (1986). Neurobehavioral evaluation of soil and structural fumigators using methyl bromide and sulfuryl fluoride. Neurotoxieology 7, 137-156. Boorman, G. A., Eustis, S. L., Elwell, M. R., Montgomery, Jr., C. A., and MacKenzie, W. F. (1990). "Pathology of the Fischer Rat. Reference and Atlas." Academic Press, San Diego. Bradley, G. J., et al. (1990). "Sulfuryl Fluoride: Four-Hour Dermal Vapor Exposure in Fischer 344 Rats." Unpublished report of The Dow Chemical Company. Breslin, W. J., Liberacki, A. B., Kirk, H. D., Bradley, G. J., and Crissman, J. W. (1993). Sulfuryl fluoride: Two-generation inhalation reproduction study in Sprague-Dawley rats. Toxicologist 13, 368. Burtis, C. A., and Ashwood, E. R. (1999). "Tietz Textbook of Clinical Chemistry," 3rd ed. Saunders, Philadelphia. Dow AgroSciences (1996). Vikane Specialty Gas Fumigant Label. Dow AgroSciences (1997). "General Information on Vikane Gas Fumigant Product." Brochure 311-56-077. Drill, V. A. (1954). "Pharmacology in Medicine." McGraw-Hill, New York. Eisenbrandt, D. L., and Nitschke, K. D. (1989). Inhalation toxicity of sulfuryl fluoride in rats and rabbits. Fundam. Appl. Toxieol. 12,540-557. Eisenbrandt, D. L., Williams, D. M., Albee, R. R., and Streeter, C. M. (1987). "Sulfuryl Fluoride (Vikane* Gas Fumigant): An Ultrastructural Assessment of the Lungs of Rats Exposed to High Concentrations of Sulfuryl Fluoride." Unpublished report of The Dow Chemical Company. Gollapudi, B. B., Samson, Y. E., and Zempel, J. A. (1990a). "Evaluation of Sulfuryl Fluoride in the Ames SalmonellalMammalian Microsome Bacterial Mutagenicity Assay." Unpublished report of The Dow Chemical Company. Gollapudi, B. B., McClintock, M. L., and Nitschke, K. D. (1990b). "Evaluation of Sulfuryl Fluoride in the Mouse Bone Marrow Micronucleus Test." Unpublished report of The Dow Chemical Company. Gollapudi, B. B., McClintock, M. L., and Zempel, J. A. (1991). "Evaluation of Sulfuryl Fluoride in the Rat Hepatocyte Unscheduled DNA Synthesis (UDS) Assay." Unpublished report of The Dow Chemical Company. Goodman, A. G., Goodman, L. S., and Gilman, A. (1980). "The Pharmacological Basis of Therapeutics," 6th ed. Macmillan, New York. Gorzinski, S. J., and Streeter, C. M. (1985). "Effect of Acute Vikane Exposure on Selected Physiological Parameters in Rats." Unpublished report of The Dow Chemical Company. Greenwood, D. A. (1940). Fluoride intoxication. Physiol. Rev. 20,582-616. Hanley, Jr., T. R., Calhoun, L. L., Kociba, R. J., and Greene, J. A. (1989). The effects of inhalation exposure to sulfuryl fluoride on fetal development in rats and rabbits. Fundam. Appl. Toxieol. 13, 79-86. Landry, T. D., and Streeter, C. M. (1983). "Sulfuryl Fluoride: Effects of Acute Exposure on Respiration in Rats." Unpublished report of The Dow Chemical Company. Lockwood, D. L. (1958). "Results of Dietary Feeding of Rats with Feed Fumigated with Sulfuryl Fluoride (Vikane)." Unpublished report of The Dow Chemical Company. Mattsson, 1. L., Albee, R. R., Eisenbrandt, D. L., and Chang, L. W. (1988). Subchronic neurotoxicity in rats of the structural fumigant, Sulfuryl Fluoride. Neurotoxieol. Terato!' 10, 127-133. Meikle, R. w., Steward, D., and Globus, O. A. (1963). Fumigant mode of action, drywood termite metabolism of Vikane fumigant shown by labelled pool technique. J. Agr. Food Chem. 11, 226-230.
1896
CHAPTER 88
Sulfuryl Fluoride
Miller, R. R., et al. (1980). "Sulfuryl Fluoride (Vikane Fumigant): An LCSO Determination." Unpublished report of The Dow Chemical Company. Mohr, U., Dungworth, D. L., and Capen, C. C. (1992). "Pathobiology of the Aging Rat," Vo!. I. International Life Sciences Institute, Washington, DC. Nitschke, K. D., and Lomax, L. G. (1989). "Sulfuryl Fluoride: Acute LCSO Study with B6C3FI Mice." Unpublished report of The Dow Chemical Company. Nitschke, K. D., and Quast, 1. F. (1990). "Sulfuryl Fluoride: Acute LC50 Study with CD-l Mice." Unpublished report of The Dow Chemical Company. Nitschke, K. D., and Quast, J. F. (1991). "Sulfuryl Fluoride: Two-Week Inhalation Toxicity Study in Beagle Dogs." Unpublished report of The Dow Chemical Company. Nitschke, K. D., and Quast, J. F. (1993). "Sulfuryl Fluoride: Thirteen-Week Inhalation Toxicity Study in CD-l Mice." Unpublished report of The Dow Chemical Company. Nitschke, K. D., and Quast, J. F. (1995). "Sulfuryl Fluoride: Two-Week Inhalation Toxicity Study in CD-l Mice." Unpublished report of The Dow Chemical Company. Nitschke, K. D., Albee, R. R., Mattsson, J. L., and Miller, R. R. (1986). Incapacitation and treatment of rats exposed to a lethal dose of sulfury I fluoride. Fundam. Appl. Toxicol. 7, 664-670. Nitschke, K. D., Beekman, M. J., and Quast, J. F. (1992). "Sulfuryl Fluoride: 13-Week Inhalation Toxicity Study in Beagle Dogs." Unpublished report of The Dow Chemical Company. NuckoIls, J. G., Smith, D. c., Walls, W. E., Oxley, D. w., Hackler, R. L., Tripathi, R. K., Armstron, C. w., and Miller, G. B. (1987). Fatalities resulting from sulfuryl fluoride exposure after home fumigation-Virginia. lAMA 258,2041-2044. Quast, J. F., Bradley, G. J., and Nitschke, K. D. (1993a). "Sulfuryl Fluoride: 18Month Inhalation Oncogenicity Study in CD-I Mice." Unpublished report of The Dow Chemical Company.
Quast, J. F., Beekman, M. J., and Nitschke, K. D. (1993b). "Sulfuryl Fluoride: One-Year Inhalation Toxicity Study in Beagle Dogs." Unpublished report of The Dow Chemical Company. Quast, J. F., Bradley, G. J., and Nitschke, K. D. (1993c). "Sulfuryl Fluoride: 2-Year Inhalation Chronic Toxicity/Oncogenicity Study in Fischer 344 Rats." Unpublished report of The Dow Chemical Company. Scheuerman, E. H. (1986). Suicide by exposure to sulfuryl fluoride. l. Forensic Sci. 31, 1154-1158. Spencer, P. J., Bradley, G. J., and Quast, J. F. (1994). "Sulfuryl Fluoride: Chronic Neurotoxicity Study in Fischer 344 Rats-Final Report." Unpublished report of The Dow Chemical Company. Taxay, E. P. (1966). Vikane inhalation. l. Occup. Med. 425-426. Torkelson, T. R. (1959). "Summary Report of Toxicological Studies with Vikane (Sulfuryl Fluoride, S02F2)." Unpublished report of The Dow Chemical Company. Torkelson, T. R., Hoyle, H. R., and Rowe, V. K. (1966). Toxicological hazards and properties of commonly used space, structural and certain other fumigants. Pest Control 1-8. Tyle, R. W. (1988). "Perspectives in Modem Toxicology," Chap. 8. Wright, London. U. S. Environmental Protection Agency (1982). "Pesticide Assessment Guidelines, Subdivision F, Hazard Evaluation: Human and Domestic Animals," pp. 98--100. U.S. Environmental Protection Agency, Washington, DC. Vaccaro, (1988). Unpublished data of The Dow Chemical Company. Vernot, E. H., et a!. (1977). Acute toxicity and skin corrosion data for some organic and inorganic compounds and aqueous solutions. Toxicol. Appl. Pharmacol. 42,417-423. Waechter, J (1991). Personal communication. Unpublished data of The Dow Chemical Company.
Index
A Abamectin, 1157 Absolute bioavailability, 569, 905 Absorption, 507-508, 905-912 distribution and, 784, 930 gastrointestinal tract, 565-566 phannacokinetics of, 563-581, 783-785 respiratory tract, 566-567 types of, 77-79 See also specific substances Acaricides, 191-194, 1187 Acequinocyl, 1208 Acetochlor, 1547-1549, 1555-1556 Acetylcholinesterases (AChEs), 33, 292, 549, 929, 1018-1020 acetylcholine synthesis, 877, 1018-1019, 1045, 1093 BuChEs and, 972, 1073-1074 carbamates, 1091-1092, 1093 electrophysiology, 943, 1051 function of, 967-968, 1045 inhibition of, 919, 942-946, 972, 979-980, 1017, 1073-1074, 1092 MUDDLES and, 1094 organophosphorous compounds, 1013, 1019 oximes, 1057 reactivation, 979-980, 1031 reactivators, 1034 release of, 877,967-968 See also Anticholinesterases ACGIH. See American Conference of Governmental Industrial Hygienists Acidity, of soil, 1781 Action, mechanisms of, 7, 377-378 Activated charcoal, 591-592 Activation-deactivation, 647 Activity level, model for, 1116-1117 Acute exposure, defined, 691-706, 888 Acute illness and injury, 629--630 Acute neurotoxicity, 269, 292-293 Acute reference dose, defined, 1772 Acylcholine acylhydrolase, 967 Additive effects, 791 Additives, 812 Adenine nucleotide transporter, 1176 Adrenal gland, 1335 Adsorption, 2,431
Aflatoxin, 802 AGDRIFf model, 363-364 Age, 609 biotransfonnation and, 873 children, 609, 685, 874-877, 887-904, 944-948 differences and, 520 pesticide classes and, 611 sensitivity and, 785-786, 873, 882 site of exposure and, 610 susceptibility and, 61--62 symptoms and, 608 toxicity and, 873-886 Agency for Toxic Substances Disease Registry, 598 Agent Orange, 48, 384 Aggregate assessment, 443 Aggregate exposure, 479-490, 684--685 Air sampling method, 431 Alachlor, 324, 911, 1543-1547, 1554-1555 Albumin, 789 Aldicarb, 776, 880, 1088, 1107-1121 Aldrin, 390, 1131-1134 Alimentary elimination, 585 Alkyl mercury compounds, 1375-1379 Allergy, 42-43, 182,300 Allethrin, 880, 1263-1264 Alternative pest management, 232-234 Altitude effects, 68 Aluminum phosphide, 323 Alzheimer's disease, 1044 Ambient air concentration, 462-466 American Conference of Governmental Industrial Hygienists (ACGIH), 1103 American Society for Testing and Materials (ASTM),360 Amidases, 540 Amines, 312, 1239-1244 Aminocarb, 776 Amitraz, 317, 390 Amitrole, 384 Anabasine, 121 Analytical methods, 672--678 Anatomical site differences, 519-520 Anilazine, 321 Anilinopyrimidine, 1701-1710 Animal health products, 406-410 Animal studies biological response, 942-947
1897
sex-specific variation, 786-788 species differences, 57 strain differences, 57-58 susceptibility of, 58--60 using small numbers, 13 See also specific pesticides Annual average daily dosage, 716 ANT. See Adenine nucleotide transporter Antagonism, 49-50, 509-510, 550 Anthelmintics, 406 Anti-vitamin K compounds, 1809-1823 Antiandrogens, 734-736 Antibiotics, 191 Anticholinesterases, 772-777,954 aggression, 272 atropine, 1056 chemistry and biochemistry, 1045 as drugs, 1044-1045 organophosphorous compounds, 1019, 1034-1035, 1043-1085 pyrethroids, 1293 toxicities of,972-974 warfare agents, 1043-1045 See also Acetylcholinesterases Antidepressants, 792 Antimalarials, 406 Antimicrobial compounds, 300, 312, 683--684 Antimony, 1390--1392 Antimony potassium tartrate, 1391-1392 Antiparasitic agents, 408-409, 881, 1157-1167 Apoptosis, 1177 Applied dose, defined, 888 Aquatic systems, 77, 360, 664--667 Aromatic acid biosynthesis, 1667-1671 Arsenic, 1392-1399 Arsenolysis, 1176 Arthropods, 182, 187-188 Artificial neural network (ANN) mode ling, 230 Aryl mercury compounds, 1379-1380 Aryloxyalkanoic acids, 104 Arylphenoxyproprionic acid, 104 Assay techniques, 974-978 Assessment endpoints, 357 ASTM. See American Society for Testing and Materials Atmospheric pressure chemical ionization (APCI), 675
1898
Index
ATP synthesis, 1171-1174, 1209-1210, 1225 Atrazine, 326, 714, 738 Atropine, 594, 980, 1056, 1103 Authorization procedures, 473 Avermectins, 317, 1157-1167 Avoidance methods, 190,279 Azadirachtin, 130-134,316-317 Azocyclotin, 1215-1216 Azoxystrobin, 1200-1202
B Bacteria antibiotics, 144-146, 149-151, 191 antimicrobials, 300, 3 I 2, 683-684 bactericides, 149-151, 323-324, 862-866 reverse mutation assay, 749-750 types of, 862-866 See also specific types, substances Baculoviruses, 867 Baits, 247-249, 257-259 BAL. See British anti-Lewisite Barium, 1357-1359 Barium carbonate, 1357-1359 Barometric pressure, 68 Batrachotoxin (BTX), 336 Baygon,755 BCF. See Bioconcentration factor Behavioral chemicals, 278-279 BEL See Biological exposure index Benchmark approach, 644--646, 805 Bendiocarb, 399 Benefin,325 Benomyl, 319, 401,1673-1694 Bensulide,325-326 Bentazon, 756 Benzene hexachloride, 598 Benzimidazoles, 406, 1673-1694 Benzodiazepine receptor, 1293 Bhopal incident, 597 Bifenthrin, 1264-1265 Bilanafos, 152-153 Biliary excretion, 80 Bioaccumulation, 661-664 Bioactivation hypothesis, 1134 Bioallethrin, 880, 1263-1264 Bioavailability, 569, 905 Biocides, 323-324 Bioconcentration factor (BCF), 596 Biocontrol, 232-235, 279-280 Biodegradability, 206,661 Biological control. See Biocontrol Biological exposure index (BEl), 599 Biological magnification, 22 Biological response, 907, 942-947 Biological rhythms, 75 Biological storage, 270 Biomarkers, 585-586, 896-900 Biomonitoring, 226, 431--432, 457, 599, 693, 758-761,941 Bioresmethrin, 1283-1284 Biotransformation, 37,45,531 age-related differences, 873 diet and, 63 extrahepatic tissues, 540-542 pesticides and, 271 phase 1,79 phase 11, 79-80
See also Metabolism Biphasic effects, 551 Bipyridyls, 324, 384 Birds, 227, 261-262 Birth defects, 380 Bismuth, 1389-1390 Bismuth subcarbonate, 1389-1390 Bismuth subsalicylate, 1390 Blasticidin-S, 144-146 Blood-brain barrier, 568, 1159-1160 BLS. See Bureau of Labor Statistics Body louse, 193 Body water, 567 Boehringer-Mannheim kits, 975 Bone marrow cells, 773 Borates, 316, 1429-1437 Boric acid, 1413-1414, 1429-1437 Botanical insecticides, 11 0-116 Boyd method, 20 Boyle's law, 6 Brain, neurology of, 1027-1029 Breast cancer, 730 British anti-Lewisite (BAL), 51 Brodifacoum, 1818-1820 Bromacil, 327 Bromadiolone, 1822 Bromethalin, 1239-1241 Bromobenzene, 38 Bromoxynil, 384, 1232-1235, 1234 Bronsted-Lowry theory, 564 BTX. See Batrachotoxin BuChEs. See Butyrylcholinesterases Bureau of Labor Statistics (BLS), 619-620, 638 Burrow fumigants, 259 Butachlor, 1550-1551, 1555 Butoxypolypropylene glycol, 317 Butyrylcholinesterases (BuChEs), 930, 967, 972, 1073-1074
c Cl A index, 20 Cadmium, 1367-1369 Cadmium chloride, 1367-1369 Caffeic acid, 801 Cage convulsants, 1143-1145 Cage effect, 66 Calander model, 449--450 Calciferol, 1825 Calcineurin, 342 Calcium channels, 341, 1002-1003 California Department of Pesticide Regulation (CDPR), 617-619, 693 California Environmental Protection Agency, 638 cAMP. See Cyclic adenosine monophosphate Cancer risks, 600--601, 715, 718, 721, 730, 756-757,799-844. See also Carcinogenicity; specific substances Capillary electrophoresis (CE), 101,677-678 Captafol, 318-319,1733 Captan, 318,547,550, 1711-1742 Carbamates, 102,314,319-320,326,390, 399--400,755,775-777,880,1087-1092, 0-110121 Carbaryl, 400, 775, 1098, 1099, 1101 Carbendazim, 1673-1694 Carbofuran, 399,775-776, 1097, 1098 Carbon hydroxylation, 538
Carbon tetrachloride, 50, 53,916-917 Carboxamides, 1193-1199 Carboxin, 321, 1194 Carboxylesterases, 923 Carcinogenic Potency Database (CPDB), 799-844 Carcinogenicity, 295,739, 747, 823-827 benchmark dose approach, 805 cell division and, 804-806 chemicals evaluated for, 803 chronic animal tests, 802-804 database, 799-844 DDT and, 1318-1319 genotoxicity,752-757 logtime-logdosage curve, 25 natural pesticides, 801 ranking of, 806-816 risks and, 600--601, 715, 718, 721, 730, 756-757,799-844 short-term, 295 toxicity and, 41 See also specific substances Cardiovascular system, 1034-1035, 1048-1049 CARES. See Cumulative and Aggregate Risk Evaluation System CAT. See Choline acetyltransferase Catalytic hydrolysis, 922 Cathartics, 592 Cats, 265 Cattle, 263-264 CDC. See Centers for Disease Control and Prevention CDPR. See California Department of Pesticide Regulation CE. See Capillary electrophoresis Cell division, 804-806 Cell membranes, 563 Cellular elimination, 585 Centers for Disease Control and Prevention (CDC),620 Central America, 626 Central nervous system, 542, 1020-1021, 1024-1031 Cerebral cholinergic signalling, 1024-1026 Cevadine, 126-127 Chemical ionization, 675 Chemical-specific exposure, 479 Chemical structure, 646-647 Chemical warfare agents, 980 ChEs. See Cholinesterases Children, 609, 685, 874-877, 887-904, 944-948 Chirality, 98, 959 Chloracetanilides, 324, 1543-1557 Chloralose, 1827-1829 Chlordane, 770-771, 1135-1138, 1179 Chlordecone, 397,731-733,880, 1139-1141 Chlordene, 1135-1138 Chlordimeform,881 Chlordithane, 1332 Chlorfenapyr, 1241-1243 Chloridazon, 1179 Chloride channel, 1292 Chloride ionophores, 1145-1151 Chlorinated cyclodienes, 390-397 Chlorinated hydrocarbons, 38, 589-590, 1180 Chlorinated insecticides, 1131, 1305-1355. See alsoDDT
Index
Chlorine, 77, 312, 1411-1413. See also specific compounds Chloro-triazines. See Cyanazine Chloroacetanilide herbicides, 536 Chlorobenzilate, 1341-1342 Chlorogenic acid, 801 Chloroneb, 321 Chloronicotinyls, 1123 Chlorophacinone, 1820-182 I Chloropicrin, 188 I Chloroquine, 406 Chlorothalonil, 1743-1757 Chlorphenoxy compounds, 384 Chlorpyrifos, 398, 756, 894, 938, 943, 945, 946, 1049 Chlorsulfuron, 327 Cholecalciferol, 1826-I 827 Choline acetyltransferase (CAT), 1018 Cholinergic system, 594, 1019-1033, 1043-1062 Cholinesterases, 607,695,789,792-793,967-986, 1092, 1110-1113. See also Acetylcholinesterases; Butyrylcholinesterases Choreoathetosis, 1293 Chromatographic method, 651-652 Chromosome aberration assay, 751 Chronic exposure, 707-726, 888 Chronic poisoning, 595-599 Chronic studies, 294-295 Chronicity index, 20-21 Cinerin, I 11 Circadian rhythms, 75 Cismethrin, 1283-1284 Classification systems, 913-917 CLOGP program, 657 Cloning, 961-963 Cloransulam-methyl, 1653-1655 Clothinanidin, 1123 Clothing, 474,497,910 CMAPs. See Compound muscle action potentials Coffee, 801 Cohort studies, 635 Colony forming units, 864 Coma, 590 Comfrey,81O Compartmental models, 571, 645-646, 930-933 Complex HI, inhibitors of, 1199-1209 Complex V, inhibitors of, 1224-1225 Compound muscle action potentials (CMAPs), 1051, 1063 Computer-aided procedures, 656-658. See also specific programs Concentration exposure, 286 Confounding factors, 382 Conjugation reactions, 99, 540, 926 Consumer right-to-know, 685-686 Contact factors, 446-447 Coppe~320-321, 1359-1363 Copper sulfate, 1361-1363 Costa Rica, 626 Coumafury1, 1816 CPDB. See Carcinogenic Potency Database Crab louse, 193 Crack and crevice treatments, 247 Crop rotation, 279 Crowding, 64-65 Cultural influences, 279, 847-851
Cumulative and Aggregate Risk Evaluation System (CARES), 438 Cumulative assessment, 443, 685 Cumulative effects, 19-22 Cushing's syndrome, 1335 Cutaneous metabolism, 526 Cyanazine, 326-327,707-726 Cyanogenesis, 823 Cyazofamid, 1205, 1206 Cyclic adenosine monophosphate (cAMP), 1013 Cyclodienes, 342-344,597, 1131-1132, 1141-1147 Cycloprothrin, 1265-1266 Cyfluthrin, 315, 1266-1267 Cyha10thrin, 315, 1267-1268 Cyhexatin, 1216-1217 Cypermethrin, 315,881, 1268 Cyphenothrin, 1268-1269 Cyprodinil, 1701-1710 Cyromazine,714 Cytochrome oxidase, 1173 Cytochrome P450, 534-538, 788-789, 920-921 Cytogenetic assay, 751
D DAG. See Diacylglycerol Danshen,791 Dazomet, 323 DBCP. See Dibromochloropropane DCBP. See Dichlorobenzophenone DDD. See Dichlorodiphenyldichloroethane DDT. See Dichlorodiphenyltrichloroethane Death, 620-621 Decision support, 278 Decontamination, 905-912 DEET. See Diethyltoluamide Defoliant studies, 1020 Deformations, 382 DEG S19 method, 672 Degradation calculation model, 451 Deguelin, 1184 Dehydroepiandrosterone, 810 De1aney clause, 684 Delayed polyneuropathy, 1064-1067 Delivered dose, defined, 888 Delta-chlordane, 1141 Deltamethrin, 336,880,881,1270-1271 Demethylation rate, 666 Department of Health and Human Services (HHS), 681 Derived compounds, 47 Dermal exposure, 425, 459-473 absorption by, 78, 515-526, 783, 895, 905-912 aldicarb, 1119-1121 decontamination, 591,910-911 dermatitis, 328-330 hydration and, 524-525 irritation, 288, 299, 300 modeling,466-471 occlusion and, 524-525 patch method, 428-429 patterns of, 299-300 risk assessment, 1119-1121 sensitization, 288-289 Derris powder, 124 DES. See Diethylstilbestrol Desferrioxamine, 1576
1899
Desulfuration reaction, 920 Deterministic-based accumulation approach, 480 Developmental effects, 375-410, 551, 714, 1297-1298 DHEA. See Dehydroepiandrosterone Diacylglycerol (DAG), 1021, 1024 Diafenthiuron, 1224-1225 Dialkyl phosphates, 916 Dialkyldithiocarbamates (EBDCs), 1759-1779 Diazepam, 1059 Diazinon, 398, 937, 938 Dibromochloropropane, 396,637,737 Dicamba, 1639-1640 Dicarboximide fungicides, 734-735 Dichlobenil, 327 Dichlofluanid, 1733-1734 Dichlorobenzene, 757 Dichlorobenzophenone, 646 Dichlorodiphenyldichloroethane (DDD), 1332 Dichlorodiphenyltrichloroethane (DDT), 11, 335-342,396,728-731,770,811,1131, 1305-1332 analogs of, 1305-1355 animals and, 1306-1318, 1321 behavioral effects, 1321 biochemical effects of, 1316-1317 carcinogenesis and, 1318-1319 DDE and, 735, 1312 dosage response, 1328-1329 enzyme induction, 44 excretion, 1314-1316, 1331 fat mobilization, 1317 first synthesized, 1305 intoxication in animals, 1306-1309 metabolism of, 1312-1314 mutation, 1318-1319 nervous system and, 1317-1318 pathology, 1321 poisoning treatment, 1331-1332 secretion in milk, 1331 storage of, 1311-1312, 1329-1330 structure of, 1306 Sweden banned, 1306 symptomatology of, 1305-1332 toxicity, 1307, 1322-1332 trade names for, 1306 treatment of poisoning, 1321-1322 Dichlorophenyltrichloroethane, 802 Dichloropropene, 323, 1840-1859 DicIosulam, 1656-1657 Dicofol, 316, 646,1342 Dicoumarol, 1809 Dieldrin, 342-343, 597, 771, 1131-1134, 1148 Dietary exposure, 63, 692 aggregate, 443-454 biotransformation and, 63 contact factors, 446-447 cumulative, 443-454 cyanazine, 710,711,719-720 Exposure 4 program, 717-718 model for, 443-454 naturally occurring chemicals, 822-834 residue factors for, 448 Diethylstilbestrol (DES), 727-746 Diethyltoluamide (DEET), 196,317,881, 1439-1456 Difenacoum, 1821
1900
Index
Difethialone, 1822-1823 Differential sensitivity, 873-874 Diflubenzuron, 401 Dihydroheptachlor, 1135-1138 Diisopropylfluorophosphate, 933 Dilute/shoot approach, 676 Dimethoate, 398 Dinitrophenols, 1227-1242 Dinocap, 401, 1228 Dinoseb, 384, 1179 Dioxin, 777 Diphacinone, 56, 1817-1818 Diphenadione, 56 Diquat, 384,698-700,1605-1621 Direct-partitioning methods, 650-651 Dissipation process, 644--647 Distribution, 563-581, 784 Dithiobiuret, 881 Dithiocarbamates, 737-738 Dithionitrobenzoic acid (DTNA), 975 Dithiopyr, 326 Diuron, 327, 390, 1521-1523 Dizocilpine, 1032 DNA adduct measurement, 599 DNA biosynthesis, 1673-1694 Dogs, 265 Domestic animals, 263-272 Dosage, 46--47 control of, 46 defined, 9, 10, 888 dose vs, 9. See also Dose duration of, 53-54 ED 50. See ED values exposure and. See Exposure LD 50. See LD values LOEL. See Lowest-observed-effect level logtime-logdosage curve, 23 NOEL. See No-observed-effect level radiation and, 17-18 response and. See Dosage-response relationships schedule of, 53 small,31-39 statistical methods, 30-31 tissue level, 28 See also Pharmacokinetics; specific substances Dosage-response relationships, 9, 286 across species, 10 chemical basis of thresholds in, 37-38 cumulative lognormal form, 18-19 curve for, 12 logprobit model and, 34-35 models of, 37 no-effect level, 10 small,34-35 small doses, 31-39 toxicity and, 40-45 See also Dosage Dose, 1-82,571-577 biomarker measurements, 896--900 defined, 4, 10, 888 dosage vs, 9. See also Dosage dosimetry, 226, 942-947 effects of small, 32-39 exposure. See Exposure pharmacology. See Pharmacokinetics physicochemical properties, 521-522 testing, 381-382
therapeutic drugs, 34 time and, 4-6, 46 volume of, 14 See also Pharmacokinetics; specific substances Draize test, 326 Drinking water, 217-222, 723 Druckrey studies, on cancer, 3, 17 DTB. See Thioimidodicarbonic diamide DTNA. See Dithionitrobenzoic acid Dunnett's test, 714 Duration exposure, 286 Dyes, 429 Dynamics, 51
E EAA. See Excitatory amino acids Earthworms, 227-228 EBDC. See Dialkyldithiocarbamates EC. See European Community Ecologic studies, 637 Ecosystem level, 361 Ecotoxicological assessment, 353-371 Ectoparasites, 193 ED values, 10-19 Edaphology, 204-205 EDB. See Ethylene dibromide EDSTAC. See Endocrine Disruptor Screening and Testing Advuisory Committee EDTA. See Ethylenediaminotetrareitic acid Effect measures, 357 Eggs, 585, 731 EHs. See Epoxide hydrolases Einstein proposal, 39 Electron transport, 1171 Electrophysiology, 1029, 1051 Elimination half-life, 572 Ellman assay, 975, 977 Emamectin, 1157 EMC. See Encephalomyocarditis Empenthrin, 1271-1272 Encephalitis, 184, 1052 Encepha10myocarditis (EMC), 66 Endocrine system, 61, 727-746 Endosulfan, 316, 756, 879, 1139 Endrin, 597, 1131-1134, 1180 Energy, and matter, 39 Environment, 52,99-101, 190,206,261,353-371, 524-526,611,643-648,692. See also Environmental Protection Agency; specific substances, standards Environmental Protection Agency (EPA), 223, 627-628,638,688-689,816--818. See also specific regulations, standards Enzymes, 537-547, 788-789, 967 acetylcholinesterase, 1014 DDTand,44 induction of, 43-44, 51, 548 inhibition of, 548-551 physiological factors, 551-552 pyrethroid action, 342 See also specific effects Epicutaneous test methods, 300 Epidemiology, 376--377,592-593 cycles, 187 fundamentals of, 634-638 organophosphate esters, 775 principles of, 635
EPN. See Ethyl p-nitrophenyl theonobenzenephosphonate Epoxide hydrolases (EHs), 539-540, 1133, 1142 EPTC. See Ethyl dipropylthiocarbamate Ergocalciferol, 1823-1826 Ergodynamics, 7 Erosion control, 205-206 Estrogens, 686,728-734 ET values, 22-23 Ethalfluralin, 325 Ethofumesate, 389 Ethopropr model, 693-695 Ethyl carbamate, 776 Ethyl dipropylthiocarbamate, 389 Ethyl p-nitrophenyl theonobenzenephosphonate (EPN),50 Ethylan, 1336--1337 Ethylene dibromide (EDB), 404, 752-754 Ethylene oxide, 323, 752, 761 Ethylene thiourea, 811 Ethylenediaminotetrareitic acid (EDTA), 51 Ethyltin, 1387-1389 Etofenprox, 1272 ETU. See Ethylene thiourea European Community (EC), 671, 672 Europoem model, 493, 495, 501-504 Exams model, 363 Excitatory amino acids, 1031-1032 Excretion, 270, 511 biomarkers, 584-586 kidneys, 785 obscure routes of, 585 pesticides and, 583-586 radioactivity, 905-906 routes of, 584-585 Exposure, 377,425-432,874 aggregate, 449, 452-454 assessments of, 363, 383, 458-459, 887-891 biomarker measurements, 896--900 blood analysis, 979-980 characterizing, 361-364 components of, 286 cumulative, 448-449 data for, 692-693 dietary, 717-718 dose, 887-900. See also Dose duration of, 888 estimations of, 363-365 humans and, 222-227 loading, 459 mechanisms of, 377, 459 methods for measuring, 363, 428-431,887-891 mixing, 459 models, 444-445 operators, 459-462 risk assessment, 820-821 routes of, 54-55, 425-426 timing, 377-378 uncertainty, 452 See also Dose; specific effects, substances Extonet model, 638 Extraction methodology, 673-674 Extrapolation methods, 359-360 Extravascular dose, 573-577 Eye contamination, 592
Index
F Famoxadone, 1205-1206 Fat, effects of, 64, 568 Fate, environmental, 643-648 FDA. See Food and Drug Administration Fechner's law, 39 Federal Food, Drug, and Cosmetic Act (FFDCA), 681,684-686,818,875 Federal Insecticide, Fungicide and Rodenticide Act (FIFRA), 190,285, 374, 383, 507, 627, 681-684,691,987,999 Federal reporting requirements, 627-628. See also specific agencies Females, 737-739, 787-788 Fenamidone, 1205, 1207 Fenazaquin, 1187-1189, 1191 Fenbutatin-oxide, 1219-1220 Fenfuram, 1196 Fenpropathrin, 1272-1273 Fenpyroximate, 1189-1190 Fenthion, 398, 1070 Fentin, 1220-1224, 1385-1387 Fentin acetate, 1385-1387 Fenvalerate, 315, 1273-1275 Fermentation processes, 106 Fetal exposure, 585, 784-786 FFDCA. See Federal Food, Drug, and Cosmetic Act FFQ. See Food Frequency Questionnaire Field scouting, 278 Field workers, 328-330 FIFRA. See Federal Insecticide, Fungicide and Rodenticide Act Fink-Heimer silver technique, 990, 996 Fipronil, 344-345 Fixed dose approach, 287 Flavin-containing monooxygenases (FMOs), 511, 532,538-539,922 Flock dispersal agent, 261 Florasulam, 1657-1660 Flourescent tracers, 429 Fluazinam, 1243-1244 Flucythrinate, 1275-1276 Flumethrin, 1276 Flumetsulam, 327-328, 1660-1661 Fluometuron, 1524-1525 Fluorine, 1407-1411 Fluoroacetamide, 1799-1801 Fluoroacetic acid, 1794-1802 Fluoroethanol, 1801 Fluoxetine,791-792 Flusilazole, 321 Flutolanil, 1196-1197 Fluvalinate, 1276-1278 FMOs. See Flavin-containing monooxygenases Folpet, 319, 1711-1742 Food, Agriculture, Conservation, and Trade Act, 682 Food and Drug Administration (FDA), 816-818. See specific regUlations, standards Food Frequency Questionnaire (FFQ), 446-448 Food Quality Protection Act (FQPA), 190,383, 480,507,671,682-686,707,741-742,793, 875,887-904,1556 Food, residues in, 671-678, 799-844 Food supply survey, 447 FOODCONTAM database, 820
Formamidines, 738, 881 Fosetyl-aluminum,321 Fosthiozate, 323 FQPA. See Food Quality Protection Act Fragmental method, 656 Free-energy parameters, 655-656 Free radical scavengers, 1576-1577 Free-solution capillary electrophoresis (FSCE), 677 FSCE. See Free-solution capillary electrophoresis Fumigants, 259, 323-324, 406-410, 462 Functional redundancy, 357 Fungicides, 105, 144-149,207,318-323, 401-406,590 American farmers and, 105 anilinopyrimidine, 1701-1710 benzimidazoles, 1673-1694 blasticidin-S, 144-146 developmental toxicity, 402-404 kasugamycin, 146-147 list of, 402-404 mildiomycin, 147-148 miscellaneous structures, 321-323 reproductive toxicity, 402-404 turfgrass, 209 validamycin A, 148-149 See also specific substances Furametpyr, 1197-1198 Furfural, 812
G GABA. See Gamma-aminobutyric acid Galton, Francis (1879), 39 Gamma-aminobutyric acid (GABA), 342-344, 1013, 1025-1031, 1131-1160, 1293 Gas chromatography, 674-675 Gastrointestinal tract, 77-78, 565-566, 592, 601, 783 Gating kinetics, 335-336 Gaussian curves, 12 Gel permeation chromatography, 674 Gender differences, 552, 737-739, 787-788, 848 Gene expression, 1028 Gene mutation assay, 751 GENEEC model, 363-364 Generic mode ling approach, 494-495 Genotoxicity,747-768 carcinogenicity, 752-757 cyanazine,713 polymorphisms, 511, 552, 788-789 predisposition, 788-789 risk assessment, 761-762 studies of exposed workers, 759-761 testing, 748-751 Geometric mean, 39 Geriatric populations, 786 Ginger Jake, 987 GIT absorption, 784 GLEAMS model, 229 Glomerular filtration, 583 Glucocorticoids, 1002 Glue boards, 260 Glufosinate, 153-155 Glutathione transferases, 540, 789, 926 Glycoproteins, 789, 1159-1160 Glyoxylate, 37 Glyphosate, 325, 1667-1671
1901
Goats, 264 Golf courses, 206, 214, 218, 221 GPe. See Gel permeation chromatography GPMT. See Guinea-pig maximization test Grading criteria, 28, 496 Grayanotoxin (GTX), 337 Greenhouses, 457-475 Groundwater, 205-206,218-222 GTP. See Guanosine triphosphate GTX. See Grayanotoxin Guanosine triphosphate (GTF), 1021-1022 Guinea-pig maximization test (GPMT), 300 Gulf War, 987,1044,1069
H Haber's rule, 3, 7, 25-27, 41 Habit alteration, 190 Halogenated hydrocarbons, 596, 597 Halophenols, 1232 Hand exposure, 430 Handbook of Pesticide Toxicology (HayeslLaws), 1095 Harmonic mean, 835 Hartley-Sielken model, 17 Hartung model, 6 HAs. See Heterocyclic amines Hayes factor, 20 Hazard quotient, 365 HCB. See Hexachlorobenzene HCF. See Health care facility HCH. See Hexachlorocyclohexane Head lice, 193 Health care facility, 605, 609 Heat dissipation, 206 Hellebore, 127 Hen test, 955-959 Henry's law, 646 Heptachlor, 771, 1135-1138 Herbal supplements, 823 Herbicides, 101, 327, 547, 590, 777 arylphenoxyproprionic acid, 104 bilanafos, 152-153 developmental toxicity, 385-389 glufosinate, 153-155 immune function, 777 list of, 385-389 miscellaneous structure, 327-328 oxphos, 1179 phenoxy, 104 reproductive toxicity, 385-389 triazine, 105 turfgrass. See Turfgrass weeds and, 215 See also specific substances HERP. See Human exposure/rodent potency Heteroatom oxygenation, 538 Heteroatom release, 538 Heterocyclic amines (HAs), 812 Hexachlorobenzene (HCB), 9, 401, 755-756 Hexachlorocyclohexane (HCH), 98, 342-344, 397, 771 Hexachloronorbornenes, 1141-1146 HHS. See Department of Health and Human Services High-resolution mass spectrometry (HRMS), 674 Historical cohort studies, 635 Home and Garden Pesticide Use Survey, 205
1902
Index
Homes. See Residential exposure Hong Kong outbreaks, 626 HormoIigosis, 33 Hormones, 103,727-746 Horses, 265 Hospital discharge survey, 622-623 House dust, 791 HRMS. See High-resolution mass spectrometry Human exposure/rodent potency (HERP), 806-816,822 Human studies arthopods, 187-188 biomonitoring, 758-761 exposure, 222-227, 800-802. See also Exposure health studies, 787-788, 845-858 HERP studies, 806-816, 822 natural defenses, 802 susceptibility, 58-60 See also Dosage; specific effects, substances Humidity, 524-525 Hydramethylnon, 317, 1208-1209 Hydrolases, 540 Hydrolysis, 922-925, 970-972 Hydrophobicity, 649-670 Hyperactivity, 590 Hypersensitivity, 42-43 Hypothalamic-pituitary effects, 737-740
IARC. See International Agency for Research on Cancer IAs. See Immunoassay techniques IDLH. See Immediately dangerous to life and health IGRs. See Insect growth factors Illness, surveillance methods, 603-638 Imazalil, 321 Imazethapyr, 326 Imidacloprid, 317, 345-346, 406, 1I23-1130 ImidazoIinones, 1641-1651 Imiprothrin, 1278 Immediately dangerous to life and health levels (IDLH), 1I03 Immunoassay techniques (IAs), 676-677 Immunotoxicity, 769-782 In vitro methodology, 907 Incubation period, 184 Independent laboratory validation, 672 Independent units, statistical, 29 Indirect action, 377 Indoor sources. See Residential exposure Induction, 51O-51I, 551 Industrial Hygiene Technical Manual (OS HA), 224 Infants, 685, 785-786 Inhalation exposure, 425-426, 430-431, 436, 460, 466 Inhibition, 509,550-551 Inorganic pesticides, 1357-1428 Insect behavior, 103-104 Insect growth factors (IGRs), 191 Insect growth regulators, 103,400-401 Insecticides, 101, 1I0-144, 207, 313, 390-401, 879-881,919-927 anabasine, 121 arthropods, 192 azadirachtin, 130-134 biological, 234-235, 316-317
botanical, 1I0-1I6 categorizing, 191-193 crack and crevice treatments, 247 developmental toxicity, 391-395 growth regulators, 103, 191,400-401 hormones, 103 list of, 391-395 microbial, 130-138 mosquito management, 194-196 neurophysiological effects, 335-346 nicotine, 1I6-121 pyrethrins, II 0-1I6 repellants, 313, 316-317 reproductive toxicity, 391-395 rotenone, 121-122 ryania, 128-130 sabadilla alkaloids, 126-128 semiochemicals, 140-144 soaps, 317 spinosad, 138-140 spot treatments, 246 thuringiensis endotoxins, 130-138 urban structural pest control, 243-249 See also specific types, substances Integrated pest management (IPM), 188-197,214, 228,243,275-282 Interactive effects, 48-52 Intermediate syndrome, 888, 1062-1064 International Agency for Research on Cancer (IARC),747 International Society of Exposure Analysis (ISEA), 438-439 International surveys, 624-627 Internet, 638 Intolerable Risk: Pesticides in our Children's Food (NRDC),875 Intradermal methods, 300 Intravenous bolus dose, 571-573, 576 Intuitive technology, 851-856 Ionizable pesticides, 651 Ionizing radiation, 69 Ioxynil, 1232, 1235-1237 IPM. See Integrated pest management Iprodione, 321 Irreversible inhibition, 549 Irrigation, 230-231 ISEA, 438-439. See International Society of Exposure Analysis Isobenzan, 1135-1138 Isodrin, 1131-1134, 1135 Isoelectrophilic windows, 666 Isolation effects, 64-65, 960-961 Isoproturon, 390, 1525-1526 Isothiazolins, 300 Isoxaben, 328 Ivermectin, 881, 1I57, 1I63-1I64
J Jasmolin, 111
K Kadethrin, 1278 Kagan method, 19 Kasugamycin, 146-147 Kefauver-Harris Act, 378 Kepone, 731-733, 1139-1I41 Ketoconazole, 404, 736
Keystone species, 358 Kidneys, 541-542,583-584,785 Kinetics, 51, 77-80 Kresoxim-methyl, 1202-1203
L LADD. See Lifetime average daily dosage Lay judgments, 851-856 LC values, 28 LCGU. See Local cerebral glucose utilization LD values, 9, 10-14 90-dose, 14, 15 confidence limits, 12 defined,286-287 ED values and, 12 Kagan method, 19 one-dose, 10-14 problems with, 10 procedures for, 13 shape of curve, 11 Leaching, 229-230 Lead, 1383-1385 Leaving group, 913 Lethality studies, 254-255, 691 Leukemia, 601 Levamisole, 791 Leydig cell tumors, 739-740 LH. See Luteinizing hormone Lidocaine, 1298 LifeLine, 438 Lifetime average daily dosage (LADD), 716 Light, effects of, 68-69 Limit of detection (LOD), 362-363, 671 Lindane, 316,397,598, 734, 771,879,1131-1134 Linearized multistage model, 810 Linuron, 390, 734 Lipid peroxidation hypothesis, 1571 Lipoid theory, 664 Lipophilic nitrogen heterocycles, 1I87-1I89 Liquid chromatographic method, 651-652, 674-676 Lithium, 1024, 1025 Lithium perfiuorooctanesulfonate, 1246 Liver glutathione (GSH), 37-38 Liver microsomal enzymes, 203-205, 537-538, 544-546 Liver tumors, 16 Livestock protection collars, 260 Loading exposure, 459 LOAEC. See Lowest observed adverse effect concentration Local cerebral glucose utilization (LGCU), 1027, 1030 LOD. See Limit of detection LOEL. See Lowest-observed-effect level Log P values, 650-658 Lognormal response, 39 Logprobit model, 16-17,34-35 Logtime-logdosage curve, 23-27 Low-volume test (LVT), 290 Lowest observed adverse effect concentration (LOAEC), 359 Lowest-observed-effect level (LOEL), 10,31 LT values, 22-23 Luke method, 672 Lung,541,1577
Index
Luteinizing hormone (LH), 721 LVET. See Low-volume test
M mAChRs. See Muscarinic receptors Mackey's postulate, 663 Macrofiora, 205 MADs. See Mosquito abatement districts Malaria, 1305 Malathion, 50, 399, 773, 778, 1789 Males, 787 Malformations, 382 Mammalian toxicity, 859-872, 1743-1757 Mancozeh, 320 Maneb,404 Margin of exposure (MOE), 5, 451-452 Margin of safety, 707, 721-722 MATe. See Maximum allowable toxic concentration Material safety data sheet (MSDS), 286 Maximum allowable toxic concentration (MATC), 359 Maximum residue limits (MRLs), 671 Maximum tolerated dose (MTD), 33, 295 Mazzotti reaction, 1163-I 164 McAlister, Donald (1879), 39 MDAe. See Multiple-dose activated charcoal MDI. See Methylene bisphenyl isocyanate MDP. See Methylenedioxyphenyl Mean residence time (MRT), 569 MECe. See Micellar electrokinetic capillary chromatography Mechanisms, of toxicity, 1289-1303 Median lethal dose. See LD values Megamouse experiment, 17 Mental retardation, 384 Mepronil, 1198 Mercury compounds, 1369-1380 MEST. See Mouse ear-swelling test Metabolism, 509-511,526,785, 1027-1028 biomarkers,585-586 chemical factors, 509, 542-551 defined,531 inhibition, 509 metabolites, 46-47, 105-106,542,585-586 organophosphorus and, 919-927, 947, 1051 pathways, 103,537-538 pesticides and, 100--101,271,531-554 storage, 44-45 toxicity, 507-511 Metam-sodium, 323,405, 1867, 1867-1879 Methamidophos poisoning, 626 Methomyl, 1097 Methoprene, 400 Methoxychlor, 316, 396-397, 733-734, 1337 Methyl bromide, 323,409-410,754-755, 1837-1847, 1867 Methyl carbamate, 776 Methyl isocyanate, 597 Methyl parathion, 398, 633, 772-773, 783 Methyl thiophanate, 405 Methylene bisphenyl isocyanate, 911 Methylene bis(thiocyanate), 321-322 Methylenedioxyphenyl (MDP), 547, 551, 1132 Metolachlor, 324, 1551-1552 Metominostrobin, 1203-1204 Metosulam, 1661-1663
Mevinphos, 695--697 Mexacarbate, 1096 Mexico, 626 Micellar electrokinetic capillary chromatography (MECC),677 Michaelis-Menten effect, 18 Michel method, 974 Microbial antagonists, 233-234 Microbial insecticides, 130--138 Microbial metabolism, 99 Microbial pest control agents (MPCAs), 859-872 Microfiora, 205 Micronucleus assay, 751 Microsomal enzyme activity, 543-547 Mildiomycin, 147-148 Milk,585 Minnesota Children's Pesticide Exposure Study (MNCPES), 898-900 Minnesota Multiphasic Personality Inventory (MMPI), 1102 Miosis, 593 Mirex, 9, 772, 1139-1141 Mites, 184 Mitochondria, 1169-1261, 1572-1573 Mitosis, 1673-1694 Mitotane, 1332 MMPI. See Minnesota Multiphasic Personality Inventory MNCPES. See Minnesota Children's Pesticide Exposure Study MOE. See Margin of exposure Molecular techniques, 191,236,961-963, 969-970 Molinate, 389,700--702,736 Monooxygenases,534,922 Monte Carlo models, 446, 451, 480 MOS. See Margin of safety Mosquito abatement districts (MADs), 189 Mosquitoes, 186-189, 194-196 Motor vehicle accidents, 598 Mouse ear-swelling test (MEST), 289 MPCAS. See Microbial pest control agents MRLs. See Maximum residue limits MRM. See Multiple-reaction mode MRT. See Mean residence time MS/MS method, 676 MSDS. See Material safety data sheet MTD. See Maximum tolerated dose MUDDLES complex, 593 Multicompartment models, 575-576 Multi enzyme systems, 5 I I Multiple-dose activated charcoal, 591 Multiple molecular forms, 969-970 Multiple-reaction mode (MRM), 676 Multiresidue methods, 672--673 Multistage model, 810 Muscarinic receptors (mAChRs), 1021-1022, 1030-1033, 1048 Muscle cells, 1033-1034 Mushroom house, 457-475 Mutagenicity, 42, 290--291, 748, 804-806, 1100--1101, 1318-1319 Myclobutanil, 322 Myopathy, 1054
1903
N NADH. See Nicotinamide adenine dinucleotide NAIN. See National Antimicrobial Information Network Narcosis, 664 Nasal tissues, 541 NASS. See National Agricultural Statistics Service National Agricultural Statistics Service (NASS), 280,638 National Agricultural Worker Survey (NAWS), 623,638 National Antimicrobial Information Network (NAIN),628 National Center for Health Statistics (NCHS), 620, 622-623 National Food Consumption Survey, 818 National Health and Nutrition Examination Survey (NHANES), 636, 897-898 National Human Activity Pattern Survey (NHAPS), 896 National Institute for Occupational Safety and Health (NIOSH), 613 National Pesticide Residue Database (NPRD), 821 National Pesticide Telecommunications Network (NPTN), 628, 638 National Public Health Surveillance System (NPHSS), 629, 632 National Research Council (NRC), 816 National Toxicology Program (NTP), 747 Natural chemicals, 801 Natural defenses, 802 Natural organic products, 233 Natural pesticides, 801,810--811 Natural predators, 235 Natural products, 106, 109-179 NAWS. See National Agricultural Worker Survey NCHS. See National Center for Health Statistics Necrosis, 1177 Nemacide,737 Nematodes, 207-211, 235 Neonicotinoids, 1123-1130 Nervous system, 1317-1318 Neuromuscular endplate, 1033-1034 Neuropathy target esterase (NTE), 292, 596, 878, 953-965, 1003, 1065-1066 Neurophysiological effects, 335-346 Neurotoxic esterase. See Neuropathy target esterase Neurotoxicity, 40-41, 270, 600, 878, 1179-1181, 1297-1298 Neurotransmitter levels, 1028-1029 NHANES. See National Health and Nutrition Examination Survey NHAPS. See National Human Activity Pattern Survey Niacin, 1577 Niclosamide, 1237-1239 Nicotinamide adenine dinucleotide (NADH), 920, 1171-1172 Nicotine, 116-121 Nicotinic receptors, 1013, 1023-1024, 1048, 1123-1124 NIOSH. See National Institute for Occupational Safety and Health Nitroaniline compounds, 324-325 Nitrofen, 389-390,740 Nitrosamines, 812
1904
Index
No-observed-adverse-effect level (NOAEL), 31, 359, 1109 No-observed-effect level (NOEL), 10, 31,193 NOAEL. See No-observed-adverse-effect level NOEL. See No-observed-effect level Nomenclature, 913-917 Non-Occupational Exposure Assessment Study (NOPES), 898 Noncompartmental models, 569-571 Noncompetitive inhibitors, 549 Noninsecticidal methods, 190-191 Nonlethal managment, 253-254 Nonoccupational Pesticide Exposure Study, 437 Nonthreshold models, 721 Nonthreshold toxicant, 286 NOPES. See Non-Occupational Exposure Assessment Study Noradrenaline, 1294 Norbormide, 1829-1830 Normal curves, 12, 18 NPHSS. See National Public Health Surveillance System NPRD. See National Pesticide Residue Database NPTN. See National Pesticide Telecommunications Network NRe. See National Research Council NTE. See Neuropathy target esterase NTP. See National Toxicology Program Nutrition, 33, 62, 64
o Occupational exposure, 27, 493-505,692,716, 720, 1068-1072. See also specific effects, substances Occupational Safety and Health Administration (OSHA),81O Ochratoxin, 813 Ocular irritation, 289-290 OECD. See Organization for Economic Cooperation and Development Office of Prevention, Pesticides, and Toxic Substances (OPPTS), 379 Olefinic suicide destruction, 538 Omethoate, 1065 Oncogeuicity. See Cancer One-compartment model, 2, 571-575 OP. See Organophosphate compounds Operator exposure data, 459-462, 500 OPIDN. See Organophosphate-induced delayed neuropathy OPP. See Ortho-phenylphenol OPPTS. See Office of Prevention, Pesticides, and Toxic Substances OPs. See Organophosphate compounds Oral exposure, 426 Ordram,777 Organic Foods Act of California, 375 Organic pollutants, 206 Organization for Economic Cooperation and Development (OECD), 360 Organochlorines, 315-316, 390-397, 575, 589, 596,770-772,879-880 Organogenesis, 291 Organometal pesticides, 1357-1428 Organophosphate compounds, 102,313,325-326, 397-399,876-879,913-927,943 AChEs and, 1013-1019
anticholinesterases, 1034-1035, 1043-1085 atropine, 1056 biomonitoring,941 BuChEsand,1073-1074 carbamates, 1091-1092 cardiac manifestations, 1048-1049 categories of, 1016-10 17 central nervous system and, 1049 chemistry of, 988-989, 1016-1018 cholinergic syndrome, 1043-1062 cyclodienes, 1132 delayed polyneuropathy, 1064-1067 diazepam, 1059 electrophysiology, 1051 encephalopathy and, 1052 esters, 772-775 impurities in, 773-774 intermediate syndrome, 1062-1064 long-term exposure, 1067-1075 metabolism, 912-927, 947 metabolites, 1051 modulation of, 1030 myopathy and, 1054 occupational exposure, 1068, 1072 OPIDN. See Organophosphate-induced delayed neuropathy oximes and, 1057 pancreatis and, 1053 Parkinsonism, 1068 pharmokinetics, 929-951 psychiatric disorders and, 1071 quantifying exposure, 941-942 respiratory failure and, 1048 suicide and, 1071 synergism, 1132 toxic actions of, 1013-1041 treatment of, 1055-1060 Organophosphate-induced delayed neuropathy (OPIDN), 292, 594, 987, 987-1012 aging process, 999 animal models, 989, 996-999 calcium channels blockers, 1002-1003 chira1 compounds, 959 clinical manifestations, 989 factors influencing the development of, 1000-1003 glucocorticoids, 1000-1003 hen test, 955-959 nervous system and, 994 neuropathologic studies of, 990-992 NTE and, 963, 1003-1004, 1065-1066 pathogenesis,999-1001 potential for, 954-964 testing for, 1004 Organotins, 1210-1224 Ortho-phenylphenol, 756-757 Oryzalin, 324 OSHA. See Occupational Safety and Health Administration Oxalic acid, 823 Oxidase inhibitors, 1529-1541 Oxidation hypothesis, 1571-1572 Oxidations, 920-922 Oxidative group transfer, 538 Oxidative phosphorylation, 1169-1261 Oximes, 1057, 1063, 1088 Oxphos, 1174-1176, 1179
Oxycarboxin, 1198-1199 Oxychlordane, 1137 Oxygenation, 538 Oxythioquinox, 317
p P450-mediated reactions, 534, 547, 920, 922 PAHs. See Polycyclic aromatic hydrocarbons Pan American Health Organization, 626 Pancreatis, 1053 Paracelsus, 2, 5 Paraesthesia, 1289-1303 Parameter uncertainty, 452 Paraoxonase (PON), 789, 973, 1025, 1063 Paraquat, 324, 384, 1559-1603 Parasites, 56, 235 Parasporal protein cystal, 862 Parathion, 398,550,772-773,877,906,920, 936-938, 1015 PARe. See Pesticide Analytical and Response Center Parkinson's disease, 788, 789, 1068 Passive dosimetry, 428-431 Passive transport, 563-564 Patch testing, 299, 428-429 Pathway-exposure factor (PEF), 890 PBO. See Piperonyl butoxide PBPK. See Physiologically-based pharmacokinetic models PCBs. See Polychlorinated biphenyls PCNB. See Pentachloronitrobenzene PCP. See Pentachlorophenol PEF. See Pathway-exposure factor PEG. See Polyethylene glycol PELs. See Permitted exposure limits Pendimethalin,324-325 Pentachloroketone, 1135 Pentachloronitrobenzene (PCNB), 322 Pentachlorophenol (PCP), 405, 1232, 1481-1509 Percutaneous absorption, 516, 905-912 clothing and, 910 decontamination, 905-912 methodology, 905-907 parathion and, 906 regional variation, 907-909 See also Dermal exposure Perfluorooctanesulfonic acid, 1244-1247 Peripheral nervous system (PNS), 600 Permethrin, 315, 881, 1278-1279 Permitted exposure limits (PELs), 810, 1103 Personal measurement, 890 Personal protection, 196-197 Pesticide Analytical and Response Center (PARC), 614 Pesticide Handlers Exposure Database, 328, 493, 495,500-501,610,693 Pesticides active ingredients, 100 animals as sentinels of, 263-265, 271-272 chemistry of, 95-107, 521-522 classification of, 97-99,191, 608-611 cost of illness, 634 defined, 632, 691 developmental toxicology, 375-410 direct-acting, 377 disposal of containers, 268 dissipation of, 471-473
Index
domestic animals, 263-265, 270-272 economic markets for, 96 environment and, 99-101, 353-371 enzymes, 44 food and, 808-809, 816--822 fungicides. See Fungicides indirect-acting, 377 insect behavior, 103-104 insecticides. See Insecticides large-scale use, 268 lethal agents, 254-255 list of, 407-408 measuring effects of, 360-361 metabolism, 100-106,271,507-511,531-554 miscellaneous, 407-408 natural products, 106, 109-179 nomenclature for, 97 parasites and, 56 poisoning, 589-601 regulatory issues, 99-101, 378-383, 681-682 repellents, 254-255 reproductive toxicology, 375-410 residues, 100-101,808-809,816--822 risk assessment, 353-354, 364-370 rodenticides. See Rodenticides root zone model, 229 skin effects, 299-330 stereochemistry, 97-99 storage of, 268 training for use, 268 treatment of poisoning in animals, 272 types of, 589-590 vector management, 191-197 vertebrate pest control, 253-254 veterinary medicine, 263-272 See also specific types, effects, standards Pesticides in the Diets of Infants and Children (NAS), 793, 875, 900, 944 PESTOTEST software, 626 pH methods, 974-975 Pharmacokinetics, 34,563-581,692,929-951 absorption. See Absorption biomarker measurements, 896 models, 577-581, 930-935, 943, 947 quantifying exposure, 941-942 risk assessment, 1115-1119 Phase IIlI reactions, 539-540 PHED. See Pesticide Handlers Exposure Database Phenobarbitals, 38 Phenolic compounds, 312, 1232-1239 Phenothrin, 315, 1281-1282 Phenoxy herbicides, 104,325-326,777, 1623-1638 Phenoxyacid herbicides, 547, 595 Phenylurea herbicides, 1521-1527 Pheromones, 103,141 PHI. See Pseudohermaphrodism index Philanthotoxins (PhTXs), 1152 Phosphine, 1861-1866 Phosphorodithioates, 929 Phosphorus, 1399-1403 Photoperiodicity, 69, 75 Photosensitization, 69-70 Phthalimido compounds, 318-319 PhTXs. See Philanthotoxins Physical suppression, 279 Physician reporting, 632
Physiochemical parameters, 521-522, 649-670 Physiological factors, 551-552 Physiologically-based pharmacokinetic models (PBPK), 577-581, 933-935, 943, 947 Phytochemicals, 197 Picoxystrobin, 1203 Piperonyl butoxide (PBO), 1461-1480 Pirenzepine, 1022 Pittsburgh Occupational Exposure Test (POET), 600 Plasma clearance, 570,572-573 Plasma membrane, 564 Plasma proteins, 78-79, 568 Plondrel, 319 PNS. See Peripheral nervous system POEM. See Predictive Operator Exposure Database POET. See Pittsburgh Occupational Exposure Test Point-of-contact measurement, 890 Poisoning, 595-599 accidental, 267-268 acute, 590-592, 592-595 diagnosis, 589-601 malicious, 267-268 pesticides, 589-601 surveillance data, 631-634 syndromes, 1289-1303 systemic, 1289-1303 therapy, 589-601, 1289-1303 See also specific effects, substances Political influences, 847-851 Polychlorinated biphenyls (peBs), 9, 813 Polychlorocycloalkane, 1131-1156 Polycyclic aromatic hydrocarbons, 36, 791 Polyethylene glycol (PEG), 592 Polymerized pinene, 328 Polymorphisms, 511, 552, 788-789, 944-948 Polyneuropathy. See Organophosphate-induced delayed polyneuropathy Polyurethane foam (PUF), 895 POMS. See Profile of Mood States PON. See Paraoxonase Porphyrin, 70-74,599,1537 Potentiation, 50-51, 509-510, 550, 791-792 Potentiometric titration method, 651 Poverty, and pesticides, 791 Pralidoxime, 594, 595, 980, 1059 Prallethrin, 1282 Predictive Operator Exposure Database (POEM), 493,500-501 Pregnancy, 61 Pressure, effects of, 68 Primary compounds, 47 Probabilistic analysis, 366--370,446,479,480-482 Procymidone, 734-735 Profile of Mood States (POMS), 1102 Prometryn, 327 Promutagens, 290 Propachlor, 1552-1554 Propanil, 777 Propargite, 317 Propoxu~697-698, 755,1088,1097,1098 Prospective cohort studies, 635 Prostaglandin synthetase, 540 Protection collars, 260 Protective clothing, 474, 475 Protein binding, 565
1905
Protein, dietary, 63-64, 565, 791-793 Protoporphyrinogen oxidase, 1529-1541 Protox, 1529 PRZM. See Pesticide root zone model Pseudohermaphrodism index (PHI), 735 Psychiatric disorders, 1071 PubMed,638 PUP. See Polyurethane foam Pyraclostrobin, 1204 Pyrethrins, 314,589,1282-1283 Pyrethroids, 102-103,314,315,400,735-736,777 anticholinesterase activity, 1293 calcineurin and, 342 calcium channels, 341 chemistry of, 1263-1288 chloride channels, 341 enzymes and, 342 GABA receptors, 341 half-life of, 1294 insecticides, 880-881, 1289-1303 poisoning, 1296--1299 resistance, 342 sodium channels, 338, 342, 1292 temperature dependence, 336--339 toxicity, 340 two classes, 1295 Pyridaben, 1190 Pyridine derivatives, 326 Pyridostigmine, 980, 1044, 1068 Pyrimidfen, 1191-1192 Pyriminil, 1802-1805 Pyrophosphoric acid, 1017
Q QSAR. See Quantitative structure-activity relationship Quantifying exposure, 674-676, 941-942 Quantitative structure-activity relationship (QSAR), 649, 661 Quantum-chemical descriptors, 666 Quinone, 1174
R Rabbit hemorrhagic disease virus, 867-868 RADAR. See Risk Assessment Duration and Recovery Radiation, 17-18,68-69 Radioactive decay, 721, 905-906 Radiometric techniques, 651,974 Radiovalidation, 672 Randomization methods, 30 Reabsorption, 583 REAM. See Residential Exposure Assessment Model Receptor characterization, 356 Recommended exposure limits (RELs), 1103 Reconstructive analysis, 890 Red squill, 159-162 Redox cycling, 1175-1176 Reduced-risk pesticides, 278-279 Reentry exposure, 462-473 Registration, of products, 686-688 Regulating Pesticides in Food, 816 Regulatory policies, 627-628, 874-875 children's health, 874-875 current, 686-689 history of, 378-379, 681-682
1906
Index
pesticides, 378-383 process, 68\-690 regulation, 874--875 See also specific legislation, substances RELs. See Recommended exposure limits Renal functions. See Kidneys Repellents, 196-197,254--256,316-317 Reproducibility, of results, 39-40 Reproductive studies, 295, 375-410, 714 Residential exposure, 435-439 children, 887-904 indoor surfaces, 894 microenvironments, 894--896 sources of, 436 symptoms of, 611 Residential Exposure Assessment Model (REAM), 438 Resistance, defined, 552-554 Resmethrin, 315,1283-1284 Respirator, 431 Respiratory systems, 458-462, 566-567, 584, 1034--1035, 1048 Retrospective studies, 635-636 RHDY. See Rabbit hemorrhagic disease virus Rice blast disease, 146 Ricin, 162-164 Ripple effects, 358 Risk assessment, 366-370 acute exposure, 691-706 aldicarb, 1115-1119 analysis of, 359, 598 animals and, 942-947 characterization process, 693, 718-720 ChEs and, 980--981 chronic exposure and, 707-726 communication of, 371 dermal exposure, 1119-1121 EPA guidelines for, 695 exposure assessments, 820--821 hazard quotient, 365 pesticides and, 353-354 pharmacokinetics and, 1115-1119 public perceptions, 845-858 RADAR study, 362 reporting requirements and, 627-628 risk management and, 473-474, 681-690 science-based approaches, 1107-1121 scoring systems, 364--365 toxicological data, 821 uncertainty and, 359, 370--371 See also specific substances, studies Risk Assessment Duration and Recovery (RADAR), 362 Rodenticides, 155-164,293,406, 1793-1836 baits, 257-259 red squill, 159-162 ricin, 162-164 salmonella bacteria, 164 scilliroside, 159-162 strychnine, 155-159 synthetic organic, 1827-1830 Rotenone, 121-122, 1181-1185, 1186 Rumen, effect of, 271 Runoff, 229-230 Rural environments, 251-262 Ryania, 128-130
S Sabadilla alkaloids, 126-128 Saccharin, 813 Safe Drinking Water Act, 741 Safe levels, defined, 684 Safrole, 551, 813 Salivation, as symptom, 593 Salmonella bacteria, 164 Sampling systems, 29, 895 Scabies mite, 193 Scenario-based assessments, 452, 890--892, 894--896 Science-based approaches, 1107-1121 Scilliroside, 159-162 Screening models, 438, 446 SDWA. See Safe Drinking Water Act Amendments Seasonal differences, 76 Seed treatments, 267-269 Seizurogenic effects, 1029-1030 Selective inhibitors, 969 Selenium, 1404--1407, 1576-1577 Semiochemicals, 140--144 Sensitivity age-related differences, 785-786, 873-882 genetic predisposition, 788-789 subpopulations, 783-798, 944--948 surveillance system, 632--633 SENSOR. See Sentinel Event Notification for Occupational Risk Sentinel Event Notification for Occupational Risk (SENSOR), 612, 613 Separation, 674--676 Serine esterases, 953 Sesoxane, 1132-1134 Sex differences, 61,584--585,786-788,848 Shake-flask procedures, 650 SHEDS. See Stochastic Human Exposure and Dose Simulation model Sheep, 264 Short -term exposure, defined, 888 Signs, assessment of, 598-600 Silent Spring (Carson), 99, 845 Simazine, 327,714 Single-species tests, 359 Site of exposure, 610 Skin. See Dermal exposure SLUD symptoms, 593 SNRs. See Supernumerary ribs Social influences, 65-66, 847-851 Sodium channels, 1289-1292 gating kinetics, 335-336 modulation, 335-341 pyrethroids, 336-338, 342, 1290--1292 TTX-sensitive, 337-338 Sodium chlorate, 1411-1413 Sodium diethylthiocarbamate, 1098 Sodium f1uoroacetate, 1794--1799 Sodium o-phenylphenate (SOPP), 756 Sodium selenate, 1404--1407 Soil, 525-526 acidity of, 1781 adsorption, 659 behavior in, 658 invertebrates, 227-228 restoration, 206 runoff, 363 termites and, 247
treatments, 247 wind erosion, 205 Solid-phase extraction (SPE), 674 SOPP. See Sodium o-phenylphenate South Carolina discharge surveys, 623 Soxhlet extraction, 673 Spacecraft, 68 SPE. See Solid-phase extraction Special review (EPA), 688--689 Specialized transport, 564--565 Species differences, 54--80,270--271,381,509, 551 Spices, 823 Spinosad, 138-140 Spot treatments, 246 Spray drift model, 363-364 Staircase method, 287 Starvation, 62--63 State-based surveillance, 613-617 Statistical methods, 28-40 Stereochemistry, 97-99 Steroid synthesis, 728-737 Stochastic Human Exposure and Dose Simulation model (SHEDS), 438 Storage procedures, 44-45, 268, 569 Strain differences, 54--79 Strawberry harvester, 483 Streptomycin, 149-151 Stressor characterization, 356 Stripping method, 906-907 Strobilurin analogs, 1199-1200 Structure-activity relationships, 658, 959 Strychnine, 155-159 Study designs, 635 Subchronic studies, 294 Submarines, 68 Substituent effects, 655--666 Substituted ureas, 1802-1805 Substrate preferences, 969 Suburban environment, 205 Succinate, 1172 Sufficient challenge, 33 Suicide, 606, 1015, 1065, 1071 Sulfluramid, 1244 Sulfosate, 325 Sulfoxidation, 922 Sulfur,920, 1403-1404, 1781-1791 Sui fur dioxide, 1784--1785 Sulfuryl fluoride, 1410--141 1,1881-1896 Sulphur, 322 Sun-Johnson hypothesis, 1132-1134 Supercritical fluid extraction, 673 Supernumerary ribs (SNRs), 382 Superoxide dismutase, 1576 Surface water, 217-218 Surveillance systems, 603--638 definitions for, 603, 632 evaluating, 628--629 limitations of, 631--634 uses for data, 633--634 Sweet clover poisoning, 1809 Swimming, 482 Swine, 264--265 Symptoms, 598, 608. See also special effects, substances Synergism, 509-510, 550, 791, 1132-1134
Index
Synthetic chemicals, 800--802 System characterization, 356 Systemic pesticides, 269 Systemic poisoning syndromes, 1289-1303
T t-Butyl trioxabicyc1ooctanes, 1144 t-Butylbicyc1ophosphorothionate, 1143 Tandem method, 676 TBOs. See t-Butyl trioxabicyc1ooctanes TBPS. See t-Butylbicyc1ophosphorothionate TCMTB. See Thiocyanomethylthiobenzothiazole TCP. See Trichloropyridinol TDE. See Tetra-chlorodiphenylethane TEAM. See Total Exposure Assessment Methodology Tebufenpyrad, 1192-1193 Telephone resources, 638 Telodrin, 597 Temperature dependence, 67, 336--339, 524, 1034-1035 TEPP. See Tetraethyl pyrophosphate Teratogenicity, 41,291-292,382 Termites, 247 Terrestial systems, 360 TESS. See Toxic Exposure Surveillance System Test-Mate models, 978 Testing animals, 57 doses, 381-382 species, 381 Testing requirements, 299 Testing strategies toxicology, 285 Tetra-chlorodiphenylethane (TDE), 1332-1335 Tetracaine, 1298 Tetrachlorodibenzo-p-dioxin (TCDD), 3, 9, 48, 384,597, 813 Tetraethyl pyrophosphate (TEPP), 1014 Tetraethylthiuram, 1098 Tetramethrin, 315, 337,1284--1285 Tetrodotoxin (TTX), 336 Thalidomide, 378 Thallium, 1380-1383 Thallium sulfate, 1380-1383 Thatch, 204 Theoretical profile shape method (TPS), 223 Therapeutic drugs, 34 Thermodynamics, 2, 6--7 Thiabendazole, 322 Thiamethoxam, 1123 Thidiazuron, 327 Thifluzamide, 1199 Thio substrates, 975 Thiocarbamates, 320, 326 Thiocholine assays, 977 Thiocyanomethylthiobenzothiazole (TCMTB), 322 Thiodan, 1139 Thiodicarb, 399 Thioimidodicarbonic diamide (DTB), 881 Thiol oxidase reaction, 976 Thiophanate methyl, 319-320 Thioureas, 1805-1809 Thiram, 400 Threshold levels, 32-33, 286, 684 Thuringiensis bacteria, 130-138, 862-863 Thyroid tumors, 739, 1556
Ticks, 184 Tiered approach, 493-494 Tin, 1385-1389 Tissue level, 28, 79 Tolerance, 552-553,686 Tolylfluanid, 1733-1734 Tordon herbicide, 1180 Total Exposure Assessment Methodology (TEAM), 435-437 TOTP. See Tri-ortho-cresyl phosphate Toxaphene, 772,1139-1141 Toxic Exposure Surveillance System (TESS), 604, 611-612 Toxic Substances Control Act (TSCA), 379 Toxicant ejector device, 260-261 Toxicity adsorption, 2, 431 age and, 873-886 allergy and, 42-43 carcinogenesis. See Carcinogenesis chronic studies, 294 cyanazine,72I-722 database for, 28-29, 691 defined, I delayed, 23-24 diet, 63 dosage-response relationships. See Dosage-response relationships dose. See Dose ED 50. See ED values factors influencing, 1-82, 45-82 formulation, 52 four fundamental variables, 46 humidity and, 76 hypersensitivity, 42-43 kinds of, 7-9, 40-45 LD values. See LD values mechanisms of, 1289-1303 metabolism. See Metabolism mutagenesis. See Mutagenesis neurotoxicity. See Neurotoxicity pharmacology. See Pharmacokinetics pyrethroids. See Pyrethroids seasonal differences, 76 statistical methods, 28-40 teratogenicity, 41,291-292,382 testing strategies, 285-296 theory of, 36 time and, 1,4--6,23,46,82 See also specific effects, substances, tests Toynbec, A., 33-34 TPS. See Theoretical profile shape method Tracking powders, 259-260 Trajectory simulation model, 223 Tralomethtin, 1285-1286 Transfer factors, 471-473 Transgenics, molecular techniques, 191 Transplacental absorption, 784 Transport mechanisms, 563 Tri-ortho-cresyl phosphate (TOTP), 988 Triadimefon, 322 Triazines, 105,312,326--327,390,536,707-726, 714, 1511-1519 Triazoles, 737 Triazolopyrimidine herbicides, 1653-1665 Tribufos, 325-326
1907
Tributyltins, 737, 1210-1214 Trichlorfon, 56, 399 Trichloropyridinol (TCP), 939 Tric1opyr, 326 Tricyclohexy1, 1214 Tridiphane, 550 Triflox ystrobin, 1205 Trifluralin, 325 Triorganotins, 1214 Triorthocresyl phosphate, 50 Trioxabicyclooctanes, 1143-1145 Triphenyltins, 1211-1212, 1214 TSCA. See Toxic Substances Control Act TTX. See Tetrodotoxin Tubular secretion, 583-584 Turfgrass, 205 allelopathy, 235-236 alternative pest management, 232 atmospheric environment, 204 biological control, 232-235 biotic environment, 205 breeding programs, 236 diseases, 208, 233 edaphology, 204--205 exposure issues, 222-228 fungicides, 209-210 genetic engineering, 237 groundwater recharge, 205-206 herbicides, 215 leaves, 203 mechanical practices, 235 molecular genetics, 236 natural organic products, 233 natural predators, 235 nematodes, 235 parasites, 235 pesticide issues, 207-232 roots, 203-204 soil restoration, 206 stems, 203 Typhus, control of, 1305
u Ubiquinol, 1172 Ultra-Iow-volume (ULV) techniques, 195-196 Ultraviolet radiation, 69 ULY. See Ultra-Iow-volume techniques Uncertainty analysis, 359, 370-371,452 Uncoupling, 1226, 1232-1239 Up-and-down method, 287 Urban environments, 243-249, 251-262 Urea, 327,390, 1802-1805 Urinary excretion, 80 U.S. Department of Agriculture, 638,681,718, 818 Use categories, 300-328
v Vaccines, 191 Vadose zone flow model, 229 Valdez Air Health Study, 438 Validamycin, 148-149 Validation algorithms, 454 Vector-borne diseases, 181-187 Vector management, 188-197 Venoms,182 Vertebrate pests, 251-262 Veterinary medicine, 263-272, 408-409 Video imaging, 429 Vikane gas fumigant, 1881 Vinclozolin, 322-323,406,734-735 Viruses, 867-868 Visible radiation, 69 Vital status statistics, 620-621 Vitamin C, 1577 Vitamin D, 1823-1827
Vitamin E, 340, 1576 Vitamin K, 1809-1823 Volatilization, 222, 230
W WAIS. See Wechsler Adult Intelligence Scale Warfare agents, 1043-1045 Warfarin, 791-792, 1810-1816, 1825 Warren relationship, 26 Water, 205-206, 217-222, 567, 723 Web sites, 689 Wechsler Adult Intelligence Scale (WAIS), 1102 Weeds, 211-214, 215 White-male effect, 849 WHO. See World Health Organization WHO Pesticide Evaluation Scheme (WHOPES), 190 Whole body method, 429-430 WHOPES. See WHO Pesticide Evaluation Scheme Wildlife, 206, 227-228
Wingspread Conference, 727,728 Wipe sampling techniques, 894 Worker exposure, 425-432 Worker protection standard (EPA), 628 Worker's compensation, 598 World Health Organization (WHO), 190,624 World Wildlife Fund, 727
X Xanthine oxidase inhibitors, 1576 Xenobiotics, 79,531,539
z Zinc, 1363-1367, 1411 Zinc chloride, 1364-1365 Zinc hexafluorosilicate, 1411 Zinc phosphide, 1365-1367 Zineb, 320, 400 Ziram,320